Oceanography and Marine Biology: An Annual Review, Volume 45 (Oceanography and Marine Biology)

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Author: R. N. Gibson | R. J. A. Atkinson | J. D. M. Gordon

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Boca Raton London New York

CRC Press is an imprint of the Taylor & Francis Group, an informa business

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International Standard Serial Number: 0078-3218

CRC Press Taylor & Francis Group 6000 Broken Sound Parkway NW, Suite 300 Boca Raton, FL 33487-2742 © 2007 by R.N. Gibson, R.J.A. Atkinson and J.D.M. Gordon CRC Press is an imprint of Taylor & Francis Group, an Informa business No claim to original U.S. Government works Printed in the United States of America on acid-free paper 10 9 8 7 6 5 4 3 2 1 International Standard Book Number-13: 978-1-4200-5093-6 (Hardcover) This book contains information obtained from authentic and highly regarded sources. Reprinted material is quoted with permission, and sources are indicated. A wide variety of references are listed. Reasonable efforts have been made to publish reliable data and information, but the author and the publisher cannot assume responsibility for the validity of all materials or for the consequences of their use. No part of this book may be reprinted, reproduced, transmitted, or utilized in any form by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying, microfilming, and recording, or in any information storage or retrieval system, without written permission from the publishers. For permission to photocopy or use material electronically from this work, please access www.copyright.com (http:// www.copyright.com/) or contact the Copyright Clearance Center, Inc. (CCC) 222 Rosewood Drive, Danvers, MA 01923, 978-750-8400. CCC is a not-for-profit organization that provides licenses and registration for a variety of users. For organizations that have been granted a photocopy license by the CCC, a separate system of payment has been arranged. Trademark Notice: Product or corporate names may be trademarks or registered trademarks, and are used only for identification and explanation without intent to infringe. Visit the Taylor & Francis Web site at http://www.taylorandfrancis.com and the CRC Press Web site at http://www.crcpress.com

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Contents Preface Inherent optical properties of non-spherical marine-like particles — from theory to observation

vii

1

Wilhelmina R. Clavano, Emmanuel Boss & Lee Karp-Boss

Global ecology of the giant kelp Macrocystis: from ecotypes to ecosystems

39

Michael H. Graham, Julio A. Vásquez & Alejandro H. Buschmann

Habitat coupling by mid-latitude, subtidal, marine mysids: import-subsidised omnivores

89

Peter A. Jumars

Use of diversity estimations in the study of sedimentary benthic communities

139

Robert S. Carney

Coral reefs of the Andaman Sea — an integrated perspective

173

Barbara E. Brown

The Humboldt Current system of northern and central Chile — oceanographic processes, ecological interactions and socioeconomic feedback

195

Martin Thiel, Erasmo C. Macaya, Enzo Acuña, Wolf E. Arntz, Horacio Bastias, Katherina Brokordt, Patricio A. Camus, Juan Carlos Castilla, Leonardo R. Castro, Maritza Cortés, Clement P. Dumont, Ruben Escribano, Miriam Fernandez, Jhon A. Gajardo, Carlos F. Gaymer, Ivan Gomez, Andrés E. González, Humberto E. González, Pilar A. Haye, Juan-Enrique Illanes, Jose Luis Iriarte, Domingo A. Lancellotti, Guillermo Luna-Jorquera, Carolina Luxoro, Patricio H. Manriquez, Víctor Marín, Praxedes Muñoz, Sergio A. Navarrete, Eduardo Perez, Elie Poulin, Javier Sellanes, Hector Hito Sepúlveda, Wolfgang Stotz, Fadia Tala, Andrew Thomas, Cristian A. Vargas, Julio A. Vasquez & Alonso Vega

Loss, status and trends for coastal marine habitats of Europe

345

Laura Airoldi & Michael W. Beck

Climate change and Australian marine life

407

E.S. Poloczanska, R.C. Babcock, A. Butler, A.J. Hobday, O. Hoegh-Guldberg, T.J. Kunz, R. Matear, D. Milton, T.A. Okey & A.J. Richardson

Author Index

479

Systematic Index

535

Subject Index

541

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Preface The forty-fifth volume of this series contains eight reviews written by an international array of authors; as usual, the reviews range widely in subject and taxonomic and geographic coverage. The editors welcome suggestions from potential authors for topics they consider could form the basis of future appropriate contributions. Because an annual publication schedule necessarily places constraints on the timetable for submission, evaluation and acceptance of manuscripts, potential contributors are advised to make contact with the editors at an early stage of preparation. Contact details are listed on the title page of this volume. The editors gratefully acknowledge the willingness and speed with which authors complied with the editors’ suggestions, requests and questions and the efficiency of Taylor & Francis in ensuring the timely appearance of this volume.

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Oceanography and Marine Biology: An Annual Review, 2007, 45, 1-38 © R. N. Gibson, R. J. A. Atkinson, and J. D. M. Gordon, Editors Taylor & Francis

INHERENT OPTICAL PROPERTIES OF NON-SPHERICAL MARINE-LIKE PARTICLES — FROM THEORY TO OBSERVATION WILHELMINA R. CLAVANO1, EMMANUEL BOSS2 & LEE KARP-BOSS2 1School of Civil and Environmental Engineering, Cornell University, 453 Hollister Hall, Ithaca, New York 14853, U.S. E-mail: [email protected] 2School of Marine Sciences, University of Maine, 5706 Aubert Hall, Orono, Maine 04469, U.S. E-mail: [email protected], [email protected] Abstract In situ measurements of inherent optical properties (IOPs) of aquatic particles show great promise in studies of particle dynamics. Successful application of such methods requires an understanding of the optical properties of particles. Most models of IOPs of marine particles assume that particles are spheres, yet most of the particles that contribute significantly to the IOPs are nonspherical. Only a few studies have examined optical properties of non-spherical aquatic particles. The state-of-the-art knowledge regarding IOPs of non-spherical particles is reviewed here and exact and approximate solutions are applied to model IOPs of marine-like particles. A comparison of model results for monodispersions of randomly oriented spheroids to results obtained for equalvolume spheres shows a strong dependence of the biases in the IOPs on particle size and shape, with the greater deviation occurring for particles much larger than the wavelength. Similarly, biases in the IOPs of polydispersions of spheroids are greater, and can be higher than a factor of two, when populations of particles are enriched with large particles. These results suggest that shape plays a significant role in determining the IOPs of marine particles, encouraging further laboratory and modelling studies on the effects of particle shape on their optical properties.

Introduction Recent advances in optical sensor technology have opened new opportunities to study biogeochemical processes in aquatic environments at spatial and temporal scales that were not possible before. Optical sensors are capable of sampling at frequencies that match the sub-metre and sub-second sampling scales of physical variables such as temperature and salinity and can be used in a variety of ocean-observing platforms including moorings, drifter buoys, and autonomous vehicles. In situ measurements of inherent optical properties (IOPs) such as absorption, scattering, attenuation and fluorescence reveal information on the presence, concentration and composition of particulate and dissolved material in the ocean. Variables such as organic carbon, chlorophyll-a, dissolved organic material, nitrate and total suspended matter, among others, are now estimated routinely from IOPs (e.g., Twardowski et al. 2005). Retrieval of seawater constituents from in situ (bulk) IOP measurements is not a straightforward problem — aquatic systems are complex mixtures of particulate and dissolved material, of which each component has specific absorption, scattering and fluorescence characteristics. In situ IOP measurements provide a measure of the sum of the different properties of all individual components present in the water column. Interpretation of optical data and its 1


WILHELMINA R. CLAVANO, EMMANUEL BOSS & LEE KARP-BOSS

successful application to studies of biogeochemical processes thus requires an understanding of the relationships between the different biogeochemical constituents, their optical characteristics and their contribution to bulk optical properties. Suspended organic and inorganic particles play an important role in mediating biogeochemical processes and significantly affect IOPs of aquatic environments, as can be attested from images taken from air- and space-borne platforms of the colour of lakes and oceans where phytoplankton blooms and suspended sediment have a strong impact (e.g., Pozdnyakov & Grassl 2003). Interactions of suspended particles with light largely depend on the physical characteristics of the particles, such as size, shape, composition and internal structure (e.g., presence of vacuoles). Optical characteristics of marine particles have been studied since the early 1940s (summarised by Jerlov 1968) and, with an increased pace, since the 1970s (e.g., Morel 1973, Jerlov 1976). In the past decade, development of commercial in situ optical sensors and the launch of several successful ocean-colour missions have accelerated the efforts to understand optical characteristics of marine particles, in particular the backscattering coefficient because of its direct application to remote sensing (e.g., Boss et al. 2004). These efforts, which have focused on both the theory and measurement of IOPs of particles, are summarised in books, book chapters and review articles on this topic (Shifrin 1988, Stramski & Kiefer 1991, Kirk 1994, Mobley 1994, Stramski et al. 2004, Jonasz & Fournier 2007, and others). Although considerable effort has been given to the subject of marine particles and their IOPs, there is still a gap between theory and the reality of measurement. Such a gap is attributed to both instrumental limitations (e.g., Jerlov 1976, Roesler & Boss 2007) and simplifying assumptions used in theoretical and empirical models (e.g., Stramski et al. 2001). The majority of theoretical investigations on the IOPs of marine particles assume that particles are homogeneous spheres. Optical properties of homogeneous spheres are well characterised (see Mie theory in, e.g., Kerker 1969, van de Hulst 1981) and there is good agreement between theory and measurement for such particles. Mie theory has been used to model IOPs of aquatic particles (e.g., Stramski et al. 2001) and in retrieving optical properties of oceanic particles (e.g., Bricaud & Morel 1986, Boss et al. 2001, Twardowski et al. 2001) with varying degrees of success. For example, while phytoplankton and bacteria dominate total scattering in the open ocean, based on Mie theory calculations for homogeneous spheres, they account for only a small fraction ( 1). A sphere is a spheroid with an aspect ratio of one.

non-spherical homogeneous particles addressing the wide range of particle sizes and indices of refraction relevant to aquatic systems is presented here. Exact analytical solutions are available for a limited number of shapes and physical characteristics (e.g., cylinders and concentric spheres larger than the wavelength and with an index of refraction similar to the medium, Aas 1984), but advances in computational power have enabled the growth of numerical and approximate techniques that permit calculations for a wider range of particle shapes and sizes (Mishchenko et al. 2000 and references therein). It is not realistic to develop a model for all possible shapes of marine particles but in order to cover the range of observed shapes, from elongated to squat geometries, a simple and smooth family of shapes — spheroids — is used here to model particles. Spheroids are ellipsoids with two equal equatorial axes and a third axis being the axis of rotation. The ratio of the axis of rotation, s, to an equatorial axis, t, is the aspect ratio, s/t, of a spheroid (Figure 1). The family of spheroids include oblate spheroids (s/t < 1; disc-like bodies), prolate spheroids (s/t > 1; cigar-shaped bodies), and spheres (s/t = 1). Spheroids provide a good approximation to the shape of phytoplankton and other planktonic organisms that often dominate the IOP signal. Furthermore, by choosing spheroids of varying aspect ratios as a model, solutions for elongated and squat shapes can easily be compared with solutions for spheres and the biases associated with optical models that are based on spheres can be quantified. This review focuses on marine particles because the vast majority of studies on IOPs of aquatic particles have been done in the marine context. However, the results presented here apply to particles in any other aquatic environment.

Bulk inherent optical properties (IOPs) Definitions Inherent optical properties (IOPs) refer to the optical properties of the aquatic medium and its dissolved and particulate constituents that are independent of ambient illumination. To set the stage for an IOP model of non-spherical particles, a brief description of the parameters that define the IOPs of particles is given here. For a more extensive elaboration on IOPs, the reader is referred to Jerlov (1976), van de Hulst (1981), Bohren & Huffman (1983) and Mobley (1994). Most of the notation used in this review follows closely that used by the ocean optics community (e.g., Mobley

3



1994). A summary of the notation along with their definitions and units of measure is provided in the Appendix (see p. 37). Light interacting with a suspension of particles can either be transmitted (remain unaffected) or attenuated due to absorption (transformed into other forms of energy, e.g., chemical energy in the case of photosynthesis) and due to scattering (redirected). Neglecting fluorescence, the two fundamental IOPs are the absorption coefficient, a(λ), and the volume scattering function (VSF), β(θ,λ), where λ is the incident wavelength and θ is the scattering angle. All other IOPs discussed here can be derived from these two IOPs. Other IOPs not discussed in the current review include the polarisation characteristics of scattering and fluorescence. While all quantities are wavelength dependent, the notation is henceforth ignored for compactness. The absorption coefficient, a, describes the rate of loss of light propagating as a plane wave due to absorption. According to the Beer-Lambert-Bouguer law (e.g., Kerker 1969, Shifrin 1988), the loss of light in a purely absorbing medium follows (Equation 11.1 in Bohren & Huffman 1983): E ( R ) = E (0 )e − aR [ W m −2 nm −1 ] ,

(1)

where E(R) is the incident irradiance at a distance R from the light source with irradiance E(0) [W m–2 nm–1]. The light source and detector are assumed to be small compared with the path length and the light is plane parallel and well collimated. The absorption coefficient, a, is thus computed from  1   E ( R )  −1 a = −   ln   [m ] .  R   E (0 ) 

(2)

This equation reveals that the loss of light due to absorption is a function of the path length and that the decay along that path is exponential. In a scattering and absorbing medium, such as natural waters, the measurement of absorption requires the collection of all the scattered light (e.g., using a reflecting sphere or tube). The volume scattering function (VSF), β(Ψ), describes the angular distribution of light scattered by a suspension of particles toward the direction Ψ [rad]. It is defined as the radiant intensity, dI(Ω) [W sr –1 nm–1] (Ω [sr] being the solid angle), emanating at an angle Ψ from an infinitesimal volume element dV [m3] for a given incident irradiant intensity, E(0): β(Ψ ) =

1 dI (Ω) [ m −1sr −1 ] . E (0 ) dV

(3)

It is often assumed that scattering is azimuthally symmetric so that β(Ψ ) = β(θ) , where θ [rad] is the angle between the initial direction of light propagation and that to which the light is scattered irrespective of azimuth. The assumption of azimuthal symmetry is valid for spherical particles or randomly oriented non-spherical particles. This assumption is most likely valid for the turbulent aquatic environment of interest here; it is assumed throughout this review and is further addressed in the following discussion. A measure of the overall magnitude of the scattered light, without regard to its angular distribution, is given by the scattering coefficient, b, which is the integral of the VSF over all (4π[sr]) angles:

4


INHERENT OPTICAL PROPERTIES OF NON-SPHERICAL MARINE-LIKE PARTICLES

b≡

∫

4π

β(Ψ )dΩ =

0

2π

∫ ∫ 0

π

β(θ, ϕ)sin θdθdϕ = 2π

0

∫

π

β(θ) sin θdθ [m −1 ] ,

(4)

0

where ϕ [rad] is the azimuth angle. Scattering is often described by the phase function, β (θ) , which is the VSF normalised to the total scattering. It provides information on the shape of the VSF regardless of the intensity of the scattered light: β(θ) −1 [sr ] . β θ ≡ b

()

(5)

Other parameters that define the scattered light include the backscattering coefficient, bb, which is defined as the total light scattered in the hemisphere from which light has originated (i.e., scattered in the backward direction): bb ≡

∫

2π

β(Ψ )dΩ = 2π

0

∫

π π 2

β(θ)sin θdθ [m −1 ] ,

(6)

and the backscattering ratio, which is defined as b b ≡ b [dimensionless]. b

(7)

Finally, the attenuation coefficient, c, describes the total rate of loss of a collimated, monochromatic light beam due to absorption and scattering: c = a + b [ m −1 ] ,

(8)

which is the coefficient of attenuation in the Beer-Lambert-Bouguer law (see Equation 1) in an absorbing and/or scattering medium (Bohren & Huffman 1983): E ( R ) = E (0 )e − cR [ W m −2 nm −1 ] .

(9)

When describing the interaction of light with individual particles it is convenient to express a quantity with dimensions of area known as the optical cross section. An optical cross section is the product of the geometric cross section of a particle and the ratio of the energy attenuated, absorbed, scattered or backscattered by that particle to the incident energy projected on an area that is equal to its cross-sectional area (denoted by Cc, Ca, Cb and C bb , respectively). For a nonspherical particle, the cross-sectional area perpendicular to the light beam, G [m2], depends on its orientation. In the case when particles are randomly oriented, as assumed here, it has been found that for convex particles (such as spheroids) the average cross-sectional area perpendicular to the beam of light (here denoted as 〈G 〉 ) is one-fourth of the surface area of the particle (Cauchy 1832). In analogy to the IOPs (Equation 8), the attenuation cross section is equal to the sum of the absorption and scattering cross sections: C c = C a + C b [m 2 ] .

5

(10)



Many theoretical texts on optics focus on optical efficiency factors, Qc,a,b,bb , in their treatment of light interaction with particles (e.g., van de Hulst 1981). Optical efficiency factors are the ratios of the optical cross sections to the particle cross-sectional area; their appeal is in that efficiency factors of compact particles are bounded (i.e., their values rarely exceed three) and their values for particles much larger than the wavelength are constant and independent of composition (see below). For non-spherical particles efficiency factors for attenuation, absorption, scattering and backscattering, respectively, are defined as (e.g., Mishchenko et al. 2002):

Qc,a,b,bb ≡

Cc,a,b,bb [dimensionless]. 〈G 〉

(11)

Other useful optical parameters are the volume-normalised cross sections defined as:

α c ,a ,b ,bb ≡

C c ,a ,b ,bb V

[ m −1 ] ,

(12)

where V [ m −3 ] is the particle volume; they provide insight into what size particle most effectively affects light per unit volume (or per unit mass, see Bohren & Huffman 1983, and Figure 6 in Boss et al. 2001). To relate IOPs to optical cross sections, efficiency factors and volume-normalised cross sections, information on particle concentration (and size distribution, see below) is required. For example, for N identical particles within a unit volume, the relations are given by: c, a, b, bb = NCc,a,b,bb = N 〈G 〉Qc,a,b,bb = NV α c,a,b,bb [m −1 ].

(13)

Characteristics of particles affecting their optical properties Three physical characteristics of homogeneous particles determine their optical properties: the complex index of refraction relative to the medium in which the particle is immersed, the size of the particle with respect to the wavelength of the incident light and the shape of the particle. For non-spherical particles, specifying the orientation of the particle in relation to the light beam is an additional requirement. To continue to set the stage for an optical model for non-spherical particles, the physical characteristics of marine particles are discussed in this section and the values that are used to parameterise them in the current study are provided. Index of refraction The complex index of refraction comprises real, n, and imaginary, k, parts: m = n + ik [dimensionless] .

(14)

The real part is proportional to the ratio of the speed of light within a reference medium to that within the particle. It is convenient to choose the reference medium to be that in which the particle is immersed, in which case the proportionality constant is one. The imaginary part of the index of refraction (referred to as the absorption index, e.g., Kirk 1994) represents the absorption of light

6



as it propagates through the particle. It is proportional to the absorption by the intra-particle material, α*[nm–1]: k=

α ∗λ [dimensionless]. 4π

(15)

These definitions are independent of particle shape. For purposes of biogeochemical and optical studies it is often convenient to group aquatic particles into organic and inorganic pools. Organic particles comprise living (viruses, bacteria, phytoplankton and zooplankton) and non-living material (faecal pellets, detritus; although these are likely to harbour bacteria). Inorganic particles consist of lithogenous minerals (quartz, clay and other minerals) and minerals associated with biogenic activity (calcite, aragonite and siliceous particles). Particles in each of these two main groups share similar characteristics with respect to their indices of refraction. Living organic particles often have a large water content (Aas 1996), making them less refractive than inorganic particles. The real part of the index of refraction of aquatic particles ranges from 1.02 to 1.2; the lower range is associated with organic particles while the upper range is associated with highly refractive inorganic materials (Jerlov 1968, Morel 1973, Carder et al. 1974, Aas 1996, Twardowski et al. 2001). The imaginary part of the index of refraction spans from nearly zero to 0.01, with the latter associated with strongly absorbing bands due to pigments (e.g., Morel & Bricaud 1981, Bricaud & Morel 1986). This review aims to primarily illustrate the effects of shape as it applies to two ‘representative’ particle types: phytoplankton with m = 1.05 + i0.01 and inorganic particles with m = 1.17 + i0.0001 (Stramski et al. 2001). Varying the real and imaginary parts of the index of refraction among the values of the two illustrative particles chosen here showed similar dependence on changes in index of refraction to those observed in spheres (van de Hulst 1981, Herring 2002) and was not found to provide additional insight into the effects of shape on IOPs. Size Size is a fundamental property of particles that determines sedimentation rates, mass transfer to and from the particle (e.g., nutrient fluxes and dissolution), encounter rates between particles and, most relevant to this review, their optical properties. Foremost, the ratio of particle size to wavelength determines the resonance characteristics of the VSF (its oscillatory pattern as a function of scattering angle) and the size for which maximum scattering per volume will occur (i.e., maximum αb). In addition, in general, the larger an absorbing particle is, the less efficient it becomes in absorbing light per unit volume (i.e., the volume-normalised absorption efficiency, αa, decreases with increasing size), often referred to as the package effect or self-shading (see Duysens 1956). In both marine and freshwater environments particles relevant to optics span at least eight orders of magnitude in size, ranging from sub-micron particles (colloids and viruses) to centimetresize aggregates and zooplankton (Figure 2). Numerically, small particles are much more abundant than larger particles. A partitioning of particles into logarithmic size bins shows that each bin includes approximately the same volume of particulate material (Sheldon et al. 1972). This observation is consistent with a Junge-like (power-law) particulate size distribution (PSD), where the differential particle number concentration is inversely proportional to the fourth power of size (Junge 1963, Morel 1973; see p. 22). Several other distribution functions have been used to represent size distributions of particles in the ocean, which include the log-normal distribution (Jonasz 1983, Shifrin 1988, Jonasz & Fournier 1996), the Weibull distribution (Carder et al. 1971), the gamma distribution (Shifrin 1988)

7



Rayleigh 0.1 nm

1 nm

Rayleigh−Gans−Debye 0.1 µm

10 nm

Van de Hulst 1 µm

10 µm

Geometric Optics 100 µm

1 mm

1 cm

Dissolved organic matter Water molecules

Suspended particulate matter Truly soluble substances Colloids Viruses Bacteria Phytoplankton: Pico-

Nano-

MicroZooplankton

Organic detritus, minerogenic particles Bubbles 10−10

10−9

10−8

10−7

10−6 10−5 Particle size (m)

10−4

10−3

10−2

Figure 2 Representative sizes of different constituents in sea-water, after Stramski et al (2004). Optical regions referred to in the text are denoted at the top axis (shading represents approximate boundaries between these regions). These boundaries vary with refractive index for a given particle size.

and sums of log-normal distributions (Risoviç 1993). Here, the focus is on particles ranging in diameter from 0.2 to 200 µm (diameter here is given by that of an equal-volume sphere). The lower bound is associated with a common operational cutoff between dissolved and particulate material — often set by a filter with that pore size — and the upper bound chosen arbitrarily to represent the upper bound of particles that can still be assumed to be distributed as a continuum in operational measurements (Siegel 1998). Two particulate size distributions are adopted (as in Twardowski et al. 2001) for the illustrative optical model used in this study: the power-law distribution and that described by Risoviç (1993). Shape Several measures have been used to characterise the shape of particles in nature; some focus on the overall shape while others concentrate on specific features such as roundness and compactness. An elementary measure of particle shape is the aspect ratio, which is the ratio of the principal axes of a particle. It describes the elongation or flatness of a particle and hence the deviation from a spherical shape (a sphere having an aspect ratio of one). Shape effects on optical properties are examined here by modelling the IOPs of spheroids of varying aspect ratios. Aquatic particles vary greatly in their shape; most notable is the striking diversity in cell shapes among phytoplankton. Hillebrand et al. (1999) provides a comprehensive survey of geometric models for phytoplankton species from 10 taxa. Two relevant results arise from their analysis: (1) the sphere is not a common shape among microphytoplankton taxa and (2) despite the apparent high diversity of cell geometries, the diverse morphologies represent variations on a smaller subset of geometric forms, primarily ellipsoids, spheroids and cylinders. Picoplankton, which are not included in the analysis of Hillebrand et al. (1999), tend to be more spherical in shape, although rod-like morphologies are also common.

8



Number of cells per size bin (N)

The authors are not aware of any published paper that provides the range of values of aspect ratios of phytoplankton cells in natural assemblages. To demonstrate the deviation from a spherical shape among phytoplankton, field data on cell dimensions of different taxonomic groups (nanoand microphytoplankton) were used to calculate aspect ratios of phytoplankton (Figure 3; data available from the California State Department of Water Resources). Aspect ratios of phytoplankton span a wide range, varying between 0.4 and 72 (Figure 3). Diatom chains, which are not included in the analysis, can have even higher aspect ratios. The frequency distribution of the aspect ratios shows that elongated shapes are a more common form compared with spheres or squat shapes (Figure 3). Inorganic aquatic particles are very often non-spherical; clay mineral particles have plate-like crystalline structures with sizes on the order of D = 0.5 µm and have aspect ratios varying between 0.05 and 0.3 (Jonasz 1987b, Bickmore et al. 2002). In nature, clays tend to aggregate and form larger particles with reduced aspect ratios. It is not possible to generalise their shapes except to say that they are extremely variable and do not look like spheres. Larger sedimentary particles such as sand and silt have aspect ratios ranging between 0.04 and 11 (derived from Komar & Reimers 1978, Baba & Komar 1981). Consistent with these observations, spheroids with aspect ratios between 0.1 and 46 are used in the analysis of IOPs of non-spherical particles presented here (98% of the cells that constitute the data in Figure 3 are within this range). Finer-scale structures that may be found in each particle do not dominate scattering, in general, as much as the effect of the

3000 2500 2000 N = 8059

1500 1000 500 0

0.5

1

2

5 10 Aspect ratio

20

50

Figure 3 Frequency distribution of aspect ratios of phytoplankton. Data are provided by the California State Department of Water Resources and the U.S. Bureau of Reclamation and are available on the Bay-Delta and Tributaries (BDAT) project website at http://baydelta.water.ca.gov/. A subset of the data was randomly selected for the analysis here and includes data collected during the period 2002–2003 from a variety of aquatic habitats: from freshwater in the Sacramento-San Joaquin Delta to estuarine environments in the Suisun and San Pablo Bays (California, USA). The data include phytoplankton from five different classes, including Bacillariophyceae (diatoms), Chlorophyceae, Cryptophyceae, Dynophyceae, and Cyanophyceae (N = 8059 cells). Phytoplankton analyses (identification, counts, and measurements of cell dimensions) were conducted at the Bryte Chemical Laboratory (California Department of Water Resources). Further information on the methods used can be found at http://iep.water.ca.gov/emp/Metadata/Phytoplankton/. The aspect ratio is calculated as the ratio between the rotational and equatorial axes of a cell based on the three-dimensional shape associated with each species as provided in Hillebrand et al. (1999). The reader is cautioned on the fact that the phytoplankton data do not include picophytoplankton (i.e., cells smaller than 2 µm) that tend to be more spherical in shape.

9



‘gross’ shape of the particle (Gordon 2006). Furthermore, Gordon (2006) found that, in theory, the total scattering of any curved shape (that is not rotationally symmetric) will behave similarly for a given particle thickness and cross-sectional area. However, when a particle exhibits sharp edges, smooth shapes are not able to reproduce the sharp spikes observed in the forward scattering (Macke & Mishchenko 1996). To allow comparisons between spheroids and spheres, particle size is used as a reference. The definition of size is often ambiguous when dealing with non-spherical particles; here the size of a spheroid is defined as the diameter of an equal-volume sphere ( D = 2 3 st 2 ). This was chosen for two main reasons: (1) popular particle sizers such as the Coulter counter are sensitive to particle volume and (2) mass, which is most often the property of interest in studies of particles, is proportional to particle volume. Size and shape, however, may not be independent attributes for aquatic particles. There appears to be a tendency for particles in ocean samples to deviate from a spherical shape as particle size increases (Jonasz 1987b). This trend has been observed for particles in both coastal (Baltic Sea) and offshore areas (Kadyshevich 1977, Jonasz 1987a). Shape effects on IOPs are examined here for two types of particulate populations: monodispersions (comprising particles with one size and one shape) and polydispersions (comprising particles with varying sizes and shapes) and are quantified by defining a bias, γ c ,a ,b ,bb, which is the ratio of the IOPs (attenuation, absorption, scattering and backscattering, respectively) of spheroids to that of spheres with the same particle volume distribution. Orientation In this review particles are assumed to be randomly oriented. IOPs of non-spherical particles, however, are strongly dependent on particle orientation (e.g., Latimer et al. 1978, Asano 1979) but data on the orientation of particles found in the natural marine environment are practically nonexistent. There are certain cases for which the assumption of random orientation may not apply because of methodological issues or because environmental conditions cause particles to align in a preferred orientation. Non-random orientation associated with methodology will be encountered when: (1) the instrument used to measure an IOP causes particles to orient themselves relative to the probing light beam (e.g., the flow cytometer in which particles are aligned one at a time within the flow chamber) and (2) when the existence of particles of a given sub-population (e.g., big diatom chains) is rare enough in the sample volume such that not all orientations are realised in a given measurement. In the latter case, averaging over many samples is necessary to randomise orientations. In the natural environment, shear flows can result in the alignment of particles with respect to the flow (e.g., Karp-Boss & Jumars 1998). When the environment is quiescent enough, large aggregates are oriented by the force of gravity as can be seen in photographs of in situ long stringers and teardrop-shape flocs (e.g., Syvitski et al. 1995). The following optical characteristics can be used to assess whether or not an ensemble of particles is randomly oriented (Mishchenko et al. 2002): (1) the attenuation, scattering and absorption coefficients are independent of polarisation and instrument orientation; (2) the polarised scattering matrix is block diagonal; and (3) the emitted blackbody radiation is unpolarised. Note that care should be applied so that the measurement procedures have minimal effect on the orientation of the particles investigated. Given that the orientation of aquatic particles is currently unconstrained we proceed in this review by assuming random orientation. Future studies, however, may find orientation effects to be important under certain conditions as was found in atmospheric studies due, for example, to orientation of particles under gravity (e.g., Aydin 2000).

10



Optical regimes A century and a half of theoretical studies on the interaction of light and particles has taught us that this interaction is strongly dependent on several parameters. First among them is the size parameter, x, which is defined as the ratio of the particle size to the wavelength: x=π

D [dimensionless], λ

(16)

where D is the particle size and λ is the wavelength of light within the medium (both with the same units), in this case water. An additional important parameter is the ratio of the speed of light within the particle to that in the medium (it is the reciprocal of the real part of the index of refraction of the particle to that of water, n). Marine particles are mostly considered to be ‘soft’; their index of refraction is close to that of water, that is, m − 1 ≈ n − 1 1. Finally, another important parameter is the phase shift parameter, ρ, which describes the shift in phase between the wave travelling within the particle and the wave travelling in the medium surrounding it and is a function of both the size parameter and the index of refraction of that particle:

(

)

ρ = 2 x n − 1 [dimensionless].

(17)

These parameters are useful to delineate optical regimes for which analytical approximations that apply to soft particles have been developed (see below). The material in this section borrows heavily from Bohren & Huffman (1983), Mishchenko et al. (2002) and Kokhanovsky (2003), where more details can be found. Many of the approximations discussed in these references are applicable to randomly oriented non-spherical particles (as in the case of marine particles) and help establish an intuition for their optical characteristics when compared with spheres. The characteristics of particles (size and index of refraction) most emphasised in Bohren & Huffman (1983), Mishchenko et al. (2002) and Kokhanovsky (2003), however, are significantly different from those of marine particles.

Particles much smaller than the wavelength The Rayleigh region (RAY) ( x 1, ρ 1, D λ ) In this optical region shape does not contribute to the optical properties of particles; for a given wavelength, the IOPs are only dependent on particle volume and its index of refraction (e.g., Kerker 1969, Bohren & Huffman 1983, Kokhanovsky 2003): 3 1 + cos 2 (θ)  β θ =  [dimensionless], 4

()

Cc =

k 2V 2 m 2 − 1

Ca =

6π

2

[m 2 ],

4 πkV [m 2 ], λ 11

(18)

(19)

(20)



Cb = Cc − Ca [m 2 ], Cbb =

Cb [m 2 ], 2

(21) (22)

where k = 2π/λ [nm–1] is the wave number. Since the IOPs are a function of only particle volume, incident wavelength and index of refraction (Equations 18–22), there is no difference between the IOPs of non-spherical particles and equal-volume spheres. In the marine environment, small organic and inorganic dissolved molecules fall within this regime.

Particles of size much larger than the wavelength The geometric optics (GO) region ( x 1, ρ 100, D λ ) In this optical region scattering is dominated by diffraction although refraction effects introduce a necessary correction for intermediate values of the size parameter (known as ‘edge effects’, e.g., Kokhanovsky & Zege 1997). An analytical solution has been derived for the attenuation cross section of absorbing particles of random shape in this region (e.g., Kokhanovsky & Zege 1997) and is given by: 

Cc = 2  1 + x 

−2  3 

〈G 〉 [m 2 ].

(23)

The absorption cross section, Ca, can also be derived analytically. In general, it is a complex function of both parts of the index of refraction and x (e.g., Kokhanovsky & Zege 1997). For sizes where kx 1, it simplifies to C a = 〈G 〉. Within the GO region, these analytical solutions imply that the attenuation, absorption, and scattering (but not the VSF) of a randomly oriented non-spherical particle will be the same as that of a sphere of the same cross-sectional area, that is, it will approach the geometric optics limit (Kerker 1969): lim γ c ,a ,b =

ρ→∞

〈G 〉 ≥ 1. G

(24)

Given that the average cross-sectional area of a sphere is always the smallest of any convex shape of the same volume, an equal-volume sphere will always underestimate the IOPs of particles much larger than the wavelength. The VSF in this regime for known shapes (including spheroids) can be obtained from ray tracing computations (see below). Particles that fall in this region in the marine environment include large diatom chains, large heterotrophs (e.g., Noctiluca sp.), mesoand macrozooplankton and macrosize aggregates, including faecal pellets.

Particles of size comparable to or larger than the wavelength The Rayleigh-Gans-Debye (RGD) (x < 1, ρ < 1, D ≈ λ) and the van de Hulst (VDH) (x > 1, 1 < ρ < 100, D > λ) regions The RGD and VDH optical regions are of particular interest because many optically relevant marine particles (e.g., phytoplankton and sediments) fall within them. However, no simple closed-form analytical solution exists for randomly oriented non-spherical particles in these regions (Aas 1984). 12



Scattering by soft particles in the RGD and VDH regions is dominated by diffraction although contributions from reflection and refraction need to be taken into account. Absorption is assumed to be independent of the real part of the index of refraction, although more recent approximations have included n effects on absorption (Kokhanovsky & Zege 1997). Simple analytical solutions for Cc, Ca and Cb have been derived for spheres and for some simple shapes by van de Hulst (1981) and Aas (1984). Shepelevich et al. (2001), following Paramonov (1994a,b), derive Cc, Ca and Cb for randomly oriented monodispersed spheroids from a polydispersed population of spheres having the same volume and cross-sectional area. A similar approach is used here to examine the IOPs of non-spherical marine-like particles but, rather than follow Shepelevich et al. (2001) who used the approximation given by van de Hulst (1981) to obtain the optical values for spheres, values for spheres are derived here directly from Mie theory. Size ranges of aquatic constituents and optical regions are provided in Figure 2 for the particular wavelength (λ = 676 nm) and the specific refractive indices (n = 1.05, 1.17) used in this review. Results for other visible wavelengths are not expected to be very different and can be deduced from the results presented here by changing the diameter while keeping x constant. Similarly, the indices of refraction used here span the range of those of marine particles thus bounding the likely results for all relevant marine particles. The sizes associated with the different optical size regions are provided in Table 1.

IOPs of monodispersions of randomly oriented spheroids Exact and approximate methods Since the 1908 paper by Mie there is now an exact solution (in the form of a series expansion) providing the optical properties of a homogeneous sphere of any size and index of refraction relevant to aquatic optics. Unfortunately, there is no equivalent converging solution for non-spherical particles for all relevant sizes. Asano & Yamamoto (1975) obtained an exact series solution for scattering by spheroids of arbitrary orientation but their solution did not converge for size parameters >30. Obtaining optical properties of non-spherical particles for the wide range of sizes exhibited by marine particles requires the use of several methods, each valid within a specific optical region. The appropriate application of each of these approaches depends on the combination of sizes, shapes and refractive indices of the particles of interest. For small particles the T-matrix method (Waterman 1971, cf. Mishchenko et al. 2000), which is an exact solution to Maxwell’s equations for light scattering, applies. This method is limited to particles with a phase shift parameter that is smaller than approximately 10 (it covers particles with phase shift parameters as large as those in the RGD region, see Table 1). As particles deviate from a spherical shape the phase shift parameter for which this method is valid decreases. For larger particles, a variety of methods that provide approximate solutions for optical properties have been used (see Mishchenko et al. 2002 for a review of the state of the art). Table 1 Size ranges roughly corresponding to the size regions defined for two different refractive indices given λ = 676 nm Size region RAY RGD VDH GO

n = 1.05

n = 1.17

Equivalent ρ

D 0.2 µm D < 5 µm 5 < D < 200 µm D 200 µm

D 0.2 µm D < 2 µm 2 < D < 65 µm D 65 µm

ρ 0.1 ρ1. The geometric cross section of an elongated cylinder, however, is very similar to that of a prolate spheroid and so prolate spheroids are used here to model larger particles. Thus, the deviation from sphericity of a particle can be expressed in terms of its aspect ratio, s/t, and diameter of its equal-volume sphere, D, and Equation 29 becomes:  −2 1  s  3 + 2  t   

()

1

s 3 t

sin −1 1 − 1−

() s t

()

−2

s t

−2

  0.22  = 1.28 D .  

(30)

Given a size D, this equation is solved to obtain s/t, which is used in the population model with aspect ratios varying as a function of size (see also Figures 2 and 3 in Jonasz 1987b).

Results for polydispersions In the following section, the modelled IOPs (c, a and b) of polydispersions of spheroids are presented. Due to the inability to obtain the VSF of spheroids throughout the size range of interest, results regarding either the VSF or the backscattering coefficient, bb, are not presented here. For polydispersions of spheroids, shape effects depend on the relative contributions of small and large particles to the population and the degree to which particles deviate from a spherical shape (as indicated by the aspect ratio). In both the power-law and Risoviç (1993) PSD simulations, with constant and varying aspect ratios, the biases of all the IOPs increase with increasing proportion of large particles in the population (i.e., as ξ → 3 or as nS : nL → 1012, Figures 13 and 14). This is a direct consequence of the nearly monotonic change in the bias as a function of size for a monodispersion (Figure 8). As expected from the results for monodispersions of spheroids, the biases in attenuation and scattering increase as the aspect ratio departs from one, the absorption bias also increases with departure from sphericity and with increasing absorption index. In most cases the biases are >1 (i.e., a spherical model will underestimate a population of spheroids), being 25



2

A

1.8

1.6

1.6

1.4

1.4

γc

γc

1.8

2

m = 1.05 + i0.01

m = 1.17 + i0.0001

D

m = 1.17 + i0.0001

F

1

1

0.8

0.8 m = 1.05 + i0.01

C

1.4

1.6

1.3 γa

γa

B

1.2

1.2

1.8

m = 1.17 + i0.0001

1.4

1.2 1.1

1.2

1 1 2

2

m = 1.05 + i0.01

E

1.8

1.6

1.6

1.4

1.4

γb

γb

1.8

1.2

1.2

1

1

0.8 0.6

0.8 3.5

3.75

4

4.25 4.5

3.5

4

4.25 4.5

ξ

ξ

x

3.75

10

Aspect ratio st 2

1

0.5

0.1

Figure 13 The bias in attenuation, γc (A, B), absorption, γa (C, D), scattering, γb (E, F), and backscattering, γbb (G, H), for a power-law polydispersion of spheroids relative to a power-law polydispersion of spheres with the same volume as a function of the power-law exponent, ξ. Each line represents a different aspect ratio, s/t (legend below the plot). The grey line with dots (legend: ‘x’) denotes the polydispersions of spheroids where the shape co-varies with size following Jonasz (1983; see text). The dotted vertical lines are used to compare equivalent size distributions in Figure 14.

1 and can be greater by as much as a factor of seven (95% of the time in Figure 9B) for specific sizes of phytoplankton-like particles. For particles with a very large absorption coefficient (unrealistic for marine particles), an asymptotic value similar to the other IOPs is reached (Herring 2002), suggesting that in general, for particles larger than the wavelength, the backscattering should be more enhanced compared with that of equal-volume spheres. Despite the complexity observed, it seems sensible to conclude that the backscattering of spheroids is likely to be significantly larger than that of equal-volume spheres for the sizes relevant to phytoplankton (Figure 8G,H). In this respect, shape may be a factor contributing to the inability to account for the bulk backscattering coefficient in the ocean, when spheres are used as a model for natural particles (e.g., Stramski et al. 2004). Indeed, Morel et al. (2002) used a mixture of prolate and oblate spheroidal particles (using the T-matrix method) to generate the phase function 31



of small phytoplankton-like particles that was more realistic in the backward directions compared with that derived from spheres. For polydispersions of particles with constant or varying shape as a function of size, the biases in attenuation, absorption and scattering have been found here to be bounded, reaching high values (270%) only for extreme shapes and size distribution parameters but generally being within about 50% of that of spheres (Figures 13 and 14). While not as large as for monodispersions, these biases are significant and most often >1, implying that populations of spherical particles perform poorly as an average, unbiased model. Diffraction-based instruments provide an opportunity to measure particle size in situ. Given that measurements are made for angular scattering and that inversions from optical measurements to obtain particle size are based on Mie theory, shape may cause significant biases for the sizing of particles. A population of non-spherical particles will appear, on average, larger (and more dispersed) than a population of equal-volume spheres (Figure 11). In addition, such an inversion will ‘create’ populations at the tail ends of the size distribution due to the fact that the non-spherical particles have no resonance pattern in the near-forward scattering as a function of angle (in contrast to spheres, see Figures 4, 16 and 17; see also Heffels et al. 1996). Shape is likely to have some effect on optical inversions that are based on Mie theory. In such inversions, IOPs are used to predict the physical characteristics of the underlying bulk particulate population. For example, the imaginary part of the index of refraction of phytoplankton has been found by inverting absorption data using measured size distributions and Mie theory (Bricaud & Morel 1986). Based on the results of this paper, the inverted k is likely to be an overestimate, with the bias increasing with increasing phytoplankton size and departure from sphericity. Similarly, an inversion of the backscattering ratio was used to obtain the real part of the index of refraction for populations of particles with a power-law size distribution, assuming spherical particles (Twardowski et al. 2001, Boss et al. 2004). Results of this work suggest that a spherical model is likely to underestimate the index of refraction as deviations from sphericity will enhance the backscattering ratio, thus increasing the bias of the inverted index of refraction. Shape effects, on the other hand, were not found to significantly change the spectral slope of the beam attenuation (Boss et al. 2001) and thus are not likely to significantly affect the inversion of this parameter to obtain information on the particulate size distribution. Given the inherent biases associated with using spheres as models for natural particles, it is sensible to predict that inversions that include nonspherical characteristics should provide an improvement compared to those based on Mie theory. This has been the case in several atmospheric studies (e.g., Dubovik et al. 2002, Zhao et al. 2003, Kocifaj & Horvath 2005). Shape has important effects on the polarisation of light scattered by marine particles but is a topic which is beyond the focus of this review. Nevertheless, it is one of the future frontiers in ocean optics, as currently there is no in situ commercial instrumentation able to measure polarised scattering. The aquatic community has largely neglected polarisation when studying particulate suspensions (with a few exceptions, e.g., Quinby-Hunt et al. 2000 and references therein). Studies by Geller et al. (1985) and Hoovenier et al. (2003) suggest that there is promise in obtaining information regarding some aspects of particle shape (e.g., departure from sphericity) by analysing certain elements of the polarised scattering matrix. For example, theoretical shape indices have been derived based on both linear (Kokhanovsky & Jones 2002) and circular (Hu et al. 2003) polarisation measurements. In particular, the latter was found to be less sensitive to multiple scattering. Both were found to be most sensitive at scattering angles in the backward hemisphere. Polarimetry shows promise especially for extreme shapes and larger particles (Macke & Mishchenko 1996). Both organic and inorganic aquatic particles are not randomly distributed among shapes but rather tend to span a limited and non-uniform range of aspect ratios, with spheres being relatively 32



rare. Given the limited amount of data available regarding shape distributions of natural particles, more measurements of shape parameters are needed; in particular, these are needed as input to improve inversion models that currently assume spherical particles. Laboratory experiments designed to measure the effects of shape on optical properties and their consistency with the predictions presented here and elsewhere are also required so that a more complete picture of the effect of shape on IOPs can be established.

Acknowledgements We are indebted to J.R.V. Zaneveld, G. Dall’Olmo and H. Gordon for helpful discussions and constructive comments on earlier drafts of this manuscript; D. Risoviç for the delight in sharing the pragmatism of representing particle size distributions; Y.C. Agrawal and A. Briggs-Whitmire for the scattering measurements and pictures of river sediment; G.R. Fournier for insight into analytical solutions to ‘the problem’; J.T.O. Kirk for resurrecting the absorption cross section triple integral that was done on a hand calculator and M.I. Mishchenko for a lifetime of T-matrix code. This project is supported by the Ocean Optics and Biology programme of the Office of Naval Research (Contract No. N00014-04-1-0710) to E. Boss and by NASA’s Ocean Biology and Biogeochemistry research programme (Contract No. NAG5-12393) to L. Karp-Boss.

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WILHELMINA R. CLAVANO, EMMANUEL BOSS & LEE KARP-BOSS Quinby-Hunt, M.S., Hull, P.G. & Hunt, A.J. 2000. Polarized light scattering in the marine environment. In Light Scattering by Nonspherical Particles: Theory, Measurements, and Applications, M.I. Mishchenko et al. (eds). San Diego, California: Academic Press, 525–554. Quirantes, A. & Bernard, S. 2004. Light scattering by marine algae: two-layer spherical and nonspherical models. Journal of Quantitative Spectroscopy and Radiative Transfer 89, 311–321. Quirantes, A. & Bernard, S. 2006. Light-scattering methods for modelling algal particles as a collection of coated and/or nonspherical particles. Journal of Quantitative Spectroscopy and Radiative Transfer 100, 315–324. Risoviç, D. 1993. Two-component model of sea particle size distribution. Deep-Sea Research I 40, 1459–1473. Roesler, C.S. & Boss, E. 2007. In situ measurement of the inherent optical properties (IOPs) and potential for harmful algal bloom detection and coastal ecosystem observations. In Real-Time Coastal Observing Systems for Ecosystem Dynamics and Harmful Algal Blooms, M. Babin et al. (eds). Paris: UNESCO Publishing, in press. Sheldon, R.W., Prakash, A. & Sutcliffe, W.H. Jr. 1972. The size distribution of particles in the ocean. Limnology and Oceanography 17, 327–340. Shepelevich, N.V., Prostakova, I.V. & Lopatin, V.N. 2001. Light-scattering by optically soft randomly oriented spheroids. Journal of Quantitative Spectroscopy and Radiative Transfer 70, 375–381. Shifrin, K.S. 1988. Physical Optics of Ocean Water. New York: American Institute of Physics, AIP translation series. Originally published as Vvedenie v optiku morya 1983. Siegel, D.A. 1998. Resource competition in a discrete environment: Why are plankton distributions paradoxical? Limnology and Oceanography 43, 1133–1146. Stramski, D., Boss, E., Bogucki, D. & Voss, K.J. 2004. The role of seawater constituents in light backscattering in the ocean. Progress in Oceanography 61, 27–56. Stramski, D., Bricaud, A. & Morel, A. 2001. Modeling the inherent optical properties of the ocean based on the detailed comparison of the planktonic community. Applied Optics 40, 2929–2945. Stramski, D. & Kiefer, D.A. 1991. Light scattering by microorganisms in the open ocean. Progress in Oceanography 28, 343–383. Syvitski, J.P.M., Asprey, K.W. & Leblanc, K.W.G. 1995. In-situ characteristics of particles settling within a deep-water estuary. Deep-Sea Research Part II: Topical Studies in Oceanography 42, 223–256. Twardowski, M.S., Boss, E., MacDonald, J.B., Pegau, W.S., Barnard, A.H. & Zaneveld, J.R. 2001. A model for estimating bulk refractive index from the optical backscattering ratio and the implications for understanding particle composition in Case I and Case II waters. Journal of Geophysical Research 106, 14129–14142. Twardowski, M.S., Lewis, M.R., Barnard, A.H. & Zaneveld, J.R. 2005. In-water instrumentation and platforms for ocean color remote sensing applications. In Remote Sensing of Coastal Aquatic Environments: Technologies, Techniques and Applications, R.L. Miller et al. (eds). Dordrecht: Springer, 69–100. van de Hulst, H.C. 1981. Light Scattering by Small Particles. New York: Dover. Volten, H., de Haan, J.F., Hoovenier, J.W., Schreurs, R. & Vassen, W. 1998. Laboratory measurements of angular distributions of light scattered by phytoplankton and silt. Limnology and Oceanography 43, 1180–1197. Voss, K.J. & Fry, E.S. 1984. Measurement of the Mueller matrix for ocean water. Applied Optics 23, 4427–4439. Waterman, P.C. 1971. Symmetry, unitarity, and geometry in electromagnetic scattering. Physical Review D 3, 825–839. Zaneveld, J.R.V., Roach, D.M. & Pak, H. 1974. The determination of the index of refraction distribution of oceanic particles. Journal of Geophysical Research 79, 4091–4095. Zhao, T.X.-P., Laszlo, I., Dubovik, O., Holben, B.N., Sapper, J., Tanré, D. & Pietras, C. 2003. A study of the effect of non-spherical dust particles on the AVHRR aerosol optical thickness retrievals. Geophysical Research Letters 30, 1317–1320.

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APPENDIX: NOTATION Notation used following Mobley (1994) closely. The actual units used are given in the text, however, only the dimensions are provided in this table for mass, M, length, L, and time, T, or angular measure as indicated. Symbol

Definition

Dimension

a b b˜ bb Ca Cb C bb Cc c D D0 E(0) E(R) f (D) G 〈G〉 I k k m N n n0 nS nL Qa Qb Qbb Qc R s s/t t V x α* αa αb αbb αc β(θ) ˜ β(θ) γa γb

Absorption coefficient Scattering coefficient Backscattering ratio Backscattering coefficient Absorption cross section of a particle Scattering cross section of a particle Backscattering cross section of a particle Attenuation cross section of a particle Attenuation coefficient Particle size represented by diameter of an equal-volume sphere Reference diameter of size range Irradiance at the light source Irradiance at distance R from the light source Particulate size distribution Geometrical cross-sectional area of a sphere Average geometrical cross-sectional area of a non-sphere Radiant intensity Imaginary part of the relative index of refraction Wave number of the incident light Complex relative index of refraction Number of particles per unit volume Real part of the relative index of refraction Number concentration of particles at the reference diameter D0 Number concentration of small particles Number concentration of large particles Absorption efficiency factor of a particle Scattering efficiency factor of a particle Backscattering efficiency factor of a particle Attenuation efficiency factor of a particle Arbitrary path length of light Rotational axis of a spheroid Aspect ratio of a spheroid Equatorial axis of a spheroid Particle volume Size parameter Specific absorption coefficient of a particle Volume-normalised absorption cross section Volume-normalised scattering cross section Volume-normalised backscattering cross section Volume-normalised attenuation cross section Volume scattering function (VSF) Volume scattering phase function Absorption bias Scattering bias

L–1 L–1 dimensionless L–1 L2 L2 L2 L2 L–1 L L M L–1 T –3 M L–1 T –3 # L–4 L2 L2 M L–1 T –3 sr –1 dimensionless L–1 dimensionless # L–3 dimensionless # L–4 # L–4 # L–4 dimensionless dimensionless dimensionless dimensionless L L dimensionless L L3 dimensionless L–1 L–1 L–1 L–1 L–1 L–1sr –1 sr –1 dimensionless dimensionless

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Symbol

Definition

Dimension

γbb γc θ λ µS µL ξ ρ τS τL υS υL ϕ Ψ Ω

Backscattering bias Attenuation bias Scattering angle Wavelength of the incident light Small-particle generalised gamma distribution parameter Large-particle generalised gamma distribution parameter Slope of the power-law size distribution Phase shift parameter Small-particle generalised gamma distribution parameter Large-particle generalised gamma distribution parameter Small-particle generalised gamma distribution parameter Large-particle generalised gamma distribution parameter Azimuth angle Angular direction into which light is scattered Solid angle into which light is scattered

dimensionless dimensionless radians (rad) L dimensionless dimensionless dimensionless dimensionless L–1 L–1 dimensionless dimensionless radians (rad) radians (rad) steradians (sr)

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50931_book.fm Page 39 Tuesday, May 1, 2007 4:43 PM


GLOBAL ECOLOGY OF THE GIANT KELP MACROCYSTIS: FROM ECOTYPES TO ECOSYSTEMS MICHAEL H. GRAHAM1,2, JULIO A. VÁSQUEZ2,3 & ALEJANDRO H. BUSCHMANN4 1Moss Landing Marine Laboratories, 8272 Moss Landing Road, Moss Landing, California 95039, U.S. E-mail: [email protected] 2Centro de Estudios Avanzados de Zonas Aridas (CEAZA - www.ceaza.cl) 3Departamento Biología Marina, Facultad de Ciencias del Mar, Universidad Católica del Norte, Coquimbo, Chile 4Centro de Investigación y Desarrollo en Ambientes y Recursos Costeros (i~mar), Universidad de Los Lagos, Casilla 557, Puerto Montt, Chile Abstract The giant kelp Macrocystis is the world’s largest benthic organism and most widely distributed kelp taxon, serving as the foundation for diverse and energy-rich habitats that are of great ecological and economical importance. Although the basic and applied literature on Macrocystis is extensive and multinational, studies of large Macrocystis forests in the northeastern Pacific have received the greatest attention. This review synthesises the existing Macrocystis literature into a more global perspective. During the last 20 yr, the primary literature has shifted from descriptive and experimental studies of local Macrocystis distribution, abundance and population and community structure (e.g., competition and herbivory) to comprehensive investigations of Macrocystis life history, dispersal, recruitment, physiology and broad-scale variability in population and community processes. Ample evidence now suggests that the genus is monospecific. Due to its highly variable physiology and life history, Macrocystis occupies a wide variety of environments (intertidal to 60+ m, boreal to warm temperate) and sporophytes take on a variety of morphological forms. Macrocystis sporophytes are highly responsive to environmental variability, resulting in differential population dynamics and effects of Macrocystis on its local environment. Within the large subtidal giant kelp forests of southern California, Macrocystis sporophytes live long, form extensive surface canopies that shade the substratum and dampen currents, and produce and retain copious amounts of reproductive propagules. The majority of subtidal Macrocystis populations worldwide, however, are small, narrow, fringing forests that are productive and modify environmental resources (e.g., light), yet are more dynamic than their large southern California counterparts with local recruitment probably resulting from remote propagule production. When intertidal, Macrocystis populations exhibit vegetative propagation. Growth of high-latitude Macrocystis sporophytes is seasonal, coincident with temporal variability in insolation, whereas growth at low latitudes tracks more episodic variability in nutrient delivery. Although Macrocystis habitat and energy provision varies with such ecotypic variability in morphology and productivity, the few available studies indicate that Macrocystis-associated communities are universally diverse and productive. Furthermore, temporal and spatial variability in the structure and dynamics of these systems appears to be driven by processes that regulate Macrocystis distribution, abundance and productivity, rather than the consumptive processes that make some other kelp systems vulnerable to overexploitation. This global synthesis suggests that the great plasticity in Macrocystis form and function is a key determinant of the great global ecological success of Macrocystis.

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MICHAEL H. GRAHAM, JULIO A. VÁSQUEZ & ALEJANDRO H. BUSCHMANN

Introduction Kelp beds and forests represent some of the most conspicuous and well-studied marine habitats. As might be expected, these diverse and productive systems derive most of their habitat structure and available energy (fixed carbon) from the kelps, a relatively diverse order of large brown algae (Laminariales, Phaeophyceae; ~100 species). Kelps and their associated communities are conspicuous features of temperate coasts worldwide (Lüning 1990), including all of the continents except Antarctica (Moe & Silva 1977), and the proximity of such species-rich marine systems to large coastal human populations has subsequently resulted in substantial extractive and non-extractive industries (e.g., Leet et al. 2001). It is therefore not surprising that the basic and applied scientific literature on kelps is extensive. Our present understanding of the ecology of kelp taxa is not uniform, as the giant kelp Macrocystis has received the greatest attention. Macrocystis is the most widely distributed kelp genus in the world, forming dense forests in both the Northern and Southern hemispheres (Figure 1). The floating canopies of Macrocystis adult sporophytes also have great structural complexity and high rates of primary productivity (Mann 1973, Towle & Pearse 1973, Jackson 1977, North 1994). Furthermore, although Macrocystis primary production can fuel secondary productivity through direct grazing, most fixed carbon probably enters the food web through detrital pathways or is exported from the system (e.g., Gerard 1976, Pearse & Hines 1976, Castilla & Moreno 1982, Castilla 1985, Inglis 1989, Harrold et al. 1998, Graham 2004). In some regions, such habitat and energy provision can support from 40 to over 275 common species (Beckley & Branch 1992, Vásquez et al. 2001, Graham 2004). Venerated by Darwin (1839), the ecological importance of Macrocystis has long been recognised. The genus, however, did not receive thorough ecological attention until the 1960s when various Macrocystis research programmes began in California, and later in British Columbia, Chile, México, and elsewhere. Since that time, several books and reviews and hundreds of research papers have appeared in both the primary and secondary literature, primarily emphasising the physical and biotic factors that regulate Macrocystis distribution and abundance, recruitment, reproductive strategies and the structure and organisation of Macrocystis communities (see reviews by North & Hubbs 1968, North 1971, 1994, Dayton 1985a, Foster & Schiel 1985, North et al. 1986, Vásquez & Buschmann 1997). This review synthesises this rich literature into a global perspective of Macrocystis ecology and such a review is timely for three reasons. First, the last review of Macrocystis ecology was done by North (1994) and thoroughly covered the literature until 1990, yet there has been significant progress on many aspects of Macrocystis ecology since that time. Second, during the last 15–20 yr the general focus of Macrocystis research (and that of kelps in general) has shifted from descriptive and experimental studies of local Macrocystis distribution, abundance and population and community structure (e.g., competition and herbivory) to comprehensive investigations of Macrocystis life history, dispersal, recruitment, physiology and broad-scale variability in population and community processes. Finally, previous reviews of Macrocystis ecology have been from an inherently regional perspective (e.g., California or Chile) and there is currently no truly global synthesis. This last aspect is of great concern because it effectively partitions kelp forest researchers into provincial programmes and limits cross-fertilisation of ideas. Such a limitation is compounded by the great worldwide scientific and economic importance of this genus, the acclimatisation of Macrocystis to regional environments, and the recent finding that gene flow occurs among the most geographically distant regions over ecological timescales (Coyer et al. 2001). Therefore, the goal here is not to review the existing Macrocystis literature in its entirety, but rather to (1) focus on progress made during the last 15 yr, (2) discuss the achievements of Macrocystis research programmes worldwide and (3) identify deficiencies in the understanding of Macrocystis ecology that warrant future investigation. 40

Washington Oregon

41 Chile

Peru

Falkland Is.

Argentina

South Georgia Is.

Gough Is.

Crozet Is. Prince Edward Is.

Tristan de Cunha Is.

South Africa

Heard Is.

Kerguelen Is.

Amsterdam/St.Paul Is.

New Zealand Bounty Is.

Chatham Is.

Antipodes Is. Campbell Is.

Tasmania Auckland Is.

South Australia

Figure 1 Global distribution of the giant kelp Macrocystis. Locations are given for distinct Macrocystis mainland and island populations determined directly from citations herein.

Baja California Mexico

California

British Columbia

Alaska


GLOBAL ECOLOGY OF THE GIANT KELP MACROCYSTIS: FROM ECOTYPES TO ECOSYSTEMS



In particular, it is now recognised that great variability exists in Macrocystis morphology, physiology, population dynamics and community interactions at the global scale and it is considered that such ecotypic variability is key to understanding the role of Macrocystis in kelp systems worldwide.

Organismal biology of Macrocystis Most of the biological processes that ultimately prove to be important in regulating the dynamics and structure of Macrocystis populations and communities (e.g., morphological complexity, photosynthesis, growth, reproductive output, gene flow) operate primarily at the scale of individual organisms. The standard means of studying Macrocystis organismal biology continues to be through laboratory studies. Clearly, laboratory studies allow researchers to address various processes under controlled environmental conditions, but in many cases the reliance on laboratory studies has been due to technical limitations in collecting organismal data in situ. Various technological advances since the 1960s (most occurring in the last two decades), however, have resulted in a surge of studies of Macrocystis evolutionary history, distribution, life history, growth, productivity and reproduction.

Evolutionary history The order Laminariales has traditionally included five families (Chordaceae, Pseudochordaceae, Alariaceae, Laminariaceae, Lessoniaceae) but various ultrastructural and molecular data suggest that subordinal classification (i.e., families, genera, and species) is in need of significant revision (Druehl et al. 1997, Yoon et al. 2001, Lane et al. 2006). For example, the Chordaceae and Pseudochordaceae should not be included in the Laminariales (Saunders & Druehl 1992, 1993, Druehl et al. 1997) and a new family has been proposed (Costariaceae; Lane et al. 2006). The order is presumed to have originated in the northeast Pacific (Estes & Steinberg 1988, Lüning 1990) and molecular studies have estimated the date of origin to be between 15 and 35 million yr ago (Saunders & Druehl 1992). Within the order, the genus Macrocystis was formerly assigned to the family Lessoniaceae (including Lessonia, Lessoniopsis, Dictyoneurum, Dictyoneuropsis, Nereocystis, Postelsia and Pelagophycus; Setchell & Gardner 1925), which was considered paraphyletic to the Laminariaceae (Druehl et al. 1997, Yoon et al. 2001). Recent molecular studies, however, have found that Lessonia, Lessoniopsis, Dictyoneurum and Dictyoneuropsis are actually in phylogenetic clades that do not include Macrocystis, and that Macrocystis, Nereocystis, Postelsia and Pelagophycus group together in a derived clade that is nested well within the Laminariaceae (Lane et al. 2006), with Pelagophycus porra being the most closely related taxon to Macrocystis. Species classification within the genus Macrocystis was originally based on blade morphology yielding over 17 species (see review by North (1971)). Blade morphology was then considered a plastic trait strongly affected by environmental conditions and subsequently all 17 Macrocystis species were synonymised with Macrocystis pyrifera (Hooker 1847). Macrocystis species were later described based on holdfast morphology ultimately leading to the current recognition of three species: M. pyrifera (conical holdfast; Figure 2A), M. integrifolia (rhizomatous holdfast; Figure 2B), and M. angustifolia (mounding rhizomatous holdfast) (Howe 1914, Setchell 1932, Womersley 1954, Neushul 1971). The fourth currently recognised species, M. laevis, was described by Hay (1986), again based on blade morphology (M. laevis has smooth fleshy blades and a M. pyriferatype conical holdfast). Four lines of evidence, however, suggest that this current classification of Macrocystis is also in need of revision: (1) M. pyrifera, M. integrifolia and M. angustifolia are interfertile (Lewis et al. 1986, Lewis & Neushul 1994; interfertility with M. laevis has not been 42



A

B

C

D

Figure 2 Macrocystis holdfast morphologies and sporophyte spacing. (A) Holdfast of pyrifera-form sporophyte from La Jolla, southern California. (Published with permission of Scott Rumsey.) (B) Holdfast of integrifolia-form sporophyte from Huasco, northern Chile. (Photograph by Michael Graham.) (C) Vertical structure of pyrifera-form population from San Clemente Island (15 m depth), southern California; note average sporophyte spacing is 3–7 m. (Published with permission of Enric Sala.) (D) Vertical structure of angustifoliaform population from Soberanes Point (3 m depth), central California; note average sporophyte spacing is 10–50 cm. (Published with permission of Aurora Alifano.)

tested); (2) intermediate morphologies have been observed in the field (Setchell 1932, Neushul 1959, Womersley 1987, Brostoff 1988); (3) in addition to blade morphology (Hurd et al. 1997), holdfast morphology is phenotypically plastic (Setchell 1932, M.H. Graham, unpublished data); and most importantly, (4) patterns of genetic relatedness among all four species are not in concordance with current morphological classification (Coyer et al. 2001). This evidence strongly supports the recognition of the genus Macrocystis as a single morphologically plastic species, with global populations linked by non-trivial gene flow. For the purpose of this review, therefore, the four currently recognised species are referred to simply as giant kelp, Macrocystis. Biogeographic studies of extant kelp in the north Pacific suggest that the bi-hemispheric (antitropical) global distribution of Macrocystis developed as the genus arose in the Northern Hemisphere and subsequently colonised the Southern Hemisphere (North 1971, Nicholson 1978, Estes & Steinberg 1988, Lüning 1990, Lindberg 1991). Alternatively, North (1971) and Chin et al. (1991) proposed a Southern Hemisphere origin of the genus, the latter via vicariant processes that have been questioned (Lindberg 1991). Recently, Coyer et al. (2001) studied the global phylogeography of Macrocystis using recombinant DNA internal transcribed spacer (ITS1 and ITS2) regions. In addition to suggesting that the morphological species description of M. pyrifera, M. integrifolia, M. angustifolia and M. laevis has no systematic support, Coyer et al. (2001) described a wellresolved phylogeographic pattern in which Southern Hemisphere Macrocystis populations nested within Northern Hemisphere populations, linked by Macrocystis populations on the Baja California Peninsula, Mexico. This pattern, and the greater genetic diversity among Macrocystis populations in the Northern Hemisphere (within-region sequence divergences 1.7% and 1.2% for ITS1 and ITS2, respectively) relative to their Southern Hemisphere counterparts (within-region sequence

43



divergences 0.8% and 0.6% for ITS1 and ITS2, respectively), supports a northern origin of the genus with subsequent range expansion to include the Southern Hemisphere (Coyer et al. 2001); Coyer et al. (2001) suggested that gene flow across the equator may have occurred as recently as 10,000 yr ago. Despite such progress, however, many questions remain regarding the evolutionary history of Macrocystis. Most importantly, how can this single, globally distributed species maintain gene flow throughout its range, yet at a regional scale exhibit relatively high geographic uniformity in such seemingly important characters as blade and holdfast morphology (i.e., ecotypes or forms)? The data of Coyer et al. (2001) suggest that simple founder effects may have resulted in the unique morphologies of the laevis form at the Prince Edward Islands (including Marion Island) and angustifolia form in Australia. The smooth-bladed laevis form has been found occasionally at the Falkland Islands (van Tüssenbroek 1989a) and a recent description from Chiloé Island, Chile (Aguilar-Rosas et al. 2003), is probably a misidentification of sporophylls as vegetative blades (Gutierrez et al. 2006). Still, despite the apparently high gene flow and morphological plasticity, the distinct forms with distinct ecologies can dominate different habitats often adjacent to each other (e.g., integrifolia form in shallow water vs. pyrifera form in deep water). The identification of which Macrocystis form is present within a region will aid in the understanding of the region’s ecology (see ‘Population’ section, p. 54). In this context, it is hypothesised that the great plasticity in Macrocystis form and function may, in fact, be an adaptive trait resulting in its great global ecological success. Studies testing this hypothesis will require a better understanding of the nature of Macrocystis morphological plasticity, including biomechanics, structural biochemistry and quantitative genetics studies of genes regulating Macrocystis form.

Distribution Macrocystis distributional patterns have been well described (especially in the Northern Hemisphere) due primarily to the large stature of Macrocystis sporophytes and ability to sense their surface canopies remotely from aircraft or satellites (Jensen et al. 1980, Hernández-Carmona et al. 1989a,b, 1991, Augenstein et al. 1991, Belsher and Mouchot 1992, Deysher 1993, North et al. 1993, Donnellan 2004). Macrocystis typically grows on rocky substrata between the low intertidal and ~25 m depth (Figure 3; Rigg 1913, Crandall 1915, Baardseth 1941, Papenfuss 1942, Scagel 1947, Guiler 1952, 1960, Cribb 1954, Chamberlain 1965, Neushul 1971, Foster & Schiel 1985, Westermeier & Möller 1990, van Tüssenbroek 1993, Schiel et al. 1995, Graham 1997, Spalding et al. 2003, Vega et al. 2005) and is distributed in the northeast Pacific from Alaska to México, along the west and southeast coasts of South America from Perú to Argentina, in isolated regions of South Africa, Australia and New Zealand and around most of the sub-Antarctic islands to 60°S (Figure 1; Crandall 1915, Baardseth 1941, Cribb 1954, Papenfuss 1964, Chamberlain 1965, Neushul 1971, Hay 1986, Stegenga et al. 1997). In unique circumstances, sexually reproducing populations can exist in deep water (50–60 m; Neushul 1971 (Argentina), Perissinotto & McQuaid 1992 (Prince Edward Islands)), in sandy habitats (Neushul 1971) and unattached populations that reproduce vegetatively can exist in the water column (North 1971) or shallow basins (Moore 1943, Gerard & Kirkmann 1984, van Tüssenbroek 1989b). High latitudinal limits appear to be set by increased wave action (Foster & Schiel 1985, Graham 1997) and decreased insolation (Arnold & Manley 1985, Jackson 1987), whereas low latitudinal limits appear to be set by low nutrients associated with warmer (non-upwelling) waters (Ladah et al. 1999, Hernández-Carmona et al. 2000, 2001, Edwards 2004) or competition with warm-tolerant species (e.g., Eisenia arborea on the Baja California Peninsula, Mexico; Edwards & Hernández-Carmona 2005). The upper shallow limits of Macrocystis populations are ultimately regulated by the increased desiccation and high ultraviolet and/or photosynthetically active radiation (PAR) of the intertidal zone (Graham 1996, Huovinen et al. 2000, Swanson & Druehl 2000), 44

45

E

B

F

C

Figure 3 Photographs of various Macrocystis populations. (A) Infrared aerial canopy photo of subtidal pyrifera-form population at La Jolla, southern California. (Published with permission of Larry Deysher/Ocean Imaging.) (B) Shallow subtidal pyrifera-form population at Mar Brava, central Chile. (Photograph by Michael Graham.) (C) Subtidal pyrifera-form population at Nightingale Island near Tristan da Cunha Island, South Atlantic Ocean. (Published with permission of Juanita Brock.) (D) Intertidal integrifolia-form population at Van Damme State Park, northern California. (Photograph by Michael Graham.) (E) Intertidal integrifolia-form population at Strait of Juan de Fuca, Washington. (Photograph by Michael Graham.) (F) Intertidal integrifolia-form population at Huasco, northern Chile. (Photograph by Michael Graham.)

D

A





although wave activity, grazing and competition with other macroalgae in shallow subtidal areas can also be important (Santelices & Ojeda 1984a, Foster & Schiel 1992, Graham 1997). At local scales, decreased availability of light and rocky substratum, and occasionally sea urchin grazing, appear to set the lower off-shore limits of Macrocystis populations (Pearse & Hines 1979, Lüning 1990, Spalding et al. 2003, Vega et al. 2005). Finally, within these upper and lower limits, the lateral distribution of Macrocystis populations typically corresponds with abrupt changes in bathymetry or substratum composition (e.g., sand channels or harbour mouths; North & Hubbs 1968, Dayton et al. 1992, Kinlan et al. 2005). There is an interesting pattern within the global distribution of Macrocystis whereby different regions may have large Macrocystis populations of one morphological form or another (Neushul 1971, Womersley 1987). For example, the integrifolia and angustifolia forms of Macrocystis are generally found in shallow waters (low intertidal zone to 10 m depth), whereas the pyrifera form is generally found in intermediate-to-deep waters (4–70 m depth) (Table 1). In the Northern Hemisphere, the integrifolia form is most commonly observed at higher latitudes north of San Francisco Bay with scattered populations found as far south as southern California (Abbott & Hollenberg 1976, M.H. Graham, personal observations), whereas the pyrifera form is most common at lower latitudes south of San Francisco Bay with scattered populations found as far north as southeast Alaska (Gabrielson et al. 2000). In South America, the integrifolia and pyrifera forms also appear to occupy shallow and deep habitats, respectively (Howe 1914, Neushul 1971). Latitudinally, however, the Southern Hemisphere Macrocystis distribution is opposite that of the Northern Hemisphere: the integrifolia form is generally found at lower latitudes, restricted to Perú México, and northern Chile (Howe 1914, Neushul 1971), whereas the pyrifera form dominates the higher latitudes of central and southern Chile (and Argentina; Barrales & Lobban 1975), but can also be found far north in Perú (Howe 1914, Neushul 1971). The pyrifera form also appears to be Table 1 Maximum depths of worldwide populations of Macrocystis ecotypes Macrocystis form

Location

angustifolia

South Australia South Africa British Columbia Northern Chile Perú Southern Chile Tasmania New Zealand St. Paul/Amsterdam Is. Crozet Is. Falkland Is. South Georgia Is. Southern California Central California Tristan da Cunha Is. Baja California Perú Southern Argentina Kerguelen Is. Gough Is. Prince Edward Is.

integrifolia

pyrifera

laevis

Depth (m) 6 8 10 8, 14 20 10 15 16 20 25 25 25 30 30 30 40 40 55 40 55 68

* Depths interpreted by Perissinotto & McQuaid (1992).

46

Reference Womersley 1954 Isaac 1937 Druehl 1978 Neushul 1971, Vega et al. 2005 Juhl-Noodt 1958* Dayton et al. 1973 Cribb 1954 Hay 1990 Delépine 1966* Delépine 1966* Powell 1981 Skottsberg 1941 Neushul & Haxo 1963 Spalding et al. 2003 Baardseth 1941 North 1971 Juhl-Noodt 1958* Neushul 1971 Grua 1964* Chamberlain 1965 Perissinotto & McQuaid 1992



most common where Macrocystis is found elsewhere in the Southern Hemisphere (e.g., Tasmania, New Zealand, various sub-Antarctic islands), except in South Australia and South Africa where the angustifolia form is common (Cribb 1954, Womersley 1954, 1987, Hay 1986, Stegenga et al. 1997). Jackson’s (1987) analyses suggested that high latitude Macrocystis sporophytes would be light limited in subtidal waters, forcing a shift in distribution to shallower water above 53° latitude. This may explain the Northern Hemisphere distributional pattern, but cannot explain why shallow-water Macrocystis is the most common form in northern Chile. Furthermore, exceptions to these patterns clearly exist. For example, pyrifera-form individuals can be found in the intertidal zone (e.g., Guiler 1952, 1960 (Tasmania), Chamberlain 1965 (Gough Island), Westermeier and Möller 1990 (southern Chile), van Tüssenbroek 1993 (Falkland Islands)), sometimes even side by side with integrifoliaform individuals (M.H. Graham, personal observations in California; J.A. Vásquez, personal observations in northern Chile). Intermediate morphologies similar to the angustifolia form of South Australia-South Africa can also be observed at intermediate depths (2–6 m) between adjacent pyrifera-form and integrifolia-form populations in central California (M.H. Graham, personal observations). Still, these global distribution patterns support the general consideration of the integrifolia and angustifolia forms as having more shallow-water affinities than the pyrifera form. Another interesting global distributional pattern is the apparent restriction of large Macrocystis forests (>1 km2) to the southwest coast of North America (Point Conception in southern California to Punta Eugenia in Baja California, Mexico; Hernández-Carmona et al. 1991, North et al. 1993), although Macrocystis forests on most of the sub-Antarctic islands have not been explored. The southwest coast of North America has broad shallow-sloping subtidal rocky platforms to support wide Macrocystis populations (up to 1 km width), whereas the regions north to Alaska and south to Patagonia have steep shores and typically support very narrow Macrocystis populations (1 µM. Such critical irradiance, temperature and nutrient thresholds were further supported by field experiments (Deysher & Dean 1986b). Although these studies did not provide data amenable to the development of probability density functions for predicting Macrocystis recruitment success as a function of variable environmental conditions, the research was vital to the development of the concept of temporal ‘recruitment windows’, during 48



which environmental factors exceeded minimum levels for successful gametogenesis and fertilisation. Deysher & Dean (1986a) also found that the growth of embryonic sporophytes to macroscopic size was inhibited at low PAR and nitrate concentrations, but these PAR levels were higher than the threshold for gametogenesis and fertilisation. This suggests that the growth of embryonic sporophytes to macroscopic size may be a stronger bottleneck in the Macrocystis life history than gametogenesis and fertilisation. The photosynthesis studies of Fain & Murray (1982) similarly identified differences in physiology between Macrocystis gametophytes and embryonic sporophytes. The process leading to Macrocystis recruitment from gametophytes can thus be divided into two functionally different stages: (1) sporophyte production (gametogenesis and fertilisation, with relatively lower light requirements) and (2) growth of sporophytes to macroscopic size (with relatively higher light requirements). It follows that the timing and success of Macrocystis recruitment will depend on both the duration of each stage and whether such durations can be extended to allow for delayed recruitment (Figure 4) similar to the concept of seed banks for terrestrial plants (Hoffmann & Santelices 1991). Laboratory and field studies indicate that the sporophyte production stage is relatively short (1–2 months) and rigid in its duration, suggesting limited potential for delayed Macrocystis recruitment via gametophytes. In California, female Macrocystis gametophytes appear to have an initial competency period of 7–10 days prior to gametogenesis (North 1987) and lose fertility after ~30 days (Deysher & Dean 1984, Kinlan et al. 2003), whereas in Chile, laboratory culture studies under ample light and nutrient conditions suggest that Macrocystis female gametophytes may remain fertile for up to 75 days (Muñoz et al. 2004). However, Macrocystis gametophytes can apparently survive indefinitely under ‘unnatural’ artificial light-quality conditions (i.e., red light only; Lüning & Neushul 1978). In California, unfertilised female gametophytes older than ~30 days have limited potential for fertilisation (Deysher & Dean 1986b) and thus recruitment, which was supported by the laboratory studies of Kinlan et al. (2003). As indicated by Kinlan et al. (2003), however, the necessary studies have not been done to determine whether this lack of fertilisation success is due to senescence of female gametophytes or of their male counterparts. Also, it has been demonstrated that zoospore swimming ability is correlated with germination success (Amsler & Neushul 1990) and a similar mechanism may affect fertilisation of oogonia by antherozoids. The demonstration of a shorter life-span (period of fertility) for Macrocystis male gametophytes relative to females would suggest the potential for cross-fertilisation among different zoospore settlements via perennial females. Another well-known aspect of sporophyte production is the minimum density of settled zoospores necessary for recruitment. Specifically, the reliance of kelps on the presence of lamoxirene as a trigger for antherozoid release (Maier et al. 1987, 2001) and the dilution of this sexual pheromone over short distances from the oogonia inherently require sufficient zoospore settlement densities (and survivorship to maturation) to ensure that males and females are close enough for fertilisation to be successful. Such ‘critical settlement densities’ were demonstrated in a series of laboratory and field experiments by Reed and his colleagues (Reed 1990, Reed et al. 1991). Specifically, Reed et al. (1991) identified 1 settled zoospore mm−2 (vs. 0.1 or 10 settled zoospores mm–2) as the minimum Macrocystis (and Pterygophora) zoospore settlement density above which fertilisation and sporophyte production could be expected. These experiments focused on recruitment from single zoospore settlement cohorts and cross-fertilisation among different zoospore settlements may result in fertilisation even if cohort settlement densities are 1 settled zoospore to yield an adult sporophyte. Reed (1990) also demonstrated that speciesdependent female maturation rates combined with species-independent pheromone activity might 49



result in chemically mediated competition among microscopic stages of kelp species, although this was only suggested for Macrocystis and Pterygophora in southern California. In addition to producing valuable life-history data, these studies clearly demonstrated the utility of combining laboratory and field experiments of kelp recruitment and resulted in a surge in studies of the ecology of kelp microscopic stages. Nevertheless, several key issues regarding the Macrocystis life history remain to be resolved. Most importantly, Macrocystis microscopic life-history stages have not been observed in the field. Microphotometric techniques have recently been developed for identifying Macrocystis zoospores based on species-specific zoospore absorption spectra (Graham 1999, Graham & Mitchell 1999). Subsequent determination of Macrocystis zoospore concentrations from in situ plankton samples led to direct studies of Macrocystis zoospore planktonic processes (e.g., Graham 2003). However, upon settlement, Macrocystis zoospores germinate into gametophytes of variable cell number and pigment concentration, negating the use of microphotometric techniques for studying postsettlement processes (Graham 2000). Fluorescently labelled monoclonal antibodies have been developed for distinguishing between Macrocystis and Pterygophora gametophytes based on cell surface antigens (Hempel et al. 1989, Eardley et al. 1990). However, the effectiveness of these tags diminishes when applied to field samples, in which kelp cells are universally coated with bacteria (D.C. Reed, personal communication). Additionally, although Kinlan et al. (2003) observed plasticity in growth of laboratory-cultured Macrocystis embryonic sporophytes under realistic environmental conditions (light and nutrients), and thus the potential for arrested development in this stage, their experiments provided no evidence of arrested development of gametophytes. This study demonstrated (1) that delayed recruitment of Macrocystis post-settlement stages is possible and (2) the general lack of understanding of the physiological processes that regulate the growth, maturation and senescence of Macrocystis microscopic stages. For example, it is considered that kelp female gametophytes living under adequate environmental conditions will have only one or very few cells, one oogonium per gametophyte, and become reproductive in the shortest period possible (e.g., Lüning & Neushul 1978, Kain 1979). In the absence of light and nutrients, female gametophytes are typically sterile and multicellular (e.g., Lüning & Neushul 1978, Kain 1979, Hoffmann & Santelices 1982, Hoffmann et al. 1984, Avila et al. 1985, Reed et al. 1991), suggesting a trade-off or antagonistic relationship between gametophyte growth and fertility. In some Chilean populations, however, female Macrocystis gametophytes grown under standard laboratory conditions (1) were multicellular, (2) produced multiple viable oogonia per gametophyte, (3) often resulted in numerous sporophytes per gametophyte and (4) took longer to mature than Californian populations (Muñoz et al. 2004). These results must be validated by additional laboratory and field studies but they did demonstrate the highly plastic physiology of Macrocystis life-history stages. Our lack of understanding of variability in the biology of Macrocystis microscopic stages, especially at the global level, is an important constraint on future progress in Macrocystis population dynamics (see ‘Population’ section, p. 54).

Growth, productivity and reproduction Recruitment processes are the main determinant of when and where Macrocystis sporophytes might occur, yet it is the survival and growth of established sporophytes that constrain sporophyte size, self-thinning, population cycles and the primary productivity and canopy structure that ultimately provide energy and habitat for Macrocystis communities. The maximum age of Macrocystis sporophytes is unknown. Individual fronds generally senesce after 6–8 months (North 1994) although van Tüssenbroek (1989c) observed maximum frond survival of 1 yr and Macrocystis sporophytes can produce new fronds from apical meristems (frond initials) retained above the holdfasts near the sporophylls (Lobban 1978a,b, van Tüssenbroek 1989c, North 50



1994). As such, the sporophytes may survive as long as they remain attached to the substratum and environmental conditions are adequate for growth. In some regions of central California and Argentina, most Macrocystis sporophytes die within a year due to high wave activity (Barrales & Lobban 1975, Graham et al. 1997), whereas in southern California, sporophytes can live up to 4–7 yr (Rosenthal et al. 1974, Dayton et al. 1984, 1999), a life-span that coincides with the periodicity of the El Niño Southern Oscillation (ENSO); in southern Chile, the life-span of Macrocystis sporophytes often exceeds 2 yr (Santelices & Ojeda 1984b, Westermeier & Möller 1990). Interestingly, the only Macrocystis populations known to recruit and senesce on an annual cycle occur in the protected waters around 42°S in Chilean fjords (Buschmann et al. 2004a). The life-span of vegetatively reproducing Macrocystis sporophytes (e.g., angustifolia and integrifolia forms) has never been determined in the field, although integrifolia-form sporophytes have been shown to survive very high levels of rhizome fragmentation (Druehl & Kemp 1982, Graham 1996) and cultivated sporophytes can live 2–3 yr (Druehl & Wheeler 1986). In any case, the life-span of Macrocystis sporophytes appears to be far less than that of other perennial kelp genera (Reed et al. 1996, Schiel & Foster 2006), for example, Pterygophora and Eisenia, which can live for 10+ yr (Dayton et al. 1984). The relatively high turnover of Macrocystis sporophytes is probably due to their massive size (up to 400 fronds per pyrifera-form sporophyte; North 1994) and the almost strict reliance of sporophyte growth and productivity on the biomass of the surface canopy (Reed 1987, North 1994, Graham 2002). Shallow-water Macrocystis sporophytes typically have lower frond numbers than deeper sporophytes (North 1994). Numerous studies have demonstrated high Macrocystis frond productivity rates with estimates of 2–15 g fixed carbon m−2 day−1 in the Northern Hemisphere (reviewed by North 1994), and values that vary between 7 and 11 g C m−2 day−1 in the southern Indian Ocean (Attwood et al. 1991). Delille et al. (2000) also observed a significant ‘draw-down’ of pCO2 when off-shore water entered a dense Macrocystis bed at the Kerguelen Islands, suggesting that the productivity of Macrocystis fronds was high enough to decrease inorganic carbon concentrations in the water column. Furthermore, Schmitz & Lobban (1976) determined that Macrocystis sporophytes can translocate photosynthates from production sources in the surface canopy to energy sinks (meristems, holdfasts, sporophylls) at rates of 55 to 570 mm h−1; the canopy typically represents the greatest contribution to total sporophyte biomass (Nyman et al. 1993, North 1994). Such high rates of productivity and translocation appear to be necessary to maintain sporophyte growth in the face of high metabolic demands (Jackson 1987) because, unlike other perennial kelp genera (e.g., Pterygophora), Macrocystis sporophytes have very limited nutrient and photosynthate storage capabilities (2 wk; Gerard 1982, Brown et al. 1997). The subsequent reliance on the surface canopy, and the vulnerability of surface canopy fronds to both physical and biological disturbance, results in considerable spatial and temporal variability in Macrocystis productivity potential, size structure and overall health. The linkage between Macrocystis sporophyte growth, productivity and biomass therefore results in a plastic response of sporophyte condition to temporal and spatial variability in resource availability (Kain 1982, Reed et al. 1996). The low storage capabilities are clearly disadvantageous during periods of suboptimal environmental conditions, such as occur seasonally in southern California (Zimmerman & Kremer 1986) and the inland waters of southern Chile (Buschmann et al. 2004a). Again, other perennial kelp genera either possess greater storage capabilities or exhibit seasonally offset periods of growth and photosynthesis in order to weather periods of low resource availability (e.g., light or nutrients; Chapman & Craigie 1977, Gerard & Mann 1979, Dunton & Jodwalis 1988, Dunton 1990). At high latitudes, like British Columbia, southeast Alaska, and the Kerguelen and Falkland Islands, Macrocystis sporophyte growth follows distinctly seasonal patterns in insolation, with frond elongation ranging from 2 to 4.7 cm day−1 during the summer maximum (Lobban 1978b, Asensi et al. 1981, Druehl & Wheeler 1986, Wheeler & Druehl 1986, Jackson 51



1987, van Tüssenbroek 1989d). At lower latitudes, like California, distinct seasonal growth patterns due to variability in insolation were not apparent (North 1971, Wheeler & North 1981, Jackson 1987, Gonzalez-Fragoso et al. 1991, Hernández-Carmona 1996). Instead, Zimmerman & Kremer (1986) described seasonal frond growth rates that corresponded with variability in ambient nutrient concentrations (nitrate), in which frond growth was maximised during winter-spring (12–14 cm day−1; upwelling periods) and minimised during summer-fall (6–10 cm day−1; non-upwelling periods). In New Zealand, minimum Macrocystis frond growth rates also occurred during summer, but were relatively high throughout the remainder of the year (Brown et al. 1997), whereas in northern Chile frond growth rates of 5–10 cm day−1 were observed with no seasonal variability (Vega et al. 2005). In many regions, light and nutrients can be present well above limiting levels throughout the year (e.g., central California or central Chile) thereby permitting continuously high Macrocystis sporophyte productivity (Jackson 1987). The reliance of Macrocystis sporophyte growth and productivity on the biomass and health of the canopy also helps to explain much of the sensitivity of Macrocystis to ENSOs, relative to that of other kelp genera (Dayton et al. 1999). There is a strong inverse relationship between water column nitrate concentrations and water temperature (Zimmerman & Robertson 1985, Tegner et al. 1996, 1997, Dayton et al. 1999, Hernández-Carmona et al. 2001). Kelp growth becomes nutrient limited below approximately 1 µM nitrate, which typically occurs in southern California when water temperatures rise above 16°C (Jackson 1977, Zimmerman & Robertson 1985, Dayton et al. 1999); the same threshold appears to occur around 18°C in Baja California, Mexico (HernándezCarmona et al. 2001). During ENSOs, depression of the thermocline shuts down nutrient replenishment via coastal upwelling and decreases the propagation of nutrients via internal waves (Jackson 1977, Zimmerman & Robertson 1985, Tegner et al. 1996, 1997). Due to its limited nutrient storage capabilities, Macrocystis canopy biomass begins to deteriorate when tissue nitrogen drops below 1.1% dry weight (Gerard 1982). When frond losses exceed frond initiation, the biomass necessary to sustain meristems is lost and the sporophytes die. Sporophyte mortality was 100% in many Macrocystis forests in southern and Baja California following the 1983 and 1997 ENSOs (Dayton et al. 1984, 1992, 1999, Tegner & Dayton 1987, Dayton & Tegner 1989, Hernández-Carmona et al. 1991, Ladah et al. 1999, Edwards 2004), although sporophytes may find refuge in deep water (Ladah & Zertuche-Gonzalez 2004) or within the benthic boundary layer (Schroeter et al. 1995). Finally, during ENSOs, regulatory control over growth of juvenile Macrocystis sporophytes shifts from light inhibition under Macrocystis surface canopies (Dean & Jacobsen 1984) to nutrient limitation (Dean & Jacobsen 1986). Extensive plasticity in sporophyte growth is by no means restricted to Macrocystis adults. Due to the high temporal variability in sporophyte growth potential and the striking differences in biomass among small and large Macrocystis sporophytes, the transition among different size classes can also be delayed in time similar to the arrested development described above for embryonic sporophytes. Santelices & Ojeda (1984a) and Graham et al. (1997) observed that Macrocystis juveniles could survive for many months under adult canopies, growing rapidly to adult size when adult densities decreased and light became available. Presumably, light levels under the canopy were adequate to meet the metabolic demands of the juveniles, but inadequate to sustain growth (Dean & Jacobsen 1984). It is unknown, however, how long juveniles or subadults can survive such conditions. Another important feature of Macrocystis growth potential is that frond initiation is indeterminate because sporophytes can tolerate sublethal biomass loss (loss of fronds) as long as meristems are present and abiotic conditions are conducive to survival (North 1994). Subsequently, sporophyte age is decoupled from sporophyte size, which can be advantageous for both young and old individuals, but disadvantageous to researchers trying to use size as a proxy of age (Santelices & Ojeda 1984b). Graham (1997) found that large Macrocystis sporophytes living in the surf zone suffered greater mortality due to wave action than those that survived sublethal loss 52



of canopy biomass, which presumably decreased overall sporophyte drag and the likelihood of detachment by waves. Finally, it has been demonstrated that the response of Macrocystis juvenile growth to variable nutrient concentrations is under genotypic control (Kopczak et al. 1991), resulting in broad latitudinal variability in sporophyte growth and recruitment potential. Again, it is interesting that such genotypic variation can occur in spite of non-trivial gene flow among Macrocystis populations (Coyer et al. 2001). The reliance of sporophyte growth on surface canopy biomass also constrains reproductive output. Unlike most other kelp genera, Macrocystis sporophyll and sorus production can occur continuously given adequate translocation of photosynthates from the surface canopy (Neushul 1963, McPeak 1981, Reed 1987, Dayton et al. 1999, Graham 2002, Buschmann et al. 2006). The number of sporophylls per fertile Macrocystis sporophyte varies from 1 to 100+ (Lobban 1978a, Reed 1987, Reed et al. 1997, Buschmann et al. 2004a, 2006), although sporophyll growth rates have yet to be determined. In Macrocystis, two processes lead to turnover of reproductive material: growth of sporophylls and production of sori on the sporophylls (Neushul 1963). Both processes decrease in magnitude following either natural or experimental loss of canopy biomass (Reed 1987, Graham 2002), although the cessation of sorus production appears to be more sensitive than sporophyll growth to biomass loss and can result in complete sporophyte sterility within 9 days of disturbance to the canopy (Graham 2002). It is unknown whether sublethal biomass loss also affects the quantity or quality of zoospores in sori or the timing of their ultimate release into the water column. Due to the continuous reliance of Macrocystis reproduction on canopy biomass, however, variability in environmental factors can also greatly affect reproductive output. Reed et al. (1996) demonstrated that nitrogen content of Macrocystis zoospores varied as a function of in situ water temperature (and presumably water column nutrient concentrations) and nitrogen content of adults, whereas the nitrogen content of Pterygophora zoospores remained relatively constant. Reed et al. (1996) argued that the ability of Macrocystis sporophytes to respond to favourable environmental conditions allowed them to be reproductively successful despite their relatively short life-span. Again, such plasticity in reproductive timing can be adaptive, especially given the apparently low cost of reproduction in kelps (DeWreede & Klinger 1988, Pfister 1992). For example, Macrocystis sporophytes living in waveexposed locations in southern Chile reproduce year-round and produce high numbers of sporophylls, whereas Macrocystis sporophytes living in nearby wave-protected populations are annuals, have increased zoospore production per sorus area and are fertile for only a few months, presumably to ensure successful zoospore settlement and fertilisation prior to the disappearance of adult plants every autumn (Buschmann et al. 2004a, 2006). Overall, Macrocystis sporophyte growth, productivity and reproduction are very responsive to variability in environmental conditions. This response differs from that of most known kelps and other algae (see review by Santelices 1990) and is probably essential to the success of Macrocystis as a competitive dominant throughout much of its global distribution. What remains to be determined, however, is how this variable physiology is expressed among the different morphological forms of Macrocystis and across the variety of habitats in which Macrocystis populations are present. For example, the integrifolia and pyrifera forms inhabit low intertidal and deeper subtidal environments, respectively, which differ strikingly in factors known to regulate Macrocystis growth, productivity and reproduction (e.g., water motion, water quality and light availability). Consequently, it is expected that these two forms will respond differently to environmental perturbations (e.g., van Tüssenbroek 1989c,e), with potentially significant consequences at the population and community levels. This scenario is further complicated by the vegetative growth capabilities of the integrifolia form, absent in the pyrifera form, because the relative contribution of vegetative growth to sexual reproduction in maintaining integrifolia-form giant kelp populations is unknown. Furthermore, kelp physiological studies presently focus on measurements of physiological processes for 53



specific structures (e.g., photosynthesis, growth, or nutrient uptake of excised laminae), and few have integrated these processes across entire sporophyte thalli (but see the translocation studies of Schmitz & Lobban (1976)). For example, translocation elements (sieve tubes and trumpet hyphae) run through the rhizomes of integrifolia-form sporophytes, which may spread over greater lengths of substratum than pyrifera-form holdfasts, potentially providing a physiological connectivity among fronds over the scale of metres. This limitation has inhibited the development of realistic carbon and nitrogen budgets for kelps and thus constrained our understanding of the physiology of entire sporophytes. This limitation is critical because it is at the level of individual sporophytes, not individual laminae, that mortality, growth and reproduction have consequences for population biology.

Population biology of Macrocystis Most of the work on Macrocystis population dynamics prior to 1990 focused on processes regulating seasonal-to-annual variability in adult sporophyte mortality (see review by North 1994), including competition (Reed & Foster 1984, Santelices & Ojeda 1984a), herbivory (Harris et al. 1984, Ebeling et al. 1985, Harrold & Reed 1985) and physical disturbance (Rosenthal et al. 1974, Dayton et al. 1984). Stimulated by the research of Reed and his colleagues (Reed 1990, Reed et al. 1988, 1991, 1992, 1996, 1997, 2004), a more population-based approach to Macrocystis biology and ecology has recently emerged in which studies have shifted to focus on reproduction, dispersal and recruitment and the consequences of these processes to the persistence of Macrocystis populations. Subsequently researchers have developed a more integrated view of Macrocystis population dynamics that unites variability in mortality agents with recruitment processes to provide a better understanding of local and global differences in Macrocystis population cycles.

Stage- and size-specific mortality Macrocystis populations do not exhibit unbounded growth (Dayton 1985a, Foster & Schiel 1985, North 1994). Although Macrocystis populations are probably never at equilibrium, Macrocystis populations often reach an apparent maximum in abundance or biomass per unit area (carrying capacity) that is determined by the availability of environmental resources (e.g., space, light and nutrients; Nisbet & Bence 1989, Burgman & Gerard 1990, Graham et al. 1997, Tegner et al. 1997). Furthermore, it has been well established that a variety of density-dependent and density-independent processes result in stage- and size-specific sporophyte mortality (reviewed by Schiel & Foster 2006) and retain Macrocystis at a population level below carrying capacity and initiate population cycling. Due to their large size and high drag, Macrocystis adult sporophytes are extremely vulnerable to removal by high water motion, and wave-induced sporophyte loss is considered the primary factor resulting in Macrocystis mortality (Foster 1982, Dayton et al. 1984, Seymour et al. 1989, Schiel et al. 1995, Graham 1997, Graham et al. 1997, Edwards 2004). The probability of a sporophyte being removed from the substratum by passing or breaking waves increases when the force (drag) experienced by the sporophyte due to water motion (related to both water velocity and the cross-sectional area of the sporophyte exposed to the flow; Seymour et al. 1989, Utter & Denny 1996) exceeds the attachment strength of the sporophyte holdfast (for whole sporophyte mortalities) or the breaking strength of individual fronds (for sublethal frond mortality; Utter & Denny 1996). High seasonal and year-to-year variability in wave intensity and sporophyte biomass therefore results in highly variable sporophyte mortality throughout the year. For example, in California, most sporophyte mortalities occur during the first large fall-winter storms (Zobell 1971, Gerard 1976, Graham et al. 1997), when adult biomass is high following long periods of low wave activity (spring to fall). It appears that sporophytes that survive these storms, but shed fronds and canopy 54



biomass, decrease their overall drag and increase the probability of surviving subsequent and often more severe storms (Graham et al. 1997). On the Chatham Islands, Macrocystis populations are only found at protected sites (Schiel et al. 1995) and never attain large sporophyte or population sizes. In southern California, uprooted sporophytes are often observed entangled with attached sporophytes, further increasing the attached sporophytes’ drag and probability of detachment (Rosenthal et al. 1974, Dayton et al. 1984) and resulting in a ‘snowball effect’ that can clear large swaths in the local population (Dayton et al. 1984). Such massive entanglements, however, appear to be rare in central California (Graham 1997), possibly due to more rapid transport of detached sporophytes out of and away from the local population. Increased sporophyte biomass, therefore, simultaneously increases both Macrocystis growth and reproductive potential (described in the Organismal biology section) and the probability of wave-induced mortality. This trade-off between fitness and survival is probably viable because of the temporal and spatial unpredictability in wave intensity experienced throughout the alga’s global distribution and its ability to survive and quickly recover from sublethal loss of biomass. Exceptions are the wave-protected annual Macrocystis populations in southern Chile in which there seems to be no trade-off between reproductive output and survival (Buschmann et al. 2006). In this case, synchronous growth, reproduction and senescence occur in the near absence of water motion. Despite the high temporal variability in wave-induced mortality, Macrocystis sporophytes exhibit distinct spatial patterns in survivorship. Wave-induced mortality of all size classes of adult sporophytes increases with both increasing wave exposure (Foster & Schiel 1985, Graham et al. 1997) and decreasing depth (Seymour et al. 1989, van Tüssenbroek 1989c, Dayton et al. 1992, Graham 1997). These patterns are primarily due to spatial variability in water motion because wave activity increases toward shallow water, the tips of rocky headlands and regions of high storm production (e.g., the relatively stable winter Aleutian low-pressure system in the Northern Hemisphere). Graham et al. (1997), however, also observed that Macrocystis holdfast growth decreased significantly along a gradient of increasing wave exposure, possibly due to greater disturbance to the Macrocystis surface canopy, which reduces translocation to haptera and thereby reduces holdfast growth (Barilotti et al. 1985, McCleneghan & Houk 1985). Thus, increased wave forces and decreased strengths of holdfast attachment can act in combination to decrease Macrocystis sporophyte survival; Graham et al. (1997) observed that Macrocystis sporophyte life-spans rarely exceeded 1 yr at their most wave-exposed sites. Although all of these described patterns may possibly exist for any Macrocystis life stage, the likelihood of wave-induced mortalities will be much lower for the smaller life stages due to both decreasing thallus size and decreasing water velocities within the benthic boundary layer. Additionally, other hydrographic factors can result in high sporophyte mortalities in relatively wave-protected regions (e.g., tidal surge, nutrient limitation, temperature and salinity stress; Buschmann et al. 2004a, 2006). Biological processes also clearly play a role in mortality of Macrocystis sporophytes. During sea urchin population outbreaks, sea urchin grazing of Macrocystis holdfasts can result in (1) detachment of adult sporophytes and their removal from the population (Dayton 1985a, Tegner et al. 1995a), (2) modification of sporophyte morphology (Vásquez & Buschmann 1997) and (3) removal of entire recruits and juvenile sporophytes (Dean et al. 1984, 1988, Buschmann et al. 2004b, Vásquez et al. 2006). Unlike some locations (e.g., the Aleutian Islands; Estes & Duggins 1995), widespread destruction of Californian and Chilean Macrocystis populations by sea urchin grazing is rare (Castilla & Moreno 1982, Foster & Schiel 1988, Steneck et al. 2002, Graham 2004). Still, sea urchin outbreaks can result in episodic deforestation of Macrocystis populations up to a scale of a few kilometres (Dayton 1985a). In healthy southern California systems, sea urchins can live in Macrocystis holdfasts and result in holdfast cavitation and thus a decrease in sporophyte attachment strength (Tegner et al. 1995a). Although episodic and small scale, the prevalence of holdfast cavitation by sea urchins increases with increasing sporophyte age, thereby increasing the vulnerability of 55



large, older sporophytes to wave-induced mortality (Tegner et al. 1995a). Infestations of Macrocystis sporophytes by epizoites and small herbivorous crustaceans (amphipods and isopods) have also been observed worldwide (North & Schaeffer 1964, Dayton 1985b). Most outbreaks of herbivorous crustaceans simply result in sublethal biomass loss (Graham 2002), which will effectively decrease sporophyte drag and thus possibly wave-induced mortality. Crustacean infestations can also occur in the holdfasts and result in increased mortality due to decreased sporophyte attachment strength (North & Schaeffer 1964, Ojeda & Santelices 1984). When carnivorous ‘picker’ fishes are absent from the water column in both California and Chile, outbreaks of epiphytic sessile invertebrates (bryozoans, kelp Pecten spp., spirorbids) often result (Bernstein & Jung 1979, Dayton 1985b), weighing down Macrocystis sporophyte canopies and either (1) increasing the likelihood of detachment due to water motion or (2) bringing surface canopy biomass into contact with grazing activities of benthic herbivores (Dayton 1985b). Although seemingly important, there are very few data concerning the importance of these processes in regulating Macrocystis mortality worldwide. Finally, although not a natural biological disturbance, human harvesting of Macrocystis canopies does not appear to have significant effects on sporophyte survival (Kimura & Foster 1984, Barilotti et al. 1985, Druehl & Breen 1986). Inter- and intraspecific competition for space and light are important in regulating the survival of Macrocystis microscopic stages (gametophytes and embryonic sporophytes) to macroscopic size (juveniles; less than tens of centimetres), and growth of Macrocystis juveniles to adult size (Schiel & Foster 2006). Smaller Macrocystis thalli are vulnerable to overgrowth by seaweeds and other kelps (Santelices & Ojeda 1984a, Vega et al. 2005), and even by conspecifics in monospecific stands (Schroeter et al. 1995, Graham et al. 1997). Intraspecific competition for space is likely to be most severe at the smaller size classes because critical zoospore settlement densities will result in high densities of microscopic embryonic sporophytes following fertilisation and the large size of adult Macrocystis holdfasts (up to 1 m diameter) necessitates that many recruits and juveniles will be smothered as nearby sporophytes grow in size. After Macrocystis sporophyte densities are initially thinned by competition for space, competition for light increases as sporophytes begin to grow to the water surface (Dean & Jacobsen 1984). Sporophytes that reach the surface will have enhanced photosynthetic rates and be able to translocate more photosynthates to basal meristems for new frond initiation (North 1994). As such, sporophytes that gain the competitive edge of a surface canopy may become even larger, increasing their likelihood of outcompeting neighbours. Water column nutrients further constrain the maximum amount of surface canopy biomass, apparently regulating the total number of Macrocystis fronds per square meter (the frond carrying capacity; North 1994, Tegner et al. 1997). The ontogenetic development of a Macrocystis cohort is, therefore, dominated by self-thinning (Schiel & Foster 2006), in which high densities of small individuals ultimately yield much lower densities of very large individuals (North 1994). The applicability of this self-thinning model in Macrocystis populations, however, has not been tested directly. North (1994) estimated the frond carrying capacity of a typical Macrocystis population to be 10 fronds m−2, whereas Tegner et al. (1997) found frond carrying capacity to vary according to oceanographic climate, being higher during cooler, nutrient-rich conditions (La Niña) and lower during warmer, nutrient-poor conditions (El Niño). Schiel et al. (1995) also observed at the Chatham Islands that a site with larger Macrocystis sporophytes had lower population densities than a site dominated by smaller Macrocystis sporophytes. Many researchers have estimated that self-thinning ultimately results in adult sporophyte densities of 1 per 10 m2 (Dayton et al. 1984, 1992, Graham et al. 1997), although the accuracy of this value has never been assessed experimentally. Furthermore, these studies have been restricted to pyrifera form populations in central, southern and Baja California. In other systems (e.g., Chile, New Zealand), pyrifera form individuals do not grow to large sizes or form large populations (Schiel et al. 1995) and conspicuous selfthinning of these populations has not been observed (Buschmann et al. 2004a, 2006). Similarly, 56



shallow-water integrifolia-form sporophytes exhibit vegetative propagation, resulting in coalescent holdfasts, and the concept of sporophyte self-thinning may be irrelevant to these populations (A. Vega & J.A. Vásquez, unpublished data). As previously described, Macrocystis microscopic stages have high light requirements and are thus highly vulnerable to inter- and intraspecific competition for light (Schiel & Foster 2006). Due to their small size, Macrocystis gametophytes and embryonic sporophytes are also highly vulnerable to sand scour (Dayton et al. 1984) and smothering by sediments (Devinny & Volse 1978) and by macro- (Dean et al. 1989, Leonard 1994) and mesograzers (Sala & Graham 2002). Finally, it should be noted that all of the above mortality agents typically result in small- to mesoscale variability in the stage and size structure of Macrocystis populations. During normal conditions, many factors are typically acting to regulate sporophyte survival in a probabilistic fashion, resulting in high variability in sporophyte abundance and size structure at the scale of tens to hundreds of metres (Edwards 2004). During episodic storms and ENSOs, however, multiple factors (e.g., wave intensity and nutrient limitation) may act simultaneously to produce massive stageand size-dependent mortalities homogeneously over broad spatial scales of 10s to 100s of km (Edwards 2004).

Dispersal, recruitment and population connectivity The field ecology of microscopic life-history stages is perhaps the most dynamic and least understood aspect of Macrocystis population biology (North 1994), and that of seaweeds in general (Santelices 1990, Amsler et al. 1992, Norton 1992). Previous life-history studies for Macrocystis indicate the potential for a wide variety of temporal and spatial variability in the time an individual remains within a life-history stage, or the time necessary to proceed to subsequent stages (Figure 4). This temporal flexibility in the life history begins with dispersal and ultimately results in variability in recruitment and thus demographic interactions within a population (Santelices 1990). Adult Macrocystis sporophytes typically produce zoospores with limited dispersal abilities (e.g., Anderson & North 1966, Dayton et al. 1984, Gaylord et al. 2002, Raimondi et al. 2004), suggesting a tight coupling between zoospore output, dispersal and recruitment (Graham 2003). Recent studies, however, have indicated that the supply of propagules of marine organisms can be decoupled from the adult demographic and genetic patterns, as propagules are dispersed far from their natal site (e.g., Roughgarden et al. 1988, Downes & Keough 1998, Wing et al. 1998, Shanks et al. 2000). This decoupling also seems to apply to Macrocystis (Reed et al. 1988, 2004, 2006, Gaylord et al. 2002), especially when the populations are not large enough for modification of currents by the canopy (Jackson & Winant 1983, Jackson 1998, Graham 2003). Because of their small size, Macrocystis zoospores will clearly be transported as far as available currents advect them (Gaylord et al. 2002). However, if adult sporophytes modify current directions and velocities, effective zoospore dispersal can be decreased, coupling zoospore supply to relative changes in the density and size structure of the adult sporophytes (Graham 2003). Subsequently, Macrocystis forests can vary between ‘open’ and ‘closed’ populations, depending on their size, isolation and geographic location (Graham 2003, Reed et al. 2004, 2006). Furthermore, Macrocystis zoospore dispersal can be enhanced by episodic periods of high zoospore production that coincide with storms (Reed et al. 1988, 1997), large population sizes (and thus high source zoospore concentrations; Reed et al. 2004, 2006) and turbulent resuspension of zoospores within the benthic boundary layer (Gaylord et al. 2002). Together, spatial and temporal variability in water motion, zoospore output and Macrocystis forest size results in high variability in the effective ranges of zoospore dispersal (Reed et al. 2006). Nevertheless, it is likely that the dispersal dynamics described for a few large Macrocystis forests in southern California are unique to this region (e.g., Point La Jolla and Point Loma at 1–8 km2; Dayton et al. 1984, Graham 2003, Reed et al. 2006) because most Macrocystis forests 57



worldwide are relatively small (90% of the taxa were more common in forested areas than deforested areas. Many of these associates had either clear trophic linkages with Macrocystis (e.g., abalones) or the associations were driven by habitat provision (e.g., kelp surfperch Brachyistius frenatus). Similar forestwide associations are found for Macrocystis populations in northern Chile, although species richness is much less than in California (Figure 6). Interestingly, the Chilean data show first that the presence of other kelp taxa (e.g., Lessonia) can also drive changes in kelp forest assemblage structure, and second that different species of kelp (e.g., Macrocystis vs. Lessonia) may differ in the quality and quantity of habitat that they provision. Furthermore, Vega et al. (2005) recently demonstrated that the morphology of the understory subtidal kelp Lessonia trabeculata varies in the presence/absence of Macrocystis, potentially resulting in additional effects of Macrocystis distribution on community structure. Still, these studies have relied on natural kelp deforestations, for example due to sea urchin overgrazing, which can produce various factors that confound variability in kelp presence. As such, the direct isolation of the importance of Macrocystis energy and habitat provision relative to that of other kelps or non-kelp macroalgae species remains elusive (Graham et al. 2007). Finally, due to the different growth rates and distribution of canopy biomass among the different forms of Macrocystis (e.g., pyrifera vs. integrifolia forms), shallow- and deepwater Macrocystis populations may also provide trophic and habitat resources in different ways. For example, interfrond distances among shallow integrifolia- and angustifolia-form sporophytes (Figure 2D) are much more homogeneous than their deeper pyrifera-form counterparts (Figure 2C), where high stipe densities are aggregated around individual sporophytes that are widely spaced. Such ecotypic variability in the spatial distribution of suitable habitat (i.e., canopy fronds) may 64



affect the nature and strength of species associations in shallow versus deep water, especially among fish taxa that preferentially use the canopy, mid-water fronds or water column spaces.

Trophic interactions and food webs During his initial observations of Macrocystis forests in southern Chile, Darwin (1839) was struck by the high diversity of species that appeared to be trophically linked to energy provision by Macrocystis. He professed “The number of living creatures of all orders, whose existence intimately depends on the kelp, is wonderful”. Indeed, all described Macrocystis forests harbour tens to hundreds of species, most of which feed either on Macrocystis-derived fixed carbon (from direct grazing to filter-feeding on Macrocystis detritus) or within some predatory trophic subweb founded upon Macrocystis-based herbivores (Graham et al. 2007). Although food web studies are rare (reviewed by Graham et al. in press), most species lists from within Macrocystis forests contain taxa from multiple trophic levels, often with numerous taxa within each level, and typically a wide variety of generalist and specialist species. It therefore seems unnecessary to review here the nature of trophic interactions within Macrocystis forests. Instead, the focus is specifically on recent studies of the fate of Macrocystis-based primary productivity and the importance of trophic interactions in the dynamics and stability of Macrocystis forests because these are topics of global interest. Gerard (1976) and Pearse & Hines (1976) estimated that although central Californian Macrocystis sporophytes had very high standing stock, most Macrocystis-based productivity entered the food web through detrital pathways. Storm waves rip entire sporophytes from the substratum, break fronds and erode senescent blades, sending detrital material of a wide size range to either the kelp forest floor or out of the forest to other systems. Large detrital pieces (i.e., drift) collect near the bottom where they are heavily grazed by asteroids, crustaceans (crabs, amphipods, isopods), snails and fishes (Leighton 1966, Feder et al. 1974, Gerard 1976, Pearse & Hines 1976, Beckley & Branch 1992, Kenner 1992, Hobson & Chess 2001; see also reviews by Castilla 1985, Foster & Schiel 1985, Graham et al. 2007). When present, large pieces of Macrocystis drift make up the primary diet of sea urchins (Leighton 1966, Castilla & Moreno 1982, Dayton et al. 1984, Vásquez et al. 1984, Castilla 1985, Ebeling et al. 1985, Harrold & Reed 1985). In Chile, Tetrapygus niger and Loxechinus albus can catch drift in Macrocystis forests (Castilla 1985, Rodriguez 2003) but to a lesser extent than Strongylocentrotus franciscanus or S. purpuratus in the northeast Pacific (Harrold & Reed 1985, Harrold & Pearse 1987). When deprived of drift, S. franciscanus or S. purpuratus abandon their normal ‘sit and catch drift’ strategy in search of attached algae (Mattison et al. 1977, Ebeling et al. 1985, Harrold & Reed 1985, Harrold & Pearse 1987), a behavioural switch not observed for other sea urchin taxa (Vásquez & Buschmann 1997, Steneck et al. 2002). Macrocystis drift is also the main component of the diet of abalone in California (Leighton 1966, Tutschulte & Connell 1988) and sea urchins and abalone are thought to compete strongly when Macrocystis drift is in short supply (Tegner & Levin 1982). In the absence of drift, abalones often decrease in abundance or disappear from the local system entirely (e.g., Graham 2004). Smaller detrital pieces (i.e., particulate organ carbon, POC) make Macrocystis-based productivity accessible to many more taxa (e.g., polychaetes, bivalves, sponges, crustaceans, ophiuroids, mysids or basically any kelp forest detrital or filter feeder; Clarke 1971, Foster & Schiel 1985, Beckley & Branch 1992, Kim 1992, Graham 2004, Graham et al. 2007). Therefore, it is not surprising that sessile filter feeders and mobile herbivores can be extremely diverse in Macrocystis forests (100+ taxa), with many taxa disappearing during local Macrocystis deforestation (Graham 2004). The direct grazing pathway is also utilised by a high diversity of kelp forest herbivores (Rosenthal et al. 1974, Gerard 1976, Pearse & Hines 1976, Moreno & Sutherland 1982, Castilla 1985, Foster & Schiel 1985, Graham 2004, Graham et al. 2007). Snails, crustaceans, asteroids and fishes can graze the benthic and water column biomass of Macrocystis sporophytes. These herbivores 65

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Figure 6 Macroscopic species abundances in areas of sea urchin overgrazing (deforested), Lessonia trabeculata forests and Macrocystis (integrifoliaform) forests for northern Chile (3–15 m depth). (A) Per cent cover of macroalgae; (B) per cent cover of sessile invertebrates.

Porifera Phragmatopoma Spionidae Balanus laevis Austromegabalanus sp. Bugula sp. Pyura chilensis Vermetidae

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Figure 6 (continued) (C) density of mobile invertebrates (no. m−2); (D) abundance (no. 20 min−1 of visual survey) of fishes. Data are means ± SD; sample sizes range from 34 to 136. Data are from Nuñez & Vásquez (1987), Vásquez et al. (1998), Salinas (2000) and W. Stotz, J. Aburto & L. Cailleaux, unpublished data.

Aplodactylus punctatus Scarthichthys viridis Doydixodon laevifrons Cheilodactylus variegatus Helcogrammoides cunninghami Pinguipes chilensis Hypsoblennius sordidus Graus nigra Chromis chrusma Paralabrax humeralis Trachurus murphyi Isacia conceptionis Prolatilus jugularis Auchenionchus microcirrhis Hemilutjanus macropthalmos Mugil cephalus Anisotremus scapularis Calliclinus genicuttatus Austromenidia laticlavia

Tetrapygus niger Tegula atra Tegula tridentata Allopetrolisthes sp. Nassarius gayi Crassilabrum crassilabrum Calyptraea trochiformis Mitrella unisfasciata Xantochorus cassidiformis Prisogaster niger Scurria sp. Pagurus sp. Ophiaycthis kroyerii Taliepus sp. Tricolia mcleani

Density (no. m−2)





generally have low per capita consumption rates (e.g., Jones 1971, Sala & Graham 2002) and probably have little impact on Macrocystis standing stock, except during population explosions (Jones 1971, Tegner & Dayton 1987, Graham 2002) or when Macrocystis sporophytes are small in size (e.g., during recruitment or recovery following disturbance; Moreno & Sutherland 1982, Harris et al. 1984, Castilla 1985, Sala & Graham 2002, Buschmann et al. 2004b). It is unknown whether the diversity of these herbivores within Macrocystis forests may enhance or lessen the effects of herbivory on Macrocystis survival and population dynamics, through complementarity or competition, respectively. It is well known, however, that sea urchins can have a great impact on Macrocystis standing stock through direct grazing (Lawrence 1975, Pearse & Hines 1979, Schiel & Foster 1986, Harrold & Pearse 1987, Vásquez & Buschmann 1997, Steneck et al. 2002, Vásquez et al. 2006). In some systems (e.g., southern Chile), Macrocystis fronds can be weighted down by epizoites and sea urchins can heavily graze water column biomass directly (Dayton 1985b). In most cases, however, the greatest impact of sea urchin grazing on Macrocystis biomass is when sea urchins aggregate on holdfasts and detach entire sporophytes, which then drift out of the system (North 1971, Foster & Schiel 1985); sea urchins (and potentially other herbivores) may then keep the system in a deforested state by grazing directly on Macrocystis recruits. Again, such overgrazing by Strongylocentrotus is apparently limited to periods of low drift availability (Ebeling et al. 1985, Harrold & Reed 1985), which occur episodically at local scales within southern California (Foster & Schiel 1988, Steneck et al. 2002, Graham 2004). Sea otters (Enhydra lutris) inhibit sea urchin overgrazing throughout the otter’s range (McLean 1962, Harrold & Pearse 1987, Foster & Schiel 1988, Watanabe & Harrold 1991), which at present is limited mostly to the Northern Hemisphere north of Point Conception (Laidre et al. 2001). Cowen et al. (1982) also observed that high wave action in central California curtailed sea urchin foraging and allowed algal recovery. In this region, seasonal variability in wave intensity was suggested as the most important factor regulating the abundance and structure of macroalgal assemblages (Cowen et al. 1982, Foster 1982). The mechanisms controlling sea urchin overgrazing in southern California (south of Point Conception), however, are controversial. In the absence of sea otters, various forms of abiotic and biotic regulation of sea urchin populations have been proposed (see Foster & Schiel 1988, Steneck et al. 2002). For example, storms and/or disease can wipe out large sea urchin aggregations over relatively broad spatial scales (Ebeling et al. 1985, Tegner & Dayton 1991, Lafferty 2004), whereas recruitment failure can limit replenishment of local sea urchin populations (Pearse & Hines 1987). Nevertheless, the most popular explanation for the lack of large-scale sea urchin barrens in the absence of sea otters in southern California is that other predators are controlling sea urchin abundance. Various kelp forest predators eat sea urchins (see review by Graham et al. 2007), although most only eat small sea urchins that are incapable of inflicting significant damage to Macrocystis holdfasts. Sheephead (Semicossyphus), lobsters (Panulirus) and the sunflower stars (Pycnopodia) appear to be the only Californian kelp forest predators other than sea otters that can feed on adult sea urchins. In some southern California kelp forests, data suggest that sheephead and lobster predation are important in controlling urchin abundance (e.g., Cowen 1983, Lafferty 2004) although the diets of these species are highly variable in space and may not include sea urchins even when sea urchins are present (e.g., Cowen 1986). Furthermore, in most cases the predation hypothesis is invoked in the absence of field experimentation (but see Cowen 1983), which is problematic since sheephead and lobsters have become relatively rare in southern California kelp forests (Dayton et al. 1998), yet deforested areas are also relatively rare (Foster & Schiel 1988). Sea urchin populations in southern California, therefore, are clearly regulated by multiple abiotic and biotic processes, probably resulting in the low frequency of sea urchin barrens in the Southern California Bight.

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The only other Macrocystis systems for which sea urchin overgrazing has been observed are in South America, although like California, large-scale overgrazing is rare (Vásquez & Buschmann 1997, Steneck et al. 2002). In northern Chile, large-scale overgrazing of Macrocystis by Tetrapygus niger appears to be limited by high water motion in the region where Tetrapygus and Macrocystis distributions overlap (Vásquez & Buschmann 1997). Sea urchin overgrazing in northern Chile is subsequently restricted to particular depth zones. Also, the asteroids Luidia and Meyenaster are solitary hunters within subtidal habitats and are important predators on Tetrapygus niger and other echinoids and asteroids (Viviani 1979, Vásquez & Buschmann 1997, Vásquez et al. 2006). Again, as in California, the role of sea urchin grazing in regulating Macrocystis populations in southern Chile is more controversial. Initial experimental results detected no effect of Loxechinus albus grazing on Macrocystis populations in the Beagle Channel (Castilla & Moreno 1982). On the other hand, Dayton (1985b) argued that Loxechinus albus grazing should significantly affect Macrocystis abundance along the protected coast of southern Chile where large asteroid predators could serve as a controlling factor and further suggested that the results of Castilla & Moreno (1982) were only relevant to the southernmost subpolar area (Beagle Channel). Additionally, density of an annual Macrocystis population in the archipelago region of southern Chile was significantly reduced from 24 to 2 sporophytes m−2 when Loxechinus albus densities exceeded 20 m−2 (Buschmann et al. 2004b). These contradictions have developed into an unresolved controversy about the ecological role of sea urchins in structuring Chilean Macrocystis populations (Vásquez & Buschmann 1997). Based on these numerous field studies spanning the global range of Macrocystis, simple trophic cascades do not seem to exist in Macrocystis-based systems. Various instances of overgrazing have been described (Steneck et al. 2002) but they are generally short-lived, observed at local scales and are often the result of overgrazing by particular trophic groups (e.g., sea urchins and amphipods). Nevertheless, due to the high diversity and productivity of Macrocystis-based food webs (Rosenthal et al. 1974, Pearse & Hines 1976, Castilla 1985, Foster & Schiel 1985, Graham 2004, Graham et al. 2007), these rare overgrazing events can have conspicuous ecological consequences (e.g., Graham 2004). The question still remains, however, as to why overgrazing is less frequent in Macrocystisbased systems than in other kelp-based systems (e.g., Aleutians, North Atlantic, Japan; Steneck et al. 2002). Three striking features are common to Macrocystis systems worldwide and may be important in buffering Macrocystis communities from overexploitation. The first is that all Macrocystis-based food webs are relatively diverse. Such high diversity, especially when it occurs at higher-order trophic levels, may provide a wider range of trophic interactions than less-diverse systems, minimising the impact of grazing by any given herbivore species. Such ecological effects of high diversity are supported by the field and experimental mesocosm studies of Byrnes et al. (2006), who found that increased predator diversity decreased the impact of an assemblage of grazers on Macrocystis biomass. In addition to the highly diverse systems of California, Beckley & Branch (1992) enumerated 200+ taxa in a Macrocystis system at the Prince Edwards Islands. Castilla (1985) identified 30+ herbivores and primary predators for the Macrocystis-based food web in the Beagle Channel, culminating with the generalist asteroid Cosmarestias lurida; Adami & Gordillo (1999) observed a similar system on the other side of the channel, although Loxechinus albus was conspicuously absent. Vásquez et al. (1998) observed similar trophic diversity in the Macrocystis system of northern Chile, which can include the Southern Hemisphere sea otter Lontra felina. Although L. felina does not feed on sea urchins (Ebensperger & Botto-Mahan 1997, Villegas 2002), these sea otters do forage on fishes, crustaceans and molluscs and may represent a diversifying component in these systems (Castilla & Bahamondes 1979). A poleward decrease in the diversity of Macrocystis systems appears to be present in both the Northern (Graham 2004, Graham et al. 2007) and Southern Hemispheres (Castilla 1985, Vásquez et al. 1998). The second commonality

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among Macrocystis systems is that they are all imbedded within high-productivity systems necessary to support Macrocystis survival, growth and reproduction. Therefore, the inherently high delivery of nutrients to global Macrocystis populations may simply override consumptive processes in regulating community structure and ecosystem processes over broad temporal and spatial scales. Finally, sea urchin recruitment appears to be more variable in space and time at low latitudes compared with high latitudes (Castilla & Moreno 1982, Foster & Schiel 1988, Buschmann et al. 2004b, Vega et al. 2005), potentially destabilising sea urchin population dynamics and decreasing the likelihood of large-scale sea urchin population explosions (Foster & Schiel 1988). Despite the numerous trophic studies of kelp forest organisms, however, there is a dearth of research on communitywide patterns of energy flow. Stable isotope methods have demonstrated the important role of detrital pathways in Macrocystis forests (e.g., Kaehler et al. 2000) and other systems (Duggins et al. 1989, Bustamante & Branch 1996, Fredriksen 2003). Macrocystis productivity is also exported to other systems (e.g., sandy beaches, deep-sea basins, coastal islands) where it may contribute greatly as an allochthonous energy source (Lavoie 1985, Inglis 1989, Vetter 1994, Harrold et al. 1998, Orr et al. 2005). The trophic consequences of Macrocystis production, however, have rarely been considered beyond the finite boundaries of the kelp forest. Clearly, Macrocystis systems are energy rich. Trophic interactions among kelp forest organisms can be conspicuous and are interesting avenues for ecological research. Yet, the emerging pattern over the last 150 yr of research is that at the community or ecosystem level, the diversity and productivity of Macrocystis systems are driven primarily by oceanographic processes that regulate the distribution, abundance and standing stock of the main foundation species, Macrocystis. It has recently been proposed, however, that the primary structural force for Macrocystis-based systems in southern California (Channel Islands National Park) is ‘top-down’ consumption (Halpern et al. 2006). Halpern et al. (2006) used satellite-derived chlorophyll-a data to estimate nutrient delivery to kelp beds as a proxy for ‘bottom-up’ processes. The effect of nutrients on algal abundance (primarily that of Macrocystis) was then determined to be significantly lower than consumptive effects. Off-shore chlorophyll-a concentrations, however, are not indicative of processes acting in the near shore (Blanchette et al. 2006) and nutrient delivery to off-shore plankton assemblages and near-shore kelp beds are two fundamentally different and negatively correlated processes (Broitman & Kinlan 2006). Additionally, it is well established that variability in Macrocystis sporophyte density (the abundance variable used by Halpern et al. 2006) is driven primarily by self-thinning and is unrelated to nutrient supply (North 1994), whereas nutrient supply and Macrocystis biomass are tightly coupled (North 1994, Tegner et al. 1996, 1997, Dayton et al. 1999). It would be interesting to know whether Halpern et al. (2006) would have obtained different results if they had used Macrocystis canopy biomass data available for the same region (Reed et al. 2006) and conducted their study beyond 1999–2002, which was the most ‘nutrient benign’ period of the last 50 yr. In fact, their study period did not include any of the conspicuous El Niño or La Niña events known to drive maxima and minima in community structure and energy flow within these systems (Dayton et al. 1999, Edwards 2004). A final concern with the approach of Halpern et al. (2006) is the inability of their correlative analyses to disentangle the confounding effects of habitat versus trophic associations. For example, one of their four conspicuous species that was correlated with Macrocystis abundance, and thus identified as a key consumer, was the striped surfperch (Embiotoca lateralis). Previous studies have repeatedly observed a negative association between E. lateralis and Macrocystis (Ebeling & Laur 1985, Holbrook et al. 1990). The mechanism underlying the association, however, is not a topdown trophic interaction but rather the negative effect of Macrocystis surface canopies on preferred foraging habitat (foliose algae) of Embiotoca lateralis. Another of the conspicuous species, the scavenger Kelletia kelletii, was found previously to be associated with sea urchin barrens rather than kelp forests (Behrens & Lafferty 2004), the exact opposite pattern from that predicted by the 70



top-down hypothesis. The confounding nature of habitat versus trophic interactions in driving kelp forest associations is probably ubiquitous among Macrocystis systems due to this species simultaneous provision of primary habitat and energy throughout much of its range (see ‘Macrocystis as a foundation species,’ p. 62). This criticism of the results of Halpern et al. (2006) does not mean that predation is unimportant in regulating the structure and dynamics of Macrocystis systems, but simply argues for greater caution when using correlative data to understand regulatory processes in this complex system.

Community consequences of climate change and kelp forest exploitation Climate change and human exploitation can affect the diversity and productivity of Macrocystis systems either by indirect modification of Macrocystis distribution, abundance and productivity or by directly modifying distribution, abundance and productivity of the flora and fauna that inhabit Macrocystis forests. Macrocystis productivity and distributional limits are largely constrained by environmental processes (see ‘Organismal biology’ section, p. 42). Studies of environmental control on Macrocystis systems, however, have been limited entirely to ecological timescales (e.g., Dayton et al. 1999). Over periods of years to decades, temporal changes in nutrient availability (as measured through sea-surface temperature proxies), sedimentation and substratum composition, storms and light have all been shown to modify the living space and carrying capacity of Macrocystis (e.g., North & Schaeffer 1964, Zimmerman & Robertson 1985, Seymour et al. 1989, North 1994, Tegner et al. 1996, 1997, Dayton et al. 1999, Edwards 2004). Such responses to short-term climate change have typically been associated with ENSOs (4- to 7-yr frequency; Dayton et al. 1999, Edwards 2004) and cycles in the Pacific Decadal Oscillation (PDO, 10- to 20-yr frequency; Dayton et al. 1999). In California, strong ENSOs affect Macrocystis populations in two primary ways: nutrient stress associated with deepening of the thermocline and destructive storm waves. The relative impacts of each of these processes, however, vary latitudinally (Edwards 2004). Conspicuous second-order community responses often result, for example, from reduction in the availability of Macrocystis standing stock and drift and subsequent overgrazing by sea urchins or crustaceans (Ebeling et al. 1985, Harrold & Reed 1985, Tegner & Dayton 1987, Graham 2002, Behrens & Lafferty 2004). Ecologically important echinoderms (e.g., sea urchins and seastars) often suffer mass mortalities that may also be associated with ENSOs (Tegner & Dayton 1987, Dayton & Tegner 1989, Dayton et al. 1992, Behrens & Lafferty 2004, Lafferty 2004). These high-frequency ENSO cycles are overlaid on longer-frequency PDO cycles, with warm PDO periods exacerbating ENSO cycles (Dayton et al. 1999); the two most destructive ENSOs on record (1982–1983 and 1997–1998) occurred during the most recent warm PDO period. Nevertheless, the most wellestablished effect of PDO cycling on Macrocystis systems is the correlation between larger Macrocystis sporophyte sizes during PDO cold periods relative to warm periods (Tegner et al. 1996, 1997). The importance of short-term oceanographic phenomena (e.g., ENSO) in regulating other Macrocystis systems is essentially unknown. During the 1997–1998 El Niño, northern Chilean Macrocystis populations increased, while black sea urchins (Tetrapygus niger) decreased (Vega et al. 2005), a pattern opposite to that observed in southern California. The 1997–1998 El Niño devastated Macrocystis populations in central Perú, decreasing sporophyte density and the diversity of associated species (Lleellish et al. 2001). The subsequent 1998–1999 La Niña, however, triggered high Tetrapygus niger recruitment and a significant increase in T. niger adult populations (Vásquez et al. 2006), which corresponded with a crash in Macrocystis populations (Vega et al. 2005). The 1997–1998 El Niño also affected northern Chilean asteroid populations. Luidia and Meyenaster are considered to be top predators in littoral benthic food chains of northern Chile, and both species prey upon Heliaster and Stichaster (Viviani 1979). Luidia and Meyenaster coexist and restrict the 71



bathymetric distribution of Stichaster and Heliaster in the intertidal and subtidal zones (Viviani 1979). Meyenaster and Luidia decreased significantly within Macrocystis populations during the 1997–1998 El Niño, potentially migrating to deeper water, whereas Heliaster and Stichaster increased in abundance during the same period (Vásquez et al. 2006). It remains to be determined whether the ENSO-driven decreases in Luidia and Meyenaster abundances, both important predators on Tetrapygus (Viviani 1979, Vásquez 1993, Vásquez & Buschmann 1997), were the ultimate causes of the 1998–1999 T. niger population explosion (Vega et al. 2005, Vásquez et al. 2006). Although nutrient deprivation is the most conspicuous intra- and interdecadal oceanographic stressor on Macrocystis physiology and survival, it has recently been shown that temperature shifts alone can have rapid impacts on Macrocystis systems from the organismal to community levels. Schiel et al. (2004) used an 18-yr dataset to study changes in the structure of a local Nereocystis kelp bed after 10 yr of increased ocean temperature (+3.5°C) due to the thermal outfall of a powergenerating station. Similar to the changes observed following deforestation in southern California (Graham 2004), Schiel et al. (2004) detected significant communitywide changes in 150 species of algae and invertebrates since the initiation of the thermal outfall. These community changes, however, were not consistent with a northern shift in the distribution of southern species, but rather a shift in the dominant canopy-forming kelp from Nereocystis to Macrocystis, and the potential shading effect of the Macrocystis surface canopy. These data demonstrate the difficulty in disentangling the direct effect of climate change on giant kelp communities from the indirect effect of climate change on the distribution, abundance and productivity of key habitat-forming and energyproducing species, like Macrocystis. Studies of the effects of natural and anthropogenic climate change on Macrocystis systems have been limited to the last few decades. The frequency and severity of ENSOs have been highly variable over geological timescales (Rosenthal & Broccoli 2004) and it has been suggested that their frequency is increasing (Diaz et al. 2001). Still, the ecological consequences of such longterm climate change to Macrocystis systems seem obvious; Schimmelmann & Tegner (1991) detected an ENSO signal in the flux of Macrocystis-derived organic carbon to the floor of the Santa Barbara Basin over 1500 yr. Less obvious, however, are interactions between long-term changes in ocean temperature, near-shore sedimentation, light and sea level that are driven by glacialinterglacial cycling (Graham et al. 2003). Macrocystis has limited depth, substratum composition and nutrient ranges, and ice age redistribution and modification of environmental conditions may have had massive impacts on Macrocystis distribution, abundance and productivity. For example, late-Quaternary sea-level rise probably led to large changes in inhabitable Macrocystis reef area around the Californian Channel Islands and mainland as broad near-shore rocky platforms became exposed, shrank and even fragmented (Graham et al. 2003, Kinlan et al. 2005). A recent study predicted that southern Californian Macrocystis kelp forest area and biomass increased up to 3-fold from the last glacial maximum to the mid-Holocene, but then rapidly declined by 40–70% during the late Holocene to current area and biomass levels (M.H. Graham, B.P. Kinlan, R.K. Grossberg, unpublished data). Furthermore, the early Holocene peak in Macrocystis distribution and abundance coincided with highly productive palaeo-oceanographic conditions, probably yielding a subsequent peak in Macrocystis productivity during that period. This shift overlapped with conspicuous changes in total biomass of kelp-associated species, such as abalone, sea urchins and turban snails in native American shell middens on the Channel Islands (Erlandson et al. 2005). The community and ecosystem consequences of such long-term climate change on Macrocystis systems can be predicted but critical tests of such predictions will require application of contemporary palaeo-ecological tools (e.g., stable isotopes) because Macrocystis sporophytes do not fossilise (Graham et al. 2003). Poor strategies of sewage discharge in the 1950s and 1960s were associated with the decimation of a few very large Macrocystis forests in southern California (North & Schaeffer 1964, North & Hubbs 1968, North 1971, Tegner & Dayton 1991). Stringent regulations, however, quickly remedied 72



the impacts. Tegner et al. (1995b) later found that nitrogenous wastes originating from breakage in sewage outfalls can actually have positive effects on Macrocystis recruitment, especially during periods of nutrient deprivation. It was also noted by Dawson et al. (1960) that an oil spill in Baja California, Mexico, had no direct impacts on Macrocystis physiology, yet positively affected Macrocystis survival by causing high local mortality of sea urchins. Effects of other pollutants, such as some metals and aqueous petroleum waste, on Macrocystis microscopic stages can inhibit microtubule dynamics, DNA replication, photosynthetic processes and overall physiology (Anderson et al. 1990, Garman et al. 1994, 1995, Reed & Lewis 1994). Despite these localised impacts, there is little evidence that chemical pollution currently restricts Macrocystis distribution, abundance and productivity over broad spatial and temporal scales. Finally, Macrocystis systems have been subjected to long-term anthropogenic exploitation, spanning a period of at least 11,000 yr (Erlandson et al. 2005). Recent attention has focused on direct exploitation of Macrocystis populations and Macrocystis-associated organisms, especially in southern California where Macrocystis has been harvested for algin extraction since the 1920s (North 1994). Californian harvests are limited to the upper 1–2 m of the water column and have been shown to have minimal impacts on sporophyte survival (see p. 56). Indeed, while there is considerable temporal variability in Macrocystis populations due to physical and biological factors, the long-term stability of the Macrocystis harvest suggests that it is one of the best-managed marine harvests of wild populations worldwide (Dayton et al. 1998). Nevertheless, in southern Chile, Macrocystis is harvested by abalone farmers who require biomass all year round and Macrocystis cultivation is now required to offset heavy exploitation of natural Macrocystis populations (Gutierrez et al. 2006). Due to the patchy distribution of integrifolia-form populations in northern Chile, Macrocystis harvesting near abalone farms has had a great impact on the dynamics of local Macrocystis populations, with subsequent effects on Macrocystis-associated communities (J.A. Vásquez, unpublished data). Although Macrocystis populations themselves appear to be relatively immune to episodic harvesting of the surface canopy, Macrocystis-associated organisms are not. Overfishing has resulted in virtual elimination of large predators in southern California Macrocystis forests (Dayton et al. 1998). The ecological impacts of overfishing on Macrocystis populations are unclear because some correlative studies suggest cascading impacts whereas others do not (Foster & Schiel 1988, Dayton et al. 1998, Steneck et al. 2002, Behrens & Lafferty 2004, Lafferty 2004, Graham et al. 2007). Although predators may or may not be more common within marine reserves (Paddack & Estes 2000, Behrens & Lafferty 2004, Lafferty 2004), predators within reserves are typically larger in size (Paddack & Estes 2000). Again, the problem lies in deciphering the various types and strengths of species interactions operating in Macrocystis forests (e.g., Macrocystis-derived habitat and energy provision compared with predation). One thing is clear, however, despite the ubiquitous role of Macrocystis-derived habitat and energy provision in enhancing kelp forest diversity and productivity worldwide, kelp forest organisms cannot survive targeted exploitation over large temporal and spatial scales (Dayton et al. 1998).

Conclusion The global scientific literature indicates that Macrocystis is an important provider of habitat and energy to its associated communities wherever it is present. It is also clear that, despite non-trivial gene flow among global Macrocystis populations, Macrocystis morphology and physiology are highly variable in response to the environmental conditions within which sporophytes recruit, grow and reproduce. These conspicuous ecotypic differences have generally led researchers to study Macrocystis population dynamics and community interactions from a regional perspective. Patterns observed for some large conspicuous giant kelp forests in southern California have subsequently 73



dominated the literature and become the paradigms against which the ecologies of other Macrocystis systems are compared. When viewed from a global perspective, however, regional differences in the results of prior descriptive and experimental studies can be reconciled by an appreciation of great plasticity in Macrocystis form and function. The origin, nature and potential restriction of such plasticity to Macrocystis are appealing paths for future research.

Acknowledgements We greatly appreciate the support of our funding agencies: J.A.V. and A.H.B. acknowledge FONDECYT (grants 1010706, 1000044 and 1040425) and the Universidad de Los Lagos; M.H.G. acknowledges the National Science Foundation (NSF 0351778 and 0407937), San Jose State University, and the Moss Landing Marine Laboratories. We also acknowledge Alonso Vega, Mariam Hernández, Pirjo Huovenin, René Espinoza, Lara Ferry-Graham, David Schiel, Sean Connell, Louis Druehl and Michael Foster for various levels of support and comments during production of this review.

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HABITAT COUPLING BY MID-LATITUDE, SUBTIDAL, MARINE MYSIDS: IMPORT-SUBSIDISED OMNIVORES PETER A. JUMARS School of Marine Sciences & Darling Marine Center, University of Maine, 193 Clark’s Cove Road, Walpole, Maine 04573, U.S. E-mail: [email protected] Abstract Mysids often dominate mobile benthic epifaunas of mid-latitude continental shelves. Macquart-Moulin & Ribera Maycas (1995) reported that the six most abundant species on western and southern European shelves are all strong diel migrators. Published daytime epibenthic sledge (sled) data from the surf zone to the shelf edge matched with published behavioural data on the most abundant species were used to test, confirm and extend that relationship to other coastal regions and to identify an association of abundant migrators with species that are important in fish diets. They also reveal another pattern: a correspondence between abundant surf-zone species and species that dominate estuarine faunas seasonally. Population concentrations at estuary mouths, sills of fjords and in the surf zone suggest a lifestyle dependent upon horizontal fluxes. Marine mysids that migrate between habitats are chronically undersampled in the field, however, and are underrepresented in food-web models. Unfortunately, no single methodology samples both pelagic and benthic individuals well and nearly all shelf measurements so far reported must be considered underestimates of local abundance. Mysids are major dietary components for many benthic and pelagic fishes, mammals, cephalopods and decapods, often for key life stages, and often because mysid migrations result in encounters with predators. Mysids can be extraordinarily omnivorous, with demonstrated capabilities to digest cellulose and diets spanning macrophyte detritus, more labile detritus, large microalgae, and smaller animals and heterotrophic protists. They can be sufficiently abundant and active to play roles in sediment transport. Contributing factors to their underappreciation have been the lack of fidelity of mysids to single habitats, coupled with higher fidelity of investigators to the study of single habitats. Sampling with classical methods has been problematic because of effective evasion by mysids, compounded by extreme patchiness associated with mysid schooling. Their frequent absence from coastal and even estuarine food-web models has not been more conspicuous because the combination of their migration and omnivory spreads their feeding impacts and because they are subsidised by horizontally imported plankton and seston and are themselves horizontally exported in the form of predator gut contents and biomass. They clearly link pelagic and benthic food webs in two important and ecosystem-stabilising ways, however, by feeding in both habitats and by succumbing in both habitats to both cruising and sitand-wait predators. Consideration of resource and predation gradients and limited data implicate horizontal, diel migrations as well, extending these linkages, especially in the onshore–offshore direction. Somewhat paradoxically, the same features that have made them difficult to study by classical means, in particular schooling, diet breadth, ontogenetic change in diet and migration between habitats, suit migrating mysids well to new, individual- or agent-based modelling approaches. Moreover, benthic observatories deploying acoustic technologies with spatial and temporal resolution sufficient to resolve individual migratory behaviours promise powerful tests of such models.

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Introduction For nearly two centuries, observations of zooplankton vertical migrations have aroused curiosity and elicited alternative and compound explanations (Pearre 2003). Selective forces evoking and altering these migrations include vertical gradients in resources, in predation risks and in environmental drivers of physiological rates (i.e., temperature and salinity). Such gradients in risks and benefits can be even steeper within the bottom boundary layer, including its upper layers of sediment (Boudreau & Jørgensen 2001), and laterally across fronts, than they are in the overlying water column. The focus of this review is on migrations between benthic and pelagic habitats by a subset of the animal community that may also move horizontally, both across and along isobaths, connecting more than two habitats. For that reason, in this review the more general term ‘habitat coupling’ is used rather than benthic-pelagic coupling (Schindler & Scheuerell 2002). Widespread use of echo sounders after the rapid advance of underwater acoustics in World War II brought attention to the ubiquity of vertical migrations and specifically to the oceanic deep scattering layer. Echo-sounder frequencies near 12 kHz that were useful for locating the bottom proved sensitive to air bladders of fishes and siphonophores. Based partly on such observations, Vinogradov (1962) developed a conceptual scheme subsequently dubbed ‘Vinogradov’s ladder’: although diel migrations from deeper than 600 m are rare, many deeper-dwelling species migrate part of the way to the surface, so that predatory interactions provide a chain or ladder for vertical redistribution of energy and materials that daily extends to depths in excess of 1000 m in the open ocean. The proliferation of acoustic Doppler current profilers (ADCPs, operating typically at 300–600 kHz; Brierly et al. 1998) and of bioacoustic instruments designed to detect zooplankton at acoustic frequencies typically ranging from 250 kHz to a few megahertz (e.g., Gal et al. 1999) is revealing the ubiquity and intensity in shallow waters of an inherently more complicated phenomenon that has been dubbed the shallow scattering layer (Kringel et al. 2003). In waters too shallow to hold a deep scattering layer, animals from many taxa have evolved foraging patterns and morphologies compatible with living in or on the bottom, usually during the day, and rising into the water column, usually at night. Although there is no need for a vertical ladder where the water is a single rung deep, early data already show the outlines of a horizontal or oblique, onshore–offshore ladder in the coastal zone. Though still quite limited in number, deliberate, multifrequency acoustic studies of shallowwater migrators suggest that water-column abundances (depth-integrated biomasses) of these migrants may frequently exceed those of the holoplankton. This suggestion led to a systematic examination of corroborative evidence for the ecological importance of these migrants. For pragmatic reasons, in this review analysis is limited to a single large taxon, the Mysidacea (commonly known as opossum shrimp), that appears in shallow, mid-latitude seas and often dominates such migrations. Similarly, the focus is limited to subtidal, coastal habitats of mid-latitudes and to species that occur outside estuaries during at least some seasons of the year. Work in other marine, estuarine and freshwater environments is cited selectively when comparable information was not at hand for mid-latitude marine systems. Literature on freshwater species or (oligohaline and mesohaline) estuarine endemics has not been reviewed for the simple reason that the importance of vertically migrating mysids in these systems is widely appreciated (e.g., Rudstam et al. 1989, Kotta & Kotta 2001a, Viitasalo et al. 2001). To avoid inflation of inferred importance by selective extraction of conspicuous examples of migration and to give some insight into migrations of individual species, a two-step process was used. The first step was to identify a few regional studies of mysids notable for the spatial or temporal extent (or both) of their epibenthic sledge sampling. Thus, this review is also focused away from hard bottoms, caves and vegetation, all habitats well exploited by mysids but ones requiring different census methods. The second step was to review characteristics of migrations in the mysid species that dominated samples in these studies. In both steps emphasis 90


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has been on studies published after the major review by Mauchline (1980), citing prior literature primarily when a particular citation was omitted by Mauchline (1980) or when focusing on information that was not summarised by Mauchline (1980). Additional information was then reviewed, substantiating the importance of mysids to coastal and estuarine ecosystems. Confluence of multiple lines of evidence for the importance of migrating mysids to both benthic and pelagic systems proved compelling. They often dominate diets of both pelagic and benthic fishes in coastal waters and estuaries, highlighting the multiple risks inherent in the migratory lifestyle. Mysids also appear to be important to the population abundances of some of their prey species, but for the most part mysids are remarkable dietary generalists when all life stages and habitat phases are included, and so are underappreciated stabilisers of the communities that they inhabit and transit (McCann & Hastings 1997). Recently, their migrations have been implicated as an important factor in sediment dynamics (Roast et al. 2004). Major habitat changes due to climate, introduced species or human intervention have often produced major changes in mysid populations that resonate through the food web. Despite underlying differences between mysid-containing food webs in fresh and marine waters, analogy with lake systems takes advantage of their closed boundaries to assess effects of mysid introduction, which are detected both up and down the food web. Another indicator of potential importance is the latitudinal range and habitat diversity over which high abundances of even single species are found; Neomysis americana (S.I. Smith, 1873) is abundant from Nova Scotia to Florida and from shelf habitats 100 m deep to salt marshes; in the last century it was introduced to the Atlantic coast of South America, where it has become an important food-web component. The importance of N. americana as food for both benthic and pelagic fishes over a broad geographic range was recognised in its original species description (Smith 1873). The question naturally arises as to why, despite engaging, comprehensive treatments of their capabilities and roles (e.g., Mauchline 1980) and intense and sustained interest among the specialists cited in this review, mysids do not figure more prominently in fisheries and oceanographic models and texts. The most direct comparison is with the largely holoplanktonic euphausiids, a group of similar body size and also large dietary breadth (but less expansion into detritivory) as a group. The summary by Mauchline (1980) of both groups followed a parallel structure for each. Tellingly, his chapter on “Mysids in the marine economy” is half as long as its counterpart for euphausiids, and only a small portion is devoted to shallow-water species that migrate. Biological oceanographic textbooks in general give an even more lopsided treatment. Reasons for this shortage of information are manifold. Shallow-water migrations are fundamentally more complicated than better-studied migrations in the open ocean or in coastal holoplankton because component populations in benthic and pelagic habitats cannot be studied by the same means and often are not sampled by a single investigator. Their natural reference frame shifts back and forth between an Eulerian fixed reference frame and a Lagrangian, water-mass-following reference frame with the change between benthic and pelagic habitats, respectively, seriously complicating description and analysis. Even when they stay within the pelagic or benthic habitat, mysids are notoriously poorly captured because of their effective evasive behaviours. More subtly, their lack of freely released eggs or larvae leaves no evidence of large mysid populations in lowflow or small-aperture capture devices, such as continuous plankton recorders, that efficiently recover those non- or weakly swimming life stages in euphausiids, decapods and fishes. Extreme patchiness of mysid populations, reinforced by schooling behaviours, make precise abundance estimates even more difficult to achieve than they are for non-schooling animals. The migratory lifestyle gives mysids access to horizontally imported pelagic food sources and leads through encounter to their local export as gut contents and assimilated biomass of fishes and decapods, effectively camouflaging their importance to local food webs and energy budgets; a large net import or export would be far more conspicuous. 91


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Migrations between the sea bed and the water column also generate semantic difficulties. Mees & Jones (1997) took a habitat point of view and defined the hyperbenthos as those animals living in the water layer immediately above the bottom. In this sense, migratory mysids spend part of their time as hyperbenthos. The term fails, however, to capture the range of habitats occupied by migratory mysids because in clear, shallow waters without bottom cover in the form of crevices or vegetation, mysids often bury themselves during daylight, disappearing from the hyperbenthos. Mysid migrations also exhibit considerable plasticity, varying in timing, intensity and vertical extent seasonally, night to night and with tides (e.g., Abello et al. 2005, Taylor et al. 2005). To pursue these migrations further from a habitat perspective thus would require more elaborate terminology than even the refinements proposed recently by Dauvin & Vallet (2006). Instead, in the present review an alternative approach is adopted that may lead more readily to quantitative models and predictions by taking the perspective of an individual migrating through habitats. It is noted that, because mysids swim actively when pelagic and may do so at times during their benthic phases, this perspective is not truly Lagrangian (in the normal physical oceanographic sense of tracking a parcel of water), although it follows that same spirit of following the entity of interest. Seeking the simplest terminology that has this behavioural focus, the term ‘emergence’ is used herein to describe the overall vertical migration behaviour between habitats and more specifically the upward component of the migration (leaving the distinction to context). This usage follows precedent for those who have focused on the migratory behaviour rather than on community structure in the hyperbenthic habitat (Saigusa 2001). When the shift is from pelagic to benthic, the term ‘re-entry’ is used in the current review, reflecting the author’s benthic background. Two recent developments promise accelerated understanding of the role of migratory mysids. One development is the continued evolution of bioacoustic instrumentation and its deployment methods, particularly in the context of high-power, high-bandwidth ocean observatories. The second advance is the rapid development of flexible, individual-based models (IBMs). Many of the same features that have made mysids difficult to study make them excellent subjects for applications and tests of IBMs (i.e., their schooling behaviours, their occupation of multiple habitats, their use of multiple food resources and their shifts in behaviour during development) (Grimm & Railsback 2005, Grimm et al. 2005). The combination of new technologies and models promises accelerated advances in understanding of the extents, causes and consequences of mysid migrations through tests of predictions about habitat usage.

Migratory capabilities, schooling and their consequences Credibility of evidence for migrations rests in some measure on the sensitivities of sensory mechanisms to guide them and on swimming capabilities. Mysids as a group are well endowed in both of these categories (Mauchline 1980). The earliest (Carboniferous to Jurassic) mysids appear to have been holopelagic, and the transition to emergence to have been marked by the evolution of statocysts with mineralised statoliths (Ariani et al. 1993), likely associated with the fitness enhancement of directional guidance in emergence and re-entry. Marine species generally (including Neomysis americana) secrete fluorite (CaF2), whereas low-salinity estuarine and freshwater species generally secrete vaterite (a CaCO3 polymorph of calcite and aragonite), although particular species provide exceptions to each generalisation that reflect their lineages (Ariani et al. 1993). Mysids have major impact on the marine fluorine cycle (Wittman & Ariani 1996), and their statoliths may be abundant enough in some fossil marine strata to warrant extraction (Voicu 1981). Likewise, calcite (transformed vaterite) from statocysts represents a substantial fraction of some Miocene Paratethys deposits in the Ponto-Caspian region, where use of calcium carbonate minerals appears to have first evolved in mysids (Ariani et al. 1993). Emergent mysids thus appear to have been very abundant in coastal ecosystems for a very long time, and they are still abundant enough to leave 92



detectable statoliths in modern shelf sediments (Enbysk & Linger 1966). In addition to a pair of statocysts for vertical orientation, a less well-identified mechanism for sensing depth is present (Rice 1961, 1964) that is sensitive to pressure changes equivalent to less than 1 m of water and that probably enables observed tidal rhythms in activity cycles (Mauchline 1980, Saigusa 2001, Gibson 2003, Taylor et al. 2005). In terms of horizontal navigation, mysids have long been known to utilise polarised light (Bainbridge & Waterman 1957, 1958), and movements of their stalked eyes are co-ordinated with information from the statocysts (Neil 1975a,b,c). Contrary to opinion in many recent references, polarisation (specifically e-vector orientation) is a useful indicator of solar azimuth throughout continental shelf depths and through most of the day, with the highest information content near dusk and dawn because of high inclination of the e-vector with respect to the horizontal (Waterman 2005). Seasonal onshore–offshore migrations have been inferred from asynchronous seasonal changes in abundance across habitats (e.g., Bamber & Henderson 1994), and polarised light probably provides the directional cue, although it often is not clear to what extent the asynchrony in local abundance is due to migration versus seasonally changing, local differences in population growth and mortality (Mees et al. 1993). For reasons that also are not yet clear, a majority of onshore–offshore migrators show winter maxima offshore, extending in high abundance into shallower water and estuaries during some or much of the period from spring to fall (Mauchline 1980). Diel homing to the same location over smaller scales has been documented experimentally in reef mysids (Twining et al. 2000). Utilisation of estuarine circulations to help maintain horizontal position on intermediate scales has also been observed (i.e., either an interaction of horizontal and vertical bias or directed navigation) (e.g., Orsi 1986, Moffat & Jones 1993, Schlacher & Wooldridge 1994, Kimmerer et al. 1998a,b), although variation in such behaviours with local conditions from year to year can be considerable (Kimmerer 2002), as can differences among mysid species at the same estuarine location (Sutherland & Closs 2001). Retention-assisting, horizontal migrations also have been observed during slack tides (Köpcke & Kausch 1996). Many mysid species are documented to be strong swimmers. Sustained swimming at 10 body lengths s−1 is not unusual, with bursts in some species exceeding 20 body lengths s−1 (Mauchline 1980). At these sustained speeds, diel vertical, diagonal or horizontal excursions on the order of 1 km would be feasible, depending on local flow velocities, so diel vertical migrations to the shelf edge are well within mysid capabilities. Habitats with flow speeds in excess of sustainable swimming speeds appear to be avoided, however, and mysids shelter behind flow obstructions and in the most slowly moving water layer directly over the bottom (Roast et al. 1998, Lawrie et al. 1999). Perhaps the most important point to emphasise in this introduction is the reason to focus on both abundance and migration. A point forgotten all too easily is that ecological importance to individuals of another species is usually a function of interspecific encounter rates (Hurlbert 1971), themselves a product of areal or volumetric abundance times relative velocity (e.g., Jumars 1993). The combination of good sensory guiding mechanisms and strong swimming capabilities would tend toward ballistic encounter during organised migrations, an advantage in feeding but a disadvantage when being preyed upon (Visser & Kiørboe 2006). Encounters in mysids are often modulated by schooling behaviours. Mysids use visual and tactile senses to form and maintain both highly polarised schools and less polarised aggregations or swarms (Ritz 1994). Very large aggregations of varying local density and orientation are termed shoals (Clutter 1969). Typical mysid schools range from 1–10 m in linear dimensions and 1–15 m3 in volume (Ritz 1994). Near the bottom, school shapes often become planar, typically with more than one layer of mysids and sometimes differing in vertical structure by sex and life stage. Moving schools tend to be elongate, whereas stationary swarms (albeit containing milling individuals) are more circular (when near the sea bed) or spherical (Clutter 1969, Wittman 1977, Ohtsuka et al. 1995). Schooling is typical of animals out from the cover of vegetation and swimming off the 93


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bottom (i.e., in the pelagic phase), even when only a few centimetres from the bottom, but schools may maintain oriented, evenly spaced formation while on the bottom. Mysids on or near the bottom typically orient into the current (Mauchline 1980). Densities in swarms are often near 105 of individuals (ind.) m−3, with mean interindividual separation distances near 2 cm; for a single layer, that spacing yields about 2500 ind. m−2 (Mauchline 1980). The first emergence event of the night shows clear schooling and a constant ascent velocity dependent on depth and local light conditions, but later emergence does not appear to be as organised; schooling may not be maintained through the night (Kringel et al. 2003, Abello et al. 2005, Taylor et al. 2005). One function of schooling is to reduce average risk per individual (Ritz 1994) to individual predators, although schooling predators or large individual predators (e.g., whales) may be quite effective in the presence of mysid schooling. It is clear from gut contents of benthic and pelagic fishes that migrating mysids still incur fatal risk and that fitness loss must be counterbalanced by even greater gain from migrations if the migrations persist. Hence, the observations made by Macquart-Moulin & Ribera Maycas (1995) in an exhaustive sampling programme of the pelagic phase of mysids throughout the water column in the northwestern Mediterranean take on particular significance: they observed that the most abundant mysids found on the continental shelves of Europe are diel migrators between the sea bed and the pelagic environment. Based on the integration of a large number of studies with varying types of sampling gear over a long period, Macquart-Moulin & Ribera Maycas (1995) concluded that six species showed high benthic abundance on the shelf: Gastrosaccus sanctus (Van Beneden, 1861), G. spinifer (Goës, 1863), Anchialina agilis (G.O. Sars, 1877), Haplostylus lobatus (Nouvel, 1951), H. lobatus var. armata (Nouvel, 1951), and H. normani (G.O. Sars, 1877). They provided compelling new data from the region near Marseille of migration to the surface in all six of these taxa. Deprez et al. (2005) regard Gastrosaccus sanctus as a synonym of G. spinifer and Haplostylus normani as a synonym of H. lobatus. Macquart-Moulin & Ribera Maycas (1995) also found strong evidence of offshore migration or transport of Anchialina agilis and Haplostylus lobatus, both captured over bottoms 700–1000 m deep, where individuals are not known to occur on the bottom. They captured pelagic Anchialina agilis in bathyal waters during the day and collected a high percentage of dead animals, suggesting that occurrence in waters deeper than 500 m is an extension beyond suitable habitat.

Methods of data collection To identify recent published records of mysid abundance, three sources have been used in this review: the Food and Agriculture Organisation of the United Nations’ Aquatic Sciences and Fisheries Abstracts (ASFA), Thomson Scientific’s Web of Science and Google Scholar. The first two sources are limited primarily to citations later than those in the review of Mauchline (1980), but the third source is expanding rapidly into older literature. Into the search fields of the first two databases, ‘mysi*’ was entered and a country name that has a continental shelf, or in the case of the United States or Canada, a state or province name, respectively. For Google Scholar ‘mysid’ was used and the place name. For ASFA and the Web of Science, the ‘and’ is a Boolean operator. For Google Scholar, it was omitted (as Google in general ignores small, common words unless they are within explicit quotation marks). From the references returned, selected were those that documented mysid abundance either over an extensive period (a year or more) or a broad geographic area or both during daytime on the basis of epibenthic sledge samples. Many of these sledge studies used multiple, vertically arrayed nets (e.g., Zouhiri et al. 1998) to get information on near-bottom vertical distributions, biased to an unknown degree by species-specific escape responses. Such samples are referred to as ‘vertically resolved, epibenthic-sledge samples’. From the data provided, mysids have been ranked in terms of their abundances, selecting the one to five abundant and 94



frequent species, using a smaller number when a natural break point in abundance occurred (a difference of an order of magnitude or more in absolute abundance), and using the largest number when a long study over a large area showed consistent dominance of one species in at least one location and season. In each case, the choices of taxa are explained. Species names in quotation marks were then used as search terms in the same three databases to determine the migratory behaviour of the most abundant species. In addition, species names were searched in the NeMys database (Deprez et al. 2004, 2005) using the inclusive list of species names (valid and invalid) for references on behaviour and as a further check for inclusivity of publications with extensive sampling of field abundance and publications on migratory behaviour. The NeMys database was also used as one source of taxonomic authority, indicated on first use of the species name in the body of this review, and for some information (especially for European species) about geographic and depth ranges. For brevity, depth ranges of the species are summarised only in tabular form (Table 1). For consistency, only benthic capture records were used. Where taxonomic ambiguities or disputes over synonymy might affect conclusions, all databases were searched under both names. Table 1 Mysid species identified as abundant in epibenthic sledge samples, along with their known depth ranges, diel migratory behaviours and the study that established their high abundance Species Mesopodopsis slabberi Schistomysis spiritus Schistomysis kervillei

Depth limits (m) (respective citations) 1–42 (Buhl-Jensen & Fosså 1991, Beyst et al. 2001) 1–116 (Buhl-Jensen & Fosså 1991, Beyst et al. 2001) 1–25 (Cornet et al. 1983, Beyst et al. 2001)

Anchialina agilis

2–493 (Bacescu 1941, Cartes & Sorbe 1995)

Gastrosaccus spinifer

1–260 (Lagardère & Nouvel 1980, San Vicente & Munilla 2000) 17–420 (Lagardère & Nouvel 1980, Dauvin et al. 2000) 10–150 (Lagardère & Nouvel 1980, Dauvin et al. 2000) 1–125 (Bacescu & Schiecke 1974, Cunha et al. 1997) 6–407 (Brattegard & Meland 1997) 5–512 (Elizalde et al. 1991, San Vicente & Munilla 2000)

Haplostylus lobatus Haplostylus normani Erythrops elegans

Schistomysis ornata Leptomysis gracilis

Diel migratory behaviour (citations) Very strong swimmer, extends vertical distribution at night (Apel 1992, Wang & Dauvin 1994) Moderately strong swimmer, extends vertical distribution at night (Apel 1992, Wang & Dauvin 1994) Weaker swimmer, extends vertical distribution at night (Apel 1992, Wang & Dauvin 1994) Strongest swimmer and migrator to limits of its benthic depth distribution; most of the population leaves the bottom every night (Macquart-Moulin & Ribera Maycas 1995) Strong swimmers and migrators (MacquartMoulin & Ribera Maycas 1995)

Source of information on abundance Beyst et al. 2001

Beyst et al. 2001

Beyst et al. 2001

Dauvin et al. 2000

Dauvin et al. 2000

Strong swimmers and migrators (MacquartMoulin & Ribera Maycas 1995)

Dauvin et al. 2000

Strong swimmers and migrators (MacquartMoulin & Ribera Maycas 1995)

Dauvin et al. 2000

May be a diel migrator (Vallet et al. 1995)

Zouhiri et al. 1998

Collected in nighttime surface samples in some seasons; may be a diel migrator (Sorbe 1991) Strong migrator (Mauchline 1980, Kaarvedt 1989)

Zouhiri et al. 1998 Cornet et al. 1983, Cunha et al. 1997 (continued on next page)

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Table 1 (continued) Mysid species identified as abundant in epibenthic sledge samples, along with their known depth ranges, diel migratory behaviours and the study that established their high abundance Species Mysideis parva

Neomysis americana

Americamysis bigelowi Erythrops erythrophthalma Metamysidopsis elongata Neomysis kadiakensis Xenacanthomysis pseudomacropsis Neomysis rayii Acanthomysis stelleri Archaeomysis kokuboi Archaeomysis japonica Iiella ohshimai Nipponomysis ornata

Depth limits (m) (respective citations) 120–519 (Bacescu & Schiecke 1974, Elizalde et al. 1991) 1–232 (Wigley & Burns 1971)

4–179 (Wigley & Burns 1971, Allen 1984) 16–450 (Petryashev 2002a) 1–14 (Clutter 1967) 1–210 (Petryashev 2005) 1–104 (Petryashev 2005) 1–79 (Petryashev 2005) 1–104 (Petryashev 2005) 0–2 (Petryashev 2005) 1–50 (Hanamura 1997) 1–5 (Takahashi & Kawaguchi 1995) 1–5 (Yamamoto & Tominaga 2005)

Diel migratory behaviour (citations)

Source of information on abundance

Non-migrator (Elizalde et al. 1991). No description found.

Cornet et al. 1983, Cunha et al. 1997

Strong diel, tidally modulated migrator, but perhaps not to the full extent of its depth range (Herman 1963, Brown et al. 2005, Taylor et al. 2005) Strong diel migrator (Williams 1972)

Wigley & Burns 1971

Migrates at least in some environments (Brunel 1979) Slight upward shift of population mode at night (Clutter 1969) Strong diel migrator (Kringel et al. 2003) Caught in mid-water trawls (Wing & Barr 1977) Caught in mid-water trawls (Wing & Barr 1977) Poorly known Strong diel migrators (Takahashi & Kawaguchi 1997) Strong diel migrators (Takahashi & Kawaguchi 1997) Strong diel migrators (Takahashi & Kawaguchi 1997) Undescribed?

Wigley & Burns 1971 Wigley & Burns 1971 Clutter 1967 Clutter 1967 Kim & Oliver 1989 Kim & Oliver 1989 Kim & Oliver 1989 Takahashi & Kawaguchi 1997 Takahashi & Kawaguchi 1997 Takahashi & Kawaguchi 1997 Hanamura & Matsuoka 2003, Yamamoto & Tominaga 2005

Given demonstrated mysid capabilities for social aggregation and movement, reported maximal local abundances per unit of volume of water are not very informative regarding typical regional abundances, and documentation of consistently high abundance over a long time or broad region is a better indicator of consistent importance. This review therefore focused on a subset of those references that provide abundance estimates from epibenthic sledge samples taken during the benthic phase (i.e., when individuals are most susceptible to capture by a sledge). Drawbacks are that these studies varied widely in the geometries of the net mouth openings used and that many of these papers reported only numbers per unit of volume filtered (as determined by flow meter). No attempt was made to express abundances per unit of volume or per unit of area when the original author did not do so. For ease of comparison, however, all areal or volumetric abundance estimates given per total area or volume of tow were converted to numbers per square or cubic metre.

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Regionally abundant mysids and their migration habits European shelves One of the most challenging environments to sample with respect to abundance and emergence behaviours is the shallow subtidal and in particular the surf zone. In 15 monthly samples with a bottom sledge hauled by hand from four sites in the Belgian surf zone, Beyst et al. (2001) found average animal densities to exceed 15 ind. m−2 and to vary in ash-free dry weight (AFDW) from 3 to >30 mg m−2. Three quarters of individuals overall were mysids, primarily of three dominant species (Mesopodopsis slabberi (Van Beneden, 1861), Schistomysis spiritus (Norman, 1860) and S. kervillei (G.O. Sars, 1885)), and mysids dominated AFDW in some seasons. As is typical of such estimates, sampling efficiency is unknown for this sledge with these species and is assumed to be 100% for purposes of the calculation, so true densities must be higher. The three-species group also dominates the Voor delta, where Gastrosaccus spinifer is also abundant (Mees et al. 1993). Patterns of diel migration in Mesopodopsis slabberi are not as well known as might be expected. The species was described from in situ observations (Wittman 1977) as active and colourless during the day, showing no visible substrata preferences and changing leadership within schools spontaneously. Wittman (1977) also noted that schools did not appear to return to the same location and that predator-evading swarms veered horizontally without changing depth unless the predator attacked from above, a behaviour that should aid in capture by an epibenthic sledge. Daytime schools swam up to 50 cm above the substratum. Although Wittman (1977) did not specifically name M. slabberi in that context, he implied that nocturnal expansion among the mysids he studied was the norm. Wang & Dauvin (1994) found M. slabberi in epibenthic sledge samples both night and day and concluded from its upward skewed distribution among vertically resolved samples that it is an active swimmer, consistent with the observations of Wittman (1977). Wang & Dauvin (1994) caught more individuals in nighttime sledge samples but remarked that it might have been because of increased capture efficiency (less evasion in the dark). Zouhiri et al. (1998) in another series of epibenthic sledge samples found crepuscular peaks in capture of M. slabberi, consistent with the idea of distribution broadening above the bottom at night (and perhaps net evasion in the light). An inference consistent with most observations and directly supported by paired benthic and pelagic samples in the Jade estuary is that M. slabberi is concentrated near the bottom during the day but spreads into the water column at night (Apel 1992). This spreading includes a horizontal component, into the intertidal zone of at least one estuary during the night (Colman & Segrove 1955). It is worth noting that whether a seaward expansion also occurs in surface waters is unknown. In a long-term study of the polyhaline zone of the very turbid Gironde estuary, however, M. slabberi was captured abundantly in surface waters during daylight (Castel 1993, David et al. 2005). In addition, in the region of South African surf-zone diatom blooms M. wooldridgei Wittman, 1992 (closely enough related that it was previously identified as M. slabberi) also migrated onshore at night to take advantage of sinking surf-zone diatoms carried offshore in rip currents (Webb & Wooldridge 1990) at the same time that another mysid species, Gastrosaccus psammodytes Tattersall, 1958, migrated offshore from its inner surf-zone, daytime habitat to take advantage of that same resource (Webb et al. 1988). Mesopodopsis slabberi appears to migrate offshore in winter but to include vertical migrations in its repertoire there at 20 m water depth (van der Baan & Holthuis 1971). Even in winter, however, this species is observed inside but near the mouths of some estuaries (Mees & Hamerlynck 1992), so the entire population does not migrate offshore seasonally, and both migration and site-dependent mortality need to be examined as components of the distributional shift. Seasonally, M. slabberi also enters tidal creeks of salt marshes at high tides in sufficient abundance to be important as a prey species there (Hampel et al. 2003a), but it 97


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is unclear to what extent active behaviour versus passive advection is responsible (Hampel et al. 2003b). Recent molecular genetic work shows some genetic differentiation among populations in the northeast Atlantic and Mediterranean and Black Seas (Remerie et al. 2006) and also shows what may be an important mysid trait allowing rapid adaptation (i.e., high intrapopulation genetic diversity). Schistomysis spiritus and S. kervillei show similar migration patterns to Mesopodopsis slabberi, including nocturnal vertical spreading (van der Baan & Holthuis 1971, Apel 1992, Wang & Dauvin 1994). Of these two congeners, nighttime expansion into very shallow water has been reported for Schistomysis spiritus (Colman & Segrove 1955). Wang & Dauvin (1994) documented near-bottom, daytime vertical distributions and near-bottom, nighttime spreading patterns in vertically resolved sledge samples that allowed them to rank swimming activity as Mesopodopsis slabberi > Schistomysis spiritus > S. kervillei, with Gastrosaccus spinifer in the same category as Schistomysis kervillei and none of these mysids in their lowest activity category. Mesopodopsis slabberi is the most widely distributed of the three species geographically, ranging from Iceland in the Atlantic to North Africa, widely through the Baltic and Mediterranean and into the Black Sea (Deprez et al. 2005). Both Schistomysis congeners have somewhat more restricted geographic distributions than Mesopodopsis slabberi, with Schistomysis spiritus ranging from the Baltic to northern France and S. kervillei ranging from the North Sea to the southern Atlantic coast of France (with a report from northwest Africa), but its habitat distribution is comparable, ranging from shallow estuarine to shelf depths (Deprez et al. 2005). Within the North Sea, S. spiritus, S. kervillei and Mesopodopsis slabberi peaked in abundance near shore (Dewicke et al. 2003). Late-summer abundances (all mysids combined) averaged near 30 m−3 and 30 mg AFDW m−3 in sledge samples from the nearshore zone. All three species, however, reached even higher densities in the polyhaline and mesohaline zones of estuaries (e.g., Castel 1993, Wang & Dauvin 1994, Delgado et al. 1997, Azeiteiro et al. 1999, Lock & Mees 1999, Dauvin et al. 2000, Mouny et al. 2000, Wittman 2001, Drake et al. 2002, Dewicke et al. 2003). In terms of winter distributions, all three species are known to occur inside estuaries (near the mouth of the Schelde; cf. Mees & Hamerlynck 1992), in shallow coastal waters of warm regions (e.g., Lock & Mees 1999) and also offshore (van der Baan & Holthuis 1971). In deeper waters of the English Channel and the European shelf, other mysid species become dominant. Dauvin et al. (2000) presented a summary of 432 epibenthic sledge samples taken at 15 stations in the English Channel, including 3 stations within the Seine estuary, and covering the years 1988–1996. Those three stations have been excluded from the analysis in this review, except to note that they support the habitat pattern observed elsewhere for M. slabberi (i.e., shallow-water marine plus polyhaline-mesohaline estuarine water). The indisputable dominant in terms of abundance and frequency of occurrence outside the Seine estuary is Anchialina agilis, with mean abundances >1 ind. m−3 at both of the deepest stations, a coarse sand off Roscoff and a medium sand off Plymouth, both at 75 m depth. The species occurred at all the stations outside the Seine. Four other species occurred at over one half of the non-estuarine stations and reached mean abundances of at least 1 ind. m−3 at a minimum of one station: Gastrosaccus spinifer, Haplostylus lobatus, H. normani and Schistomysis ornata (G.O. Sars, 1864). Other studies in the same region by the same group of investigators appear consonant with these broad conclusions (e.g., Vallet et al. 1995, Vallet & Dauvin 1998, 2001), although Zouhiri et al. (1998) clearly showed Erythrops elegans (G.O. Sars, 1863) to be co-dominant with Anchialina agilis and Schistomysis ornata in autumn samples from the 75-m station near Plymouth, so Erythrops elegans has been added to the list of species for investigation of migratory habits in this review. Samples off Arcachon, France (Cornet et al. 1983), and off Aveiro, Portugal (Cunha et al. 1997), support the ubiquity and abundance of Anchialina agilis at shelf depths ≤125 m and the inclusion of Erythrops elegans as a frequent and abundantly caught mysid. They also support adding Leptomysis gracilis (G.O. Sars, 1864) as a 98



frequent co-dominant and occasional dominant at 52–125 m depth. Both of these sampling efforts also found relatively high abundances of Mysideis parva Zimmer, 1915 at 85–120 m depth, so this species has been added to the list of abundant species (Table 1). Anchialina agilis, Gastrosaccus spinifer, Haplostylus lobatus and H. normani are included in the list of six species provided by Macquart-Moulin & Ribera Maycas (1995), along with incontrovertible evidence of their emergence at night. As its specific name implies, Anchialina agilis is an exceptionally strong swimmer. As for most shelf mysids, direct observations are rare. Wittman (1977) described A. agilis during the day in diving depths to be inactive and colourless and to cling to leaves of Zostera without showing any reaction to natural predators or to touch by a diver. This description is incompatible with inferences from vertically resolved epibenthic sledge samples, where Anchialina agilis is usually nearly uniformly distributed among vertically arrayed nets near the bottom (e.g., Zouhiri & Dauvin 1996). Either the behaviour of this species varies spatially or they show an eventual escape response to the sledge that is strong enough to randomise their vertical distribution at the sledge mouth. Given their ubiquity, observations with a remote underwater vehicle or other camera system should be feasible to resolve their daytime behaviours beyond comfortable depths for divers. Both A. agilis and Haplostylus normani showed decreases of abundance in bottom trawls (Zouhiri & Dauvin 1996) and increases in surface plankton tows (Macquart-Moulin & Ribera Maycas 1995) at night, when Anchialina agilis showed remarkable concentration in the very surface 10 cm of the water column (Champalbert & Macquart-Moulin 1970). Few other species are so reduced in abundance in nighttime epibenthic sledge samples (Zouhiri & Dauvin 1996), prompting the conclusion that most of the A. agilis and Haplostylus normani populations undergo diel migration (Macquart-Moulin & Ribera Maycas 1995, Vallet et al. 1995). Anchialina agilis has been caught at the surface in water columns 1000 m deep, likely due to cross-isobath advection of surface waters during emergence (Macquart-Moulin & Patriti 1993, Macquart-Moulin & Ribera Maycas 1995), yet it is difficult from extant data to exclude the possibility of an active horizontal component to the migration as a contributor (Macquart-Moulin & Ribera Maycas 1995). Suggestive of accidental expatriation is the capture in bathyal waters during the day of dead specimens (MacquartMoulin & Ribera Maycas 1995). Macquart-Moulin & Ribera Maycas (1995) also concluded that the whole population of Gastrosaccus spinifer became pelagic at night. Of the migrators discussed by Macquart-Moulin & Ribera Maycas (1995) and included under the abundance criteria in this review, Anchialinja agilis is distributed from the North Sea to the northern Mediterranean and is captured in estuaries only as stray specimens. Haplostylus lobatus and H. normani, with synonymy that has been disputed (Deprez et al. 2005), together cover a range from the Porcupine Bight to the northern Mediterranean and also are rare in estuaries. Gastrosaccus spinifer ranges from Norway, in all the seas surrounding the British Isles and into the northern Mediterranean and along the adjacent North African coast, with disjunct reports from Ivory Coast and the South Shetland Islands in the Southern Ocean (Deprez et al. 2005). It is found further into estuaries than the others but generally in reduced numbers compared with the nearby shelf (e.g., Buhl-Jensen & Fosså 1991). Macquart-Moulin & Ribera Maycas (1995) did not list Erythrops elegans, Schistomysis ornata, Leptomysis gracilis or Mysideis parva among the most abundant shelf mysids and thus did not assess their migratory capabilities. Little published information is available on migration in Erythrops elegans. Zouhiri et al. (1998) on the basis of vertically resolved epibenthic sledge samples listed it as a weak swimmer, along with Gastrosaccus spinifer and Schistomysis ornata, because it tended to be caught in the lower nets. Vallet et al. (1995), quoted in Zouhiri & Dauvin (1996), suggested that Erythrops elegans migrates on a diel cycle (up at night) from the bottom boundary layer to the surface. Mauchline (1980) found negative evidence in the Clyde Sea and Loch Etive for diel migration of S. ornata and in the Clyde for Erythrops elegans. He did not comment on migration of Mysideis 99


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parva. In an extensive series of vertical plankton tows from the continental shelf of western France over the span of 2 yr, Beaudouin (1979) reported Schistomysis ornata only from February tows near the Gironde. Sorbe (1991) reported this species to be abundant from 91 to 179 m depth off Arcachon in southwest France. Based on qualitative plankton tows over the 91-m station, he reported vertical migration in this species and suggested based on size-frequency data that the species migrates seasonally across isobaths. Zouhiri et al. (1998) found S. ornata and Erythrops elegans in high abundance in the western English Channel, but their data showing nighttime disappearance of mysids as a group from sledge samples (with crepuscular peaks in abundance) came from a station in the eastern channel. They captured both Schistomysis ornata and Erythrops elegans primarily in the lower two sampling nets of their epibenthic sledge, prompting the classification of these two species in the group of weakest swimmers. Dauvin et al. (2000) captured Schistomysis ornata only in nighttime sledge tows. Vallet & Dauvin (2001), however, captured roughly equal biomasses of this species in day and night tows with peak autumn abundances reaching 21 ind. m–2. Vallet et al. (1995), quoted in Zouhiri & Dauvin (1996), suggested that S. ornata remained planktonic all day. Kaartvedt (1989) in the abstract of his study on mysid migration in Masfjorden, Norway, listed S. ornata as a vertical migrator, but in that paper reported capture in 57 mostly nighttime Isaacs-Kidd mid-water trawls of only 10 individuals (7 in the shallowest region sampled, above a bottom 35–40 m deep and 3 singletons elsewhere) but made reference to unpublished data from Kaartvedt et al. (1988) to support the conclusion of frequent migration. Earlier observations, summarised by Tattersall (1938), suggested that a small subset of breeding individuals enters the water column at night. Also supporting a case for reduced migration in this species compared with the others already explored is its reported occurrence in the guts of relatively few fish species in the three databases queried (i.e., Gibson & Ezzi 1980, Mauchline 1980, Astthorsson, 1985), but Mauchline (1980) included several more predators, including herring (Clupea harengus), so the issue of the degree of diel migration is not well settled in terms of the fraction of the population participating, the seasonality of the phenomenon, its short-term frequency or the height above bottom at which potentially enhanced swimming activity occurs. Schistomysis ornata also appears to be able seasonally to congregate at a coastal front, presumably via horizontal migration, to take advantage of the concentrated food resources there (Dewicke et al. 2002). Mauchline (1980) cited abundant evidence of diel vertical migration in Leptomysis gracilis, a highly mobile swarmer. Subsequent observations underscore the diel migratory activities of L. gracilis (into the water column at night; e.g., Kaartvedt 1985, 1989). Zouhiri & Dauvin (1996), however, captured more individuals in epibenthic sledges at night than during the day, suggesting that part of the population stays on or returns to the bottom and that it may be more easily captured at night. They also observed this species to be concentrated in the lower nets of vertically resolved tows, nominally indicating lower activity but perhaps indicating a species-specific escape response. Very little information is available on the migratory behaviour of Mysideis parva. Elizalde et al. (1991) reported that the species was concentrated in the lower nets of their sledge samples, indicating weak swimming ability, though nocturnal activity cycles have not been ruled out. In terms of predator gut contents, it has been reported only from thornback rays (Raja clavata) (see Mauchline 1980), lending some support to the conclusion that little migration occurs. Erythrops elegans is somewhat narrowly distributed latitudinally in the North Atlantic (not reported from northern Norway, Iceland or Morocco) but it is found broadly in the northern Mediterranean. Although it is found in some fjords, it has not been reported from shallower estuaries or from salinities much below that of sea water. Schistomysis ornata occurs in the North Atlantic off Iceland, from Norway to France along the European west coast and in coastal seas surrounding the British Isles. To the east it extends into the Baltic. It is also reported from Morocco, but not from the Mediterranean. Like Leptomysis gracilis, it occurs frequently in fjords. Leptomysis gracilis is distributed from Norway south around the British Isles, through the Baltic and broadly in the 100



northern Mediterranean. It occurs frequently in fjords and in some shallower estuaries. Mysideis parva has been reported on only a few occasions and only at strictly marine sites stretching from southern Ireland to the Ionian Sea. Erythrops elegans, Schistomysis ornata and Leptomysis gracilis co-occur in western Norwegian fjords, where their depth ranges are generally narrower and shallower than the inclusive ones quoted in Table 1 (Fosså & Brattegard 1990). Erythrops elegans was collected by Fosså & Brattegard (1990) at 32–100 m depth and had a median depth of occurrence of 40 m. In the fjords, Schistomysis ornata and Leptomysis gracilis had depth ranges of 32–350 m and 32–166 m and median depths of occurrence of 66 and 89 m, respectively. The four shallowest stations in this wide-area survey were at 32, 40, 74 and 100 m depth. Even more interesting is the horizontal distribution in detailed studies of a single fjord of western Sweden (Buhl-Jensen & Fosså 1991). Erythrops elegans, Leptomysis gracilis, Mesopodopsis slabberi, Schistomysis ornata and S. spiritus all reached high abundances on one or both of the sill stations. Erythrops elegans and Mesopodopsis slabberi occurred only on the sill. Leptomysis gracilis had highest abundance on the sill but also occurred throughout the fjord. Schistomysis ornata peaked on the shelf just beyond the sill (with secondand third-ranked abundances on the sill) but was distributed across all stations except the shallowest, most upstream station (depth 33 m). Erythrops erythrophthalma (Goës, 1864) showed a similar pattern, with high abundance just outside the sill, at the inner sill station, and the highest abundance shown by any mysid (11 ind. m−2) just inshore of the sill at 72 m depth.

Northwest Atlantic shelves Regrettably, systematic, intensive epibenthic sledge sampling has not been practised so frequently on other shelves. On the west side of the Atlantic, the collection of the U.S. National Marine Fisheries Service (NMFS) from the U.S. Atlantic coast remains indisputably the most comprehensive (Wigley & Burns 1971). It includes samples with an epibenthic sledge (‘bottom skimmer’) and 11 other kinds of samplers. Although this study and a follow-up analysis (Wigley & Theroux 1981) included abundant core samples from areas further south, sledge samples were limited to the region between Nova Scotia and Long Island. The NMFS survey left no room for doubt about the single most dominant species (Wigley & Burns 1971): “N[eomysis]. americana is the most common mysid inhabiting the northeastern coastal waters of the United States and undoubtedly the most abundant mysid in the western North Atlantic Ocean. …The NMFS [National Marine Fisheries Service] collection originally contained over 2 million specimens…”. The next most abundant mysids in the NMFS collection were Erythrops erythrophthalma, with 4573 specimens, and Americamysis bigelowi (Tattersall, 1926), with 2031 specimens (Wigley & Burns 1971). No other species yielded >382 specimens. No areal or volumetric abundance estimates were attempted by Wigley & Burns (1971), who were quite sensitive to issues of sampling bias (e.g., Wigley 1967). Mauchline (1980) cited abundant evidence that Neomysis americana is a frequent migrator in shallow water but noted (p. 74) that Whiteley (1948) “found no evidence of a regular diel migration in Neomysis americana on Georges Bank where the depth at which they were living, as deep as 75 m, was greater than in the coastal regions where this species is known to migrate fairly regularly”. Brown et al. (2005) from an extensive collection of zooplankton samples documented seasonal emergence, peaking in April and May, over the period from 1995 to 1989. In most years, peak abundances captured in these plankton samples were 0.1–1 ind. m−3. The present author and coworkers have been collecting emergence-trap and acoustic data on this species in the Damariscotta River estuary, in the U.S. mid-coast region of Maine for over 5 yr and at depth ranges of 10–20 m; it undergoes diel, tidally modulated emergence from approximately late June until early November, although emergence may be weak or absent on any particular day (Abello et al. 2005, Taylor et al. 2005, P.A. Jumars, unpublished observations). Neomysis americana appears to be a strongly diel 101


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migrator in some seasons in most habitats from which it has been reported (e.g., Calliari et al. 2001). Neomysis americana also shows spectacular ability to aggregate (up to 2500 ind. m−2) in bottom-water salinity fronts of estuaries (Schiariti et al. 2006). Mauchline (1980) did not comment on migration in Americamysis bigelowi. The 10-year study of Williams (1972) left no doubt, however, that A. bigelowi migrates above the bottom in large numbers. Further support comes from its presence in the guts of Atlantic silversides (Menidia menidia) in 20 m of water off New York (Warkentine & Rachlin 1989). Mauchline (1980) cited Brunel (1979) for documentation of vertical migration of Erythrops erythrophthalma in the Gulf of St. Lawrence. More recently, Carter & Dadswell (1983) reported planktonic capture of E. erythrophthalma in the very turbid Saint John River estuary, New Brunswick, year-round, including all life stages, but they took no benthic samples. Beaudouin (1979) reported it from only three of her vertical plankton hauls off Gascogne during one winter. It is among the species reported by Astthorsson (1985) from cod (Gadus morhua) guts, but the small number of observations leaves the regularity and depth limits of its diel migratory status uncertain. When this review was written, Deprez et al. (2005) had not entered many geographic data on mid-latitude mysids outside Europe and Africa so, herein, original reports are cited instead. The natural range of Neomysis americana is from the Gulf of St. Lawrence to northeastern Florida (Williams et al. 1974). Such a broad geographic range that includes high abundances at both its extremes of latitude strongly suggests a very successful opportunist or generalist, a conclusion supported by its invasion of coastal waters and estuaries of South America, where it was first reported from Uruguay (González 1974), but has spread south at least as far as San Blas, Argentina (Orensanz et al. 2002). Some fishes have come to depend on this resource (e.g., Sardiña & Lopez Cazorla 2005a,b), and some sympatric copepod populations have declined (Hoffmeyer 2004, although she does not attribute the effect to N. americana; M.S. Hoffmeyer, personal communication). Americamysis bigelowi is known from Georges Bank southward to Florida (Wigley & Burns 1971). The species frequently co-occurs with, but in substantially lower abundance than, Neomysis americana (e.g., Allen 1984). These two species are capable of substantial carnivory (Fulton 1982). Americamysis bigelowi (as Mysidopsis bigelowi) was thought to range into the Gulf of Mexico to the Texas coast, but is replaced by a pair of closely related species in the Gulf of Mexico (Price et al. 1994): Americamysis alleni Price, Heard & Stuck, 1994 and A. stucki Price, Heard & Stuck, 1994. The former species is found in poly- and mesohaline estuaries and the surf zone to 15 m, whereas the latter species has a deeper distribution out to the shelf edge (Price et al. 1994). Deprez et al. (2005) give distributional data for European Erythrops erythrophthalma, which is found off Greenland and Svalbard, down the Norwegian coast and around the British Isles, off western France and along the shelf and slope of the northern Mediterranean. The species is found southward along the U.S. east coast as far as Delaware, peaking in abundance at 60–100 m depth in evidence from the NMFS collection (Wigley & Burns 1971), and is widespread in the Arctic basin (Petryashev 2002a,b). All three of the most abundant mysid species in the NMFS collection showed high abundance on Georges Bank, with E. erythrophthalma most abundant on its southern flank, just above the 100-m isobath. Erythrops erythrophthalma is found primarily on the middle and outer shelf (Wigley & Burns 1971, Petryashev 2002a). Both Neomysis americana and Americamysis bigelowi are abundant in estuaries (e.g., Herman 1963, Allen 1984). In the NMFS survey, peak abundances on the shelf for A. bigelowi were at 30–60 m depth (Wigley & Burns 1971). The bathymetric distribution of Neomysis americana is unusual. In grab samples from the Gulf of Maine, Wigley and Burns (1971) found this species at highest abundance from 30 to 60 m depth, noting its more common presence in grabs taken during daylight. Within Cape Cod Bay, however, N. americana apparently has a shallower abundance peak at 10–29 m depth (Maurer & Wigley 1982). In the southern United States, the species rarely is captured in benthic samples offshore (Wigley & Burns

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1971), occurring more frequently in the lower, middle and upper reaches of estuaries (e.g., Williams 1972, Zagursky & Feller 1985). This onshore shift, opposite in direction to that of many other species with temperature or latitude, led Williams et al. (1974) to suspect (sub)speciation, but morphological evidence did not support this interpretation, although the issue now merits reexamination with molecular methods (Audzijonyte & Väinölä 2005, Remerie et al. 2006). Over both the shelf and offshore shoals, within coastal embayments and near inlets, abundance peaks of N. americana are reported in winter-spring, consistent with an overwintering generation breeding then, although overwintering is not excluded as a possibility within estuaries as well (Carter & Dadswell 1983). Estuarine abundance of N. americana generally peaks in summer from the midAtlantic states northward, and the number of generations per year can increase to three in the southern part of the species range (Cowles 1930, Whiteley 1948, Hulbert 1957, Herman 1963, Hopkins 1965, Williams 1972). In South Carolina, N. americana is found year-round in subtidal estuary channels and in shallow ocean waters, and its peak populations in estuaries are shifted earlier in the year (DeLancey 1987, Johnson & Allen 2005, D.M. Allen, personal communication).

Northeast Pacific shelves Elsewhere, information is much more fragmentary. In his classic study of nearshore mysids in the surf zone of southern California, Clutter (1967) found the two most abundant mysids to be Metamysidopsis elongata Holmes, 1900 (up to about 2000 ind. m−3 at 6 m below MLLW (mean lower low water)) and Neomysis kadiakensis Ortmann, 1908 (up to about 180 ind. m−3 at 8 m below MLLW); both are swarming species. Clutter (1969) reported only a subtle upward shift in median population position of Metamysidopsis elongata at night, and he reported no observations on Neomysis kadiakensis at night. Little more has been written after Clutter’s studies about the behaviour and distribution of Metamysidopsis elongata except its essential fatty acid requirements (Kreeger et al. 1991) and its role as a prey species for juvenile white seabass (Atractoscion nobilis) (Donahoe 1997). A few observations on abundance patterns and laboratory culture of its Atlantic subspecies have appeared (Tararam et al. 1996, Gama et al. 2002, and references therein). Neomysis kadiakensis appears to follow an analogous pattern to the European Mesopodopsis slabberi in abundance along the west coast of the United States. Clutter described Neomysis kadiakensis as occurring in kelp beds as well as over open sand. This species can reach higher abundances in estuaries. Dean et al. (2005) in a salt marsh within the San Francisco estuary noted that it dominated mysid abundance there over the full year, with a spring peak in abundance at 244 ind. m−3. They measured a large net import of N. kadiakensis into the salt marsh, where instantaneous mortality was calculated as 0.29 day−1. Kringel et al. (2003) in northern Puget Sound over a muddy bottom at 20 m water depth observed coherent emergence and re-entry events of N. kadiakensis associated with estimated biovolumes of 4–5 × 103 mm3 m−3. Moreover, these nocturnal emergence events appear to have dominated the holoplankton in abundance (Kringel et al. 2003). Vertical migration of this species is widespread in Puget Sound, but in the deeper reaches individuals do not appear to migrate all the way to the sea bed (Thorne 1968). It seems unlikely that Clutter (1969) could have missed such strong nocturnal emergence, so N. kadiakensis likely differs in diel migration patterns along its range. Neomysis kadiakensis is distributed from the Gulf of Alaska to southern California (Petryashev 2005). Kim & Oliver (1989) specifically studied schooling crustaceans in regions where gray whales (Eschrichtius robustus) fed in the Bering and Chukchi Seas. In diving observations concentrated in the Bering Sea, they reported swarms of Xenacanthomysis pseudomacropsis Tattersall, 1933, Neomysis rayii Murdoch, 1885 and Exacanthomysis arctopacifica Holmquist, 1981 and sampled them with various means at depths from 3 to 24 m. Petryashev (1992) regarded E. arctopacifica

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as a junior synonym of Acanthomysis stelleri Derzhavin, 1913. In Kim & Oliver’s study, Xenacanthomysis pseudomacropsis usually dominated and reached abundances of 600 ind. m−3. Acanthomysis stelleri was one to two orders of magnitude less abundant, and Neomysis rayii was reported only from two sites at the southeast extreme of the Bering Sea, where it reached intermediate abundances between those of the other two species. Direct observations of diel migration apparently are lacking, but Wing & Bar (1977, quoted in Mauchline 1980) captured Xenacanthomysis pseudomacropsis, Acanthomysis sp. and Neomysis rayii in mid-water trawls from the Chukchi Sea. Neomysis rayii is a known winter diet component in common murres (Uria aalge) and marbled murrelets (Brachyramphus marmoratus) off southeast Alaska (Sanger 1987, DeGange 1996) and Acanthomysis spp. are listed as additional diet components of the latter (DeGange 1996), although it is not clear how far the birds make excursions toward the bottom or mysids make excursions off the bottom to effect their encounters. Neomysis rayii and Acanthomysis spp. are also taken by gray whales (Eschrichtius robustus) further to the south, off Vancouver Island (Darling et al. 1998). Neomysis rayii must have been caught often enough in plankton samples to be considered a pelagic crustacean by some (McConnaughey & McRoy 1979), but simultaneous benthic and pelagic samples over diel cycles would do much to clarify these issues for all three of the Alaskan species. Xenacanthomysis pseudomacropsis occurs from central Japan to British Columbia, Neomysis rayii shares that range and extends it to southern California, and Acanthomysis stelleri has the narrowest geographic range from northern Japan to the easternmost portion of the Aleutian peninsula. All three species extend into the Chukchi Sea but not above the latitude of Wrangel Island (Petryashev 2002a).

Northwest Pacific shelves Takahashi & Kawaguchi (1995, 1997) on the Pacific coast of northern Honshu, Japan, took isobathparallel, epibenthic sledge samples from the shallowest submerged station they could sample at the lowest level of the spring tide out to 100 m from the tide line, to about 5 m water depth. Tows were stratified by distance from shore in 10-m increments and were repeated monthly between March 1992 and January 1993. Three species clearly dominated, with the dominant varying with season and depth: Archaeomysis kokuboi Ii, 1964, A. japonica Hanamura, Jo & Murano, 1996 and Iiella ohshimai (Ii, 1964). Archaeomysis kokuboi moved with the tide to stay in the shallowest position, barely immersed, and showed a peak abundance of 511 ind. m−2 in 31 July in the 0- to 10-m range from the water’s edge. (The areal abundance estimate in the present review includes no correction for sampling efficiency but simply divides the number in the tow by its 60-m2 area.) Archaeomysis kokuboi emerged at night, expanding its distribution offshore, with the reproductively most valuable members of the population showing less tendency to do so. Archaeomysis japonica occupied the next depth stratum and did not migrate with the tides but also emerged at night. The species reached peak abundance in the 10- to 20-m interval from the tide line in June at 60 ind. m–2. Iiella ohshimai was found primarily in the deepest samples as juvenile stages but showed some shoaling of its distribution in summer and also emerged at night. Iiella ohshimai reached peak abundance of 1.3 ind. m−2 in the 20- to 30-m distance from the shoreline in August, but in other months more than half of the captured individuals were found further offshore. All three species spent daytime hours buried in the sand, and all three species showed depth segregation of life stages. The Archaeomysis kokuboi population had breeding females as its shallowest members, whereas the other two species’ populations had juveniles as their shallowest members. Archaeomysis kokuboi and A. japonica are extensively exploited as food by surf-zone fishes and, as expected from their burrowing behaviour during the day, are taken mostly at night by both benthic and pelagic fishes that converge on this environment (Takahashi et al. 1999). Moreover, adult females of A. kokuboi find spatial refuge in the extremes of shallow water (Takahashi et al. 2004).

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Yamamoto & Tominaga (2005) took epibenthic sledge samples by boat from the shallow surf zone of the Seto Inland Sea three times a month from May to August in three successive years. Iiella ohshimai and Nipponomysis ornata Ii, 1964 co-dominated, with mean densities of 1.38 and 1.28 ind. m−2, respectively. One of the three dominant fish species fed primarily on the (daytime) epibenthic species N. ornata, whereas the other two ingested Iiella ohshimai more frequently, based on gut contents of daytime-collected fishes (Yamamoto & Tominaga 2005). In a separate study of Japanese flounder (Paralichthys olivaceus) from the same region, Yamamoto et al. (2004) found them to prefer the epifaunal Nipponomysis ornata. In another site in the same general region of the Seto Inland Sea, sampled monthly by epibenthic sledge between May and October for two of the same years, N. ornata showed much greater dominance over Iiella ohshimai, and mysid abundance peaked at 250 ind. m−2 in May–June (Hanamura & Matsuoka 2003). Of the four mysid species that dominated the nearshore subtidal in these Japanese studies, only Archaeomysis kokuboi is listed by Petryashev (2005) in his biogeographic summary. He listed it as being West Pacific, low boreal. Hanamura (1997) described its geographic limits as being from central Hokkaido to Honshu in northern Japan and those of A. japonica as being shallow subtidal to 50 m from Kyushu to Hokkaido. Suh et al. (1995) described an offshore migration in A. kukuboi in eastern Korea in the afternoon, with onshore migration in the morning, but their study was done before Hanamura (1997) resolved differences between some closely related species of Archaeomysis. Hanamura et al. (1996) also confirmed the emergence of Archaeomysis japonica at night. It appears that no information on diel migration of Nipponomysis ornata is available.

Other regions In other regions, it was not possible to attempt the two-step methodology for lack of comparable abundance estimates or information on migratory behaviour or lack of both, but it is clear that mysids are ubiquitous at and beyond the range of latitudes considered here. Patagonian fjords, as but one example, contain a rich mysid fauna (Brandt et al. 1997), with evidence of diel migrations (Antezana 1999), and merit more intensive study. In many cases, neuston or plankton samples confirm a pelagic phase, but neither the connection to benthic populations nor phasing of migrations is known (e.g., Sawamto 1987). A few more of these geographically scattered reports are mentioned in the following discussions of particular issues of ecological roles of mysids and drivers of their vertical migrations. Mysid ‘umwelt’ Information in the three databases used is clearly biased geographically toward Europe and North America, but where data are available on all three aspects, there is strong association among diel migration by a mysid species, its high relative abundance among mysids in the same habitat and its use as food by fishes and other animals. There is no reason to doubt that additional data would add additional species to this list or that it could already be enlarged through other databases, but some trends already are apparent. The clear risk during migration in species with this syndrome must clearly be accompanied and outweighed by substantial fitness gains in nutrition, dispersal, reproductive encounter and other aspects of life in order for these species to be among the most abundant mysids. As but one example of these ‘other’ potential gains or reduced losses, mysids living in macrophyte beds or over dense diatom mats may be driven out by low oxygen concentrations at night (Ledoyer 1969). One other pattern is apparent immediately from the data: many of the abundant shelf species of mysids show even higher abundances in the convergence zones at estuary mouths (Figure 1) than they do on shelves, with varying seasonal penetrations and population irruptions into polyhaline, mesohaline and even oligohaline reaches of estuaries.

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Erythrops elegans Erythrops erythrophthalma Gastrosaccus spinifer Leptomysis gracilis Neomysis americana Schistomysis ornata

Coastal plains estuary with an inlet

Fjord with a sill

Americamysis alleni Mesopodopsis slabberi Neomysis americana Neomysis kadiakensis Schistomysis kervillei Schistomysis spiritus

Mesopodopsis slabberi Isobath Americamysis bigelowi

Shelf edge Increasing latitude

Figure 1 Species that meet abundance criteria in the text and that occur either in the surf zone and estuaries or on the mid-shelf and in fjords or show both patterns (Neomysis americana). Mesopodopsis slabberi apparently also invades fjords from the surf zone, whereas Americamysis bigelowi apparently invades estuaries from the mid-shelf as surmised from seasonal abundance patterns.

Multiple drivers of emergence and their differing relative importance among species, locations and times are as certain for mysids as they are for holoplankton (Pearre 2003), but strong involvement of visual predators is clear from the characteristic pattern of emergence near dusk and re-entry near dawn. What makes the emergent lifestyle unique is relative, Eulerian immobility during the benthic phase, during which overlying waters are replaced. Holoplankton arguably can get a similar subsidy in regions of high vertical shear, and euphausiids are routinely captured in epibenthic sledge samples from mid-shelf depths and deeper (e.g., Cunha et al. 1997), but it will take a particular combination of vertical shear and migration timing to remain in a region of high horizontal velocity (e.g., Barber & Smith 1981). Daytime location of mysids under the surf zone, under inlets, on sills (Figure 1) and under particular portions of estuaries (e.g., Kaartvedt 1989, Cunha et al. 1999) attests to mysid virtuosity in utilising horizontal fluxes and resisting displacement. Horizontal subsidy has been easiest to visualise over abrupt changes in topography such as seamounts (reviewed by Genin 2004). The zooplankters that happen to occur over seamounts and shoals as their downward diel migrations begin are subject to intense predation. In the mysid case, the topography can be less steep, but the process is analogous and it is the mysids rather than their food taxa that do the migrating; the large phytoplankton, other protists and small zooplankton that chance to be advected above a mysid-rich area at night are subject to intense predation. Just as predators resident on seamounts can produce patches of reduced zooplankton abundance in those waters that passed over a seamount at night (Haury et al. 2000), nocturnal, pelagic patches of high mysid abundance of diel migrators can be predicted to produce patches of reduced abundance in their holoplanktonic prey. Preferred daytime mysid habitat underlies wave-driven currents, coastal currents, tidal currents, buoyancy-driven currents and topographically driven flow convergences over shoals, narrows and sills, where resultant flows replace overlying waters at high frequency. The contrast between mysids and euphausiids in fjords (Kaartvedt et al. 1988, Kaartvedt 1989) illuminates the benefits of the mysid lifestyle. Euphausiids are clearly less capable of maintaining or returning to horizontal

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co-ordinates of high horizontal flux or of high utility as refuge from predators and so are less able to profit from them than are mysids. Mysids thus fit into a broader pattern of horizontal import subsidy but have some special adaptations that enhance their gains. Genin (2004, his ‘feed-rest’ hypothesis) noted that fishes on seamounts can rest in the bottom boundary layer or behind flow obstructions when they are not feeding. Mysids as omnivores can improve on feed-rest by feeding in the pelagic environment and both feeding and resting on the bottom. They can cause ‘trophic focusing’ (Genin 2004) even where topography is not steep. The reason that this focusing is not as evident as it could be, in turn, is that mysids clearly suffer extensive mortality from mobile decapod and fish predators that horizontally export much of the horizontal import subsidy that they gather. If mysids fed only on locally produced prey and on a narrower range of resources, or if they were not fed on as heavily, their impacts would be far easier to detect but they would be less important as stabilisers in the context of new food-web theory (Montoya et al. 2006, Rooney et al. 2006). The likelihood that they return to similar habitat but not the same location when they re-enter the benthic habitat suggests that they also lend dynamic stability to benthic community structure; a local disturbance in terms of mortality of mysids can recover literally overnight compared with the need in many other benthic species for larval recruitment to occur. Besides the typical pattern of daytime re-entry and nighttime emergence, a second indication of the importance of visual predation as a driver of emergence is release from re-entry in especially turbid waters (Carter & Dadswell 1983, Castel 1993) and association of high mysid abundances with high turbidity zones of estuaries (e.g., Kimmerer et al. 1998b, Roast et al. 2004, Schiariti et al. 2006), which often leads to shoaling of populations in upper, more turbid reaches (e.g., Hulbert 1957). This association with and benefit from turbidity may underlie the paradoxical southward shoaling of peak abundances of Neomysis americana on the continental shelf. South of Cape Cod, particularly where barrier bars and islands develop along this passive continental margin, most fine material delivered by rivers is trapped inside estuaries, and what little does get delivered has a short residence time on the shelf. Dependence by juveniles on macrophyte detritus (either algal or angiosperm) may also contribute to this southward shoaling because macroalgal substrata become scarcer southward and plant and macroalgal fragments are among the particles largely trapped in estuaries. Mysids are not absent from shallow or clear waters, but generally adopt one or more of three strategies where they cannot hide within or below turbid waters: burying themselves in the bottom, hiding in vegetation or other cover or schooling. If the reaction is to visual predation and not turbidity per se, other optical phenomena that impede image formation will also benefit mysids (e.g., image distortion through salinity or thermal microstructure) (suggested by M.J. Perry, personal communication) and wave speckle and bubble clouds in the surf zone. Some mysids do congregate at salinity fronts (Kotta & Kotta 2001b, Schiariti et al. 2006). Of course there are other potential reasons, such as enhanced resource concentrations from electrostatically induced coagulation, from salting out of organics or from frontal circulation patterns. Day-night shifts in activity level (frequency and duration of movement) and height above the bottom occur even in those species not known to migrate to surface waters, with increases typical of nighttime (Fosså 1986). In the lowermost bottom boundary layer, because of the rapidly increasing horizontal velocity and hence fluxes with increasing distance from the sea bed, even modest changes in height above the bottom can yield large differences in exposure to resources. Coming out of the sediments (for the mysid species that bury) as well as activity and height changes can also produce major changes in encounter rates with predators. Saigusa (2001) provided a highresolution method for examining activity cycles, that is, by pump sampling from a depth 50 cm below the water surface simultaneously with pump sampling 50 cm above the bottom and collecting sequential 30-min samples from both streams. Saigusa (2001), at Akkeshi on the Pacific Coast of

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Hokkaido in water ~3 m deep at high tide, Nipponomyis toriumi (Murano, 1977), in 25 days of continuous sampling, showed a strong nocturnal periodicity in capture 0.5 m off the bottom, whereas the surface pump showed weak or no diel periodicity. Both sampling streams showed tidal periodicity in capture rates. At Ushimado on the Seto Inland Sea, in water of similar depth, where only surface water was sampled, an 18-day time series showed Siriella japonica Ii, 1964 to have distinct nocturnal periodicity in capture, again modulated by tides. A great deal of diversity in tidal and diel periodicity and height of emergence above the bottom is expected of mysids across species, locations and times as there is ample modulation of both risks and benefits on diverse scales. Particularly striking in the time series analysis by Saigusa (2001) is the ubiquity of periodicity in pump capture of potential prey of mysids and thus, in potential, for their capture by mysids.

An appreciation of mysids Further evidence of mysid importance in the coastal marine economy Food-web roles Roles that mysids play in feeding their predators are increasingly recognised. A search on ‘mysi*’ and ‘feeding’ in ASFA currently returns about 900 citations of which >90% concern mysids in diets of other animals. Mysids constitute particularly large fractions in the diets of many fishes in the 3–15 cm length category. In a collection of nearly 500 beam trawl samples from the Westerschelde estuary (Hostens & Mees 1999), mysids occurred in >50% of the (mostly juvenile) fish stomachs analysed and constituted >10% of the diets of two goby species (Pomatoschistus lozanoi and P. minutus), garfish (Belone belone), two gadids (bib, Trisopterus luscus, and whiting, Merlangius merlangus), two flatfish species (Pleuronectes platessa and Platichthys flesus), herring (Clupea harengus), seabass (Dicentrarchus labrax), sea snail (Liparis liparis), hook-nose (Agonus cataphractus) and tub gurnard (Trigla lucerna). Beach seining in the southern Sea of Japan produced similar dietary prominence in the 19 fish species recovered, with 67% of individuals feeding on mysids (Inoue et al. 2003). As other prominent examples of mysid dominance in gut contents, juvenile cod (Gadus morhua) 50% cover (Wilkinson 2004). 185


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Table 1 The number of zooxanthellate corals recorded to date at selected locations in the Andaman Sea Location Andaman Islands Andaman Islands Andaman Islands Andaman and Nicobar Islands Andaman and Nicobar Islands Nicobar Islands Nicobar Islands Mergui Archipelago Mergui Archipelago West coast of Thailand West coast of Thailand

No. of coral species

No. of coral genera

57 49 187 110 203 64 ND 51 65 268 353

23 15 56 45 ND 19 40 26 ND 66 69

Reference Pillai 1972 Reddiah 1977 Turner et al. 2001 Pillai 1983 Wilkinson 2004 Reddiah 1977 Scheer 1971 Summarised by Pillai 1972 Wilkinson 2004 Phongsuwan in press Turak et al. 2005

Note: ND, not described.

Table 1 summarises the level of diversity of corals throughout the region. The values are very variable even for a single region and reflect, in the Andaman Islands for example, limited sampling in earlier studies. In several of the publications cited, corals have been designated variously as hermatypes or non-hermatypes, zooxanthellate or non-zooxanthellate and scleractinian or nonscleractinian which makes cross-comparison between studies very difficult In addition, it is likely that figures given for the Mergui Archipelago and the Nicobars are underestimates because no comprehensive surveys have been carried out here or on the northwest tip of Sumatra. While Andaman Sea reefs are some of the most diverse in the Indian Ocean they have previously been reported as being less diverse than those in the Philippines and Indonesia. Recent surveys in western Thailand, however, now report 353 coral species (Turak et al. 2005) which brings the Andaman Sea into the Indo-Pacific ‘coral triangle’ of high biodiversity centred on Indonesia to the east. Interestingly, a number of these newly recorded coral species were previously known only from the Pacific. Despite limited sampling Wallace (1999) noted at least 55 species of Acropora in the Andaman Sea, a figure which is exceeded in the Indian Ocean only in the ‘eastern Indian Ocean’ which boasts 71 species and ranks alongside the most diverse areas of the world. Of Acropora species recorded in the Andaman Sea, 51 are regarded as widespread and are present in 12 of the 29 biogeographic areas described in the Indo-Pacific.

Physiological attributes of reef corals in the Andaman Sea Early work on zooxanthellae densities in corals suggested that they were remarkably constant, ranging from 1 × 106 to 2.5 × 106 algae cm−2 (Drew 1972). Subsequent work has shown more variability particularly in corals from turbid waters such as those which characterise shallow inshore areas of the Andaman Sea. In reef-flat corals from Phuket, algal densities per square centimetre of coral tissue ranged from 0. 6 × 107 to 1.4 × 107 in Porites lutea, 0.4 × 107 to 1.8 × 107 in the faviid Goniastrea retiformis and 0.8 × 107 to 2.6 × 107 in the faviid G. aspera and agaricid Coeloseris mayeri (Brown et al. 1999). Massive corals from inshore turbid waters around Singapore (B. Goh personal communication) and Java (Suharsono & Soekarno 1983) have similar high algal densities. Non-massive reef-slope corals at Phuket also show relatively high algal densities ranging from 186


CORAL REEFS OF THE ANDAMAN SEA — AN INTEGRATED PERSPECTIVE

5 × 106 cm−2 (Psammocora digitifera) to 6 × 106 cm−2 (Mycedium elephantotus) (Satapoomin 1993). A potential influence on algal density is the nutrient content of surrounding sea water because experimental studies have shown that the density of algae increases on exposure to elevated nutrient concentrations (Stambler et al. 1991). The concentration of dissolved nutrients at Phuket is high all year round due to drainage from extensive mangrove areas to the north and oceanic upwelling off shore (Janekarn & Hylleberg 1989). Brown et al. (1999) compared levels of dissolved nutrients at Phuket with other reef sites worldwide and showed that values of nitrate, nitrite and phosphate recorded at Phuket were high, ranking alongside other inshore locations in the Florida Keys and Barbados which experienced eutrophication. In addition, particulate load is elevated in inshore waters, frequently reaching 20–40 mg l−1 (Scoffin et al. 1992, Panutrakul 1996). Particulates are derived from fine terrigeneous sediments composed of clay minerals. It is likely that these particulates, coated with bacteria and microalgae, present added potential nutrients to particulate-feeding corals in the region (Anthony 1999). As a result corals have the potential to benefit from both autotrophic and heterotrophic feeding. The algal genus Symbiodinium, which forms a symbiotic relationship with corals, is diverse and molecular analyses have shown that there is considerable variation at the level of the ribosomal RNA genes. It comprises two clades, one known as phylotype A and the other which includes phylotypes B–F (LaJeunesse 2001). Of all the phylotypes so far evaluated phylotype D appears to be the most thermotolerant (Rowan 2004). Interestingly, phylotype D is present in many shallowwater corals and zoanthids throughout the Indian Ocean (Goodson 2000, Burnett 2002) whereas this phylotype is relatively rare on the southern Great Barrier Reef and in the Caribbean (LaJeunesse et al. 2003). On extensive reef flats of the east coast of Phuket, Thailand phylotype D was identified in all six species (four genera) sampled. In most cases D was present to the exclusion of all other phylotypes apart from Acropora pulchra which contained a mixture of C and D. Interestingly, in Goniastrea aspera, which is a cosmopolitan intertidal reef species throughout the Indo-Pacific, colonies from the southern Great Barrier Reef contained phylotype C (LaJeunesse et al. 2003) although only phylotype D has been found in this species at Phuket (Goodson 2000). Since the rigours of intertidal living are similar on the Great Barrier Reef to those in Thailand it appears that something other than environmental constraints might be acting to produce this distribution pattern. Burnett (2002) believes that biogeography may play a major role in the distribution of symbionts in the zoanthid Palythoa in the Indian Ocean. If this is also true for corals, then the presence in a wide variety of coral species of the most thermotolerant algal symbiont known could have important implications for the impacts of global warming. Another noteworthy feature of corals in the Andaman Sea (in this case specifically massive corals from the west coast of Thailand) is their rate of skeletal extension, which appears to be higher than any other recorded in the Indo-Pacific. Buddemeier & Kinzie (1976) described extension rates for massive corals in optimal settings around the world as 10–15 mm yr−1 with the rate for Porites spp. being slightly higher. Lough & Barnes (2000) compared the extension rates of P. lutea from 44 Indo-Pacific reefs (29 on the Great Barrier Reef; 14 in the Hawaiian Archipelago and 1 close to Phuket, Thailand). The Thai study was carried out by Scoffin et al. (1992). Skeletal extension and calcification was significantly higher in corals from Phuket compared with those from other Indo-Pacific sites, being two to three times higher than corals from the Great Barrier Reef and at least a third higher than Hawaiian corals. According to Lough & Barnes (2000) such differences were strongly linked to sea temperature with the highest sea temperature being found in the Andaman Sea. In the original study of Thai corals Scoffin et al. (1992) showed that skeletal extension rates in P. lutea ranged from 1.35 cm yr−1 at off-shore sites to 3.25 cm yr−1 at inshore locations. The significant difference in extension rates between sites was attributed to wave energy, with linear extension decreasing along a gradient of increased hydraulic energy. Clearly environmental factors are important in controlling coral skeletogenesis but one factor which has not been 187


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taken into account above is the nutrient content of sea water. Earlier descriptions in this review of sites in the Andaman Sea have highlighted the extent of turbid, high-nutrient, sheltered inshore reef settings which may have a bearing on the remarkably high zooxanthellae densities recorded in shallow-water corals. If the synthesis of organic matrix (upon which crystals of calcium carbonate are deposited in the coral skeleton) is a product of the zooxanthellae, as earlier work on calcification suggests (Johnstone 1980), then the observed rapid skeletal extension of corals in the Andaman Sea may be a result not only of high sea temperature and reduced wave energy but also of high nutrients. It is not clear how universal these growth characteristics might be in the Andaman Sea but it is likely that, because of the similarity of environmental conditions throughout the region, it is more widespread than the west coast of Thailand. The high tidal range (2–5 m) at many sites in the Andaman Sea, which favours extensive intertidal reefs in the region, results in many coral species being exposed to rigorous environmental conditions during low spring tides. During these periods, corals will be exposed to extremely high sea temperatures, high solar radiation, and reduced salinities. To survive in these settings the corals have developed remarkable environmental tolerances to temperature, light, desiccation and salinity. Protective mechanisms range from behavioural responses which induce rapid tissue retraction (Brown et al. 2002a) to photoprotective mechanisms such as xanthophyll interconversion and biochemical defences that include antioxidant enzyme and heat-shock protein production in both coral host and symbiotic alga (Brown et al. 2002b). Not surprisingly all these mechanisms are particularly well developed in intertidal corals in the region which are capable of withstanding up to 3-h aerial exposure at low spring tides. Relatively few assays of such defences have been monitored worldwide, but in the few studies that have been carried out production of stress responses by intertidal corals from the Andaman Sea (e.g., Goniastrea aspera) are many-fold higher than subtidal species such as Montastraea faveolata from the Caribbean when exposed to similar stress levels (Downs et al. 2000, Brown et al. 2002b). It is interesting to note that although there has been marked coral bleaching in years with anomalous sea temperatures such as 1991, 1995 and 2003, the intertidal reefs around Phuket, Thailand, have shown no significant mortality (Brown & Phongsuwan 2004) — a result which is testimony to the fact that these corals are well endowed with effective environmental defences and have the potential to acclimatise and adapt to varied environments.

Natural and human influences on coral reefs in the Andaman Sea The coral reefs of the Andaman Sea have been described as some of the most diverse, extensive and least disturbed by human intervention in the Indian Ocean (Wallace & Muir 2005) particularly with respect to the Andamans and Nicobars and Myanmar (Wilkinson 2004). There is generally a low level of human interference in the region apart from Aceh in Sumatra, where damaging fishing practices have negatively affected reefs (A.H. Baird personal communication). Thai reef scientists report very few incidences of damaging fish practices in their waters (U. Satapoomin personal communication) but highlight the establishment of marine protected areas as a factor which minimises such activities. Within the Andaman Sea there are at least 100 nominated marine protected areas in the Andamans and Nicobars, 2 in Myanmar, 13 in Thailand, 2 on the northwest coast of peninsular Malaysia and 1 on the northwest tip of Sumatra (Spalding et al. 2001). Enforcement procedures, however, are not rigorous in the majority of these designated areas. Despite a low level of human influences shallow reefs are subject to natural disturbances which include exposure to high solar radiation during aerial exposure, lowered sea level during Indian Ocean Dipole events, decreased salinity from heavy rain at low tide, tsunami-related damage and volcanic uplift. There is evidence of marked partial mortality as a result of exposure of faviid 188


CORAL REEFS OF THE ANDAMAN SEA — AN INTEGRATED PERSPECTIVE

colonies to high solar radiation at low tide on reefs in the Andamans, west coast of Thailand and Malaysia (Brown et al. 1994) and this phenomenon, which is induced by solar bleaching, is probably widespread throughout the region. Unusually low spring tides also cause coral mortality with sealevel depressions of 1994 and 1997–1998 (>20 cm experienced over 8–10 months) resulting in considerable coral mortality on shallow reefs of western Thailand (Brown & Phongsuwan 2004). Again, these events affect reefs throughout the region, particularly those on the northwest coast of Sumatra where the sea-level depression of 1997–1998 exceeded 30 cm for several months (Webster et al. 1999). This review has already highlighted the fact that earthquakes and resultant tsunamis have been a recurrent historic feature of the Andaman Sea although the 26 December 2004 earthquake was unprecedented in scale. Uplifted reef flats (raised by 1–2 m) were reported as a result of the earthquake on the southwest coast of Simeulue, off the west coast of Sumatra (Sieh 2005) and also in the Andaman Islands (Searle 2006). Reefs uplifted by the earthquake on the west coast of islands along the 1000-km arc from Sumatra to the northern most Andamans represent a significant loss to the gene pool of physiologically resistant corals in this area with many thousands of colonies now well above the high-tide mark. A rapid assessment survey of the 700-km coastline of west Thailand (Department of Marine and Coastal Resources 2005) 4 days after the earthquake revealed very little damage to coral reefs as a result of tsunami waves. Of the 174 sites visited up to 105 were unaffected or showed very little damage, with 30 sites displaying low-level damage (11% of coral cover affected). A further 16 sites showed moderate damage (31–50% cover affected) while 23 sites were severely damaged (>50% coral cover affected). The type of damage fell into three categories: (1) overturned massive corals, (2) broken branching corals and (3) sedimentation effects. The northernmost coastline and its off-shore islands were more severely impacted than the south — apart from Phi Phi Island, with shallow reefs on wave-exposed islands and shorelines that are most vulnerable to wave-induced damage. Similar results were obtained in Aceh although at this location deeper-living massive Porites corals, with poor attachment to the substratum, were dislodged while many shallow-living colonies were unaffected (Baird et al. 2005). At damaged locations in Thailand it is predicted that recovery will be relatively rapid (5–10 yr) based on monitoring of reefs affected by storm surges in 1986 (Phongsuwan 1991, Satapoomin et al. 2006). Reports of human damage to reefs are restricted to the Thai coastline because the majority of research has been carried out here. They include the effects of tin-smelting operations (Brown & Holley 1982), tin dredging (Chansang et al. 1992), land reclamation and associated dredging (Brown et al. 1990, Clarke et al. 1993) and tourism-related activities. While all these perturbations may cause localised damage, from which coral reefs may recover once the environmental stressor is removed, global warming represents a much more serious threat to all reefs in the Andaman Sea. Already a significant increase in sea temperature has been noted in the Andaman Sea over the last 50 yr (Brown et al. 1996), leading to claims that by the late 1990s coral bleaching might be seen on an annual basis in the region (Hoegh Guldberg 1999). In fact, this has not been the case because the two most severe bleaching events were in 1991 and 1995 in Thai waters. While the effects of global warming cannot be underestimated there seems to be little evidence at present to support so gloomy a prognosis. Not only has there been no annual bleaching of corals but also there has been limited mortality on both off-shore and inshore reefs. Presently the coral reefs of the Andaman Sea are in very good condition with inshore corals displaying remarkable physical tolerances. These in-built tolerances will, however, be severely tested if sea temperatures continue to rise. Rather than being the first reefs to succumb to global warming inshore reefs in the region may be protected by both turbid waters and their broad array of environmental defences acquired over centuries. Offshore reefs in clearer waters are likely to be more susceptible to global warming, at least initially,

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because they live in a much more benign environment and have not acquired the same level of physiological defences as their inshore counterparts.

Conclusions The coral reefs of the Andaman Sea are not only extensive and largely undisturbed but also boast a relatively high diversity of corals — the levels of which are likely to be revised upward in the future because of the present limited surveillance. Such high diversity is probably a consequence of the complex geological history of the area which has led to habitat disturbance that promoted increased species diversity over time. Features highlighted in this review, such as the dynamic and complicated hydrography, the marked tidal ranges and regular sea-level anomalies, will also have resulted in corals and symbiotic algae developing remarkable environmental tolerances. It is now abundantly clear that corals thrive in turbid settings such as those in the Andaman Sea and nearby Java (Bak & Meesters 2000) where they appear to have acclimatised/adapted to living under high sediment loads. Indeed Potts & Jacobs (2000) suggested that success in turbid habitats allowed corals to radiate out to more ‘typical’ oceanic habitats. They further proposed that a variety of turbid inshore habitats have been continuously available through geological time providing ecological and evolutionary continuity as well as refugia for corals during non-optimal periods of reef growth. The Andaman Sea could well have been such a refuge in the past and indeed may act as such in the future if sea temperatures continue to rise.

Acknowledgements I acknowledge the support of staff at Phuket Marine Biological Center over the last 26 years. Thanks are given to the Natural Environment Research Council (NERC), United Kingdom, the Leverhulme Trust, United Kingdom, the Royal Society, United Kingdom, and the U.K. Department for International Development for research support. Thanks also to Peter Hunter at the National Oceanographic Centre, Southampton, for compilation of Figure 3 and to Dr. Patrick Hyder of the Meteorological Office for discussion of the oceanography of the Andaman Sea.

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THE HUMBOLDT CURRENT SYSTEM OF NORTHERN AND CENTRAL CHILE OCEANOGRAPHIC PROCESSES, ECOLOGICAL INTERACTIONS AND SOCIOECONOMIC FEEDBACK MARTIN THIEL1,2, ERASMO C. MACAYA1, ENZO ACUÑA1, WOLF E. ARNTZ3, HORACIO BASTIAS1, KATHERINA BROKORDT1,2, PATRICIO A. CAMUS4,5, JUAN CARLOS CASTILLA5,6, LEONARDO R. CASTRO7, MARITZA CORTÉS1, CLEMENT P. DUMONT1,2, RUBEN ESCRIBANO8, MIRIAM FERNANDEZ5,9, JHON A. GAJARDO1, CARLOS F. GAYMER1,2, IVAN GOMEZ10, ANDRÉS E. GONZÁLEZ1, HUMBERTO E. GONZÁLEZ10, PILAR A. HAYE1,2, JUAN-ENRIQUE ILLANES1, JOSE LUIS IRIARTE11, DOMINGO A. LANCELLOTTI1, GUILLERMO LUNA-JORQUERA1,2, CAROLINA LUXORO1, PATRICIO H. MANRIQUEZ10, VÍCTOR MARÍN12, PRAXEDES MUÑOZ1, SERGIO A. NAVARRETE5,9, EDUARDO PEREZ1,2, ELIE POULIN12, JAVIER SELLANES1,8, HECTOR HITO SEPÚLVEDA13, WOLFGANG STOTZ1, FADIA TALA1, ANDREW THOMAS14, CRISTIAN A. VARGAS15, JULIO A. VASQUEZ1,2 & J.M. ALONSO VEGA1,2 1

Facultad de Ciencias del Mar, Universidad Católica del Norte, Larrondo 1281, Coquimbo, Chile E-mail: [email protected] 2Centro de Estudios Avanzados en Zonas Áridas (CEAZA), Coquimbo, Chile 3Alfred Wegener Institute for Polar and Marine Research, Columbusstrasse, 27568 Bremerhaven, Germany 4Facultad de Ciencias, Universidad Católica de la Santísima Concepción, Casilla 297, Concepción, Chile 5Center for Advanced Studies in Ecology and Biodiversity (CASEB), Santiago, Chile 6Facultad de Ciencias Biológicas, Pontificia Universidad Católica de Chile, Casilla 114-D, Santiago, Chile 7Laboratorio de Oceanografía Pesquera y Ecología Larval (LOPEL), Departamento de Oceanografía, Universidad de Concepción, Concepción, Chile 8Center for Oceanographic Research in the Eastern South Pacific (COPAS), Departamento de Oceanografía, Facultad de Recursos Naturales y Oceanografía, Universidad de Concepción, Estación de Biología Marina, PO Box 42, Dichato, Chile 9Coastal Marine Research Station, Departamento de Ecología, Pontificia Universidad Católica de Chile, Casilla 114D, Santiago, Chile 10Instituto de Biología Marina, Universidad Austral de Chile, PO Box 567, Campus Isla Teja, Valdivia, Chile 11Instituto de Acuicultura, Universidad Austral de Chile, PO Box 1327, Puerto Montt, Chile 12Departamento de Ciencias Ecológicas, Facultad de Ciencias, Universidad de Chile, Las Palmeras 3425, Casilla 653, Nuñoa, Santiago, Chile 13Departamento de Geofísica, Facultad de Ciencias Físicas y Matemáticas, Universidad de Concepción, Casilla 160-C, Concepción, Chile 14School of Marine Sciences, University of Maine, Orono, Maine 04469-5741, U.S. 15Unidad de Sistemas Acuáticos, Centro de Ciencias Ambientales EULA, Universidad de Concepción, Concepción, Chile 195


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Abstract The Humboldt Current System (HCS) is one of the most productive marine ecosystems on earth. It extends along the west coast of South America from southern Chile (~42°S) up to Ecuador and the Galapagos Islands near the equator. The general oceanography of the HCS is characterised by a predominant northward flow of surface waters of subantarctic origin and by strong upwelling of cool nutrient-rich subsurface waters of equatorial origin. Along the coast of northern and central Chile, upwelling is localised and its occurrence changes from being mostly continuous (aseasonal) in northern Chile to a more seasonal pattern in southern-central Chile. Several important upwelling centres along the Chilean coast are interspersed with long stretches of coast without or with sporadic and less intense upwelling. Large-scale climatic phenomena (El Niño Southern Oscillation, ENSO) are superimposed onto this regional pattern, which results in a high spatiotemporal heterogeneity, complicating the prediction of ecological processes along the Chilean coast. This limited predictability becomes particularly critical in light of increasing human activities during the past decades, at present mainly in the form of exploitation of renewable resources (fish, invertebrates and macroalgae). This review examines current knowledge of ecological processes in the HCS of northern and central Chile, with a particular focus on oceanographic factors and the influence of human activities, and further suggests conservation strategies for this high-priority large marine ecosystem. Along the Chilean coast, the injection of nutrients into surface waters through upwelling events results in extremely high primary production. This fuels zooplankton and fish production over extensive areas, which also supports higher trophic levels, including large populations of seabirds and marine mammals. Pelagic fisheries, typically concentrated near main upwelling centres (20–22°S, 32–34°S, 36–38°S), take an important share of the fish production, thereby affecting trophic interactions in the HCS. Interestingly, El Niño (EN) events in northern Chile do not appear to cause a dramatic decline in primary or zooplankton production but rather a shift in species composition, which affects trophic efficiency of and interactions among higher-level consumers. The low oxygen concentrations in subsurface waters of the HCS (oxygenminimum zone, OMZ) influence predator-prey interactions in the plankton by preventing some species from migrating to deeper waters. The OMZ also has a strong effect on the bathymetric distribution of sublittoral soft-bottom communities along the Chilean coast. The few long-term studies available from sublittoral soft-bottom communities in northern and central Chile suggest that temporal dynamics in abundance and community composition are driven by interannual phenomena (EN and the extent and intensity of the OMZ) rather than by intra-annual (seasonal) patterns. Macrobenthic communities within the OMZ are often dominated in biomass by sulphide-oxidising, mat-forming bacteria. Though the contribution of these microbial communities to the total primary production of the system and their function in structuring OMZ communities is still scarcely known, they presumably play a key role, also in sustaining large populations of economically valuable crustaceans. Sublittoral hard bottoms in shallow waters are dominated by macroalgae and suspension-feeder reefs, which concentrate planktonic resources (nutrients and suspended matter) and channel them into benthic food webs. These communities persist for many years and local extinctions appear to be mainly driven by large-scale events such as EN, which causes direct mortality of benthic organisms due to lack of nutrients/food, high water temperatures, or burial under terrigenous sediments from river runoff. Historic extinctions in combination with local conditions (e.g., vicinity to upwelling centres or substratum availability) produce a heterogeneous distribution pattern of benthic communities, which is also reflected in the diffuse biogeographic limits along the coast of northern-central Chile. Studies of population connectivity suggest that species with highly mobile planktonic dispersal stages maintain relatively continuous populations throughout most of the HCS, while populations of species with limited planktonic dispersal appear to feature high genetic structure over small spatial scales. The population dynamics of most species in the HCS are further influenced by geographic variation in propagule production (apparently caused by local differences in primary production), by temporal variation in recruit supply (caused by upwelling 196


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events, frontal systems and eddies), and topographically driven propagule retention (behind headlands, in bay systems and upwelling shadows). Adults as well as larval stages show a wide range of different physiological, ecological and reproductive adaptations. This diversity in life-history strategies in combination with the high variability in environmental conditions (currents, food availability, predation risk, environmental stress) causes strong fluctuations in stocks of both planktonic and benthic resources. At present, it remains difficult to predict many of these fluctuations, which poses particular challenges for the management of exploited resources and the conservation of biodiversity in the HCS. The high spatiotemporal variability in factors affecting ecological processes and the often-unpredictable outcome call for fine-scale monitoring of recruitment and stock dynamics. In order to translate this ecological information into sustainable use of resources, adaptive and co-participative management plans are recommended. Identification of areas with high biodiversity, source and sink regions for propagules and connectivity among local populations together with developing a systematic conservation planning, which incorporates decision support systems, are important tasks that need to be resolved in order to create an efficient network of Marine Protected Areas along the coast of northern-central Chile. Farther offshore, the continental shelf and the deep-sea trenches off the Chilean coast play an important role in biogeochemical cycles, which may be highly sensitive to climatic change. Research in this area should be intensified, for which modern research vessels are required. Biodiversity inventories must be accompanied by efforts to foster taxonomic expertise and museum collections (which should integrate morphological and molecular information). Conservation goals set for the next decade can only be achieved with the incorporation of local stakeholders and the establishment of efficient administrative structures. The dynamic system of the HCS in northern-central Chile can only be understood and managed efficiently if a fluent communication between stakeholders, administrators, scientists and politicians is guaranteed.

Introduction The deep-blue colour of the water observed for a long time past gave place to a green colour, and on the whole there was a great change in the general character of the surface fauna, pointing to the nearness of a great continent, similar to what was observed off Japan and elsewhere. On November 18 [1875], the water was very green in colour, and the ship occasionally passed through large red or brown patches, which the tow-net showed to be due to immense numbers of red copepods, hyperids, and other Crustacea. Murray (1895), on approaching Valparaíso aboard H.M.S. Challenger

Such a vivid language was rarely used by John Murray to refer to the abundance of planktonic organisms in surface waters. Only for the surface plankton from the Agulhas Bank off South Africa did he employ similar colourful language, referring to “myriads of Zoeae and a few larger Megalopae”. Scientists studying the plankton ecology of the Eastern Boundary Currents (EBCs) are used to the sight of these dense accumulations of zooplankton, which are often found in sharply defined patches. It is the intense upwelling of nutrient-rich waters in the EBCs that fuels the extraordinary high primary production (PP) in the EBCs, which forms the basis of the food web supporting some of the largest fisheries of the world. However, the frequency and intensity of upwelling within the EBCs varies, mainly depending on large-scale climatic forcing, latitudinal/seasonal signals and local factors, such as the width of the shelf, coastal topography, and sources of upwelled waters (Thomas et al. 2004). Although the overall importance of upwelling in these large marine ecosystems is relatively well known, the effects of temporal and spatial variability of upwelling on the ecology and productivity of the planktonic and benthic communities remain poorly understood. Herein these effects are explored, using the Humboldt Current System (HCS), one of the most 197


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productive EBCs (Halpin et al. 2004, Montecino et al. 2005), as a model system. In the present paper knowledge on the HCS in northern-central Chile is reviewed because this area is characterised by a complex set of temporally changing and geographically variable conditions that represent a particular challenge to science and management (see also Artisanal benthic fisheries and following sections, p. 278ff). The aims of this review are 4-fold: to (1) review current knowledge, (2) reveal gaps in information, (3) indicate conservation priorities, and (4) propose future research avenues. The HCS, by some authors named the Peru-Chile Current System, extends from ~42°S up to about the equator (Montecino et al. 2005). The main oceanographic features of this system are often described as cold nutrient-rich waters being transported northward and nutrient-enriched subsurface waters upwelled along the shorelines of Ecuador, Peru and northern Chile. Occasionally the nutrient-supply engine of the HCS is interrupted by influx of warm and nutrient-depleted equatorial waters; during such events (El Niño, EN) the northward flow of cool nutrient-rich waters is suppressed and upwelling intensity is often reduced (W. Palma et al. 2006). Individual cycles of this El Niño Southern Oscillation (ENSO) last for several years, but predictability of EN events is still very limited. On shorter temporal scales, there is a seasonal (predictable) pattern of climate and oceanography at high latitudes, which becomes aseasonal (and less predictable) at mid- and low latitudes (Blanco et al. 2001, Carr et al. 2002). It can be argued that the predictability of oceanographic conditions is lowest in the mid-region of the HCS (from 18°S to about 32°S; see also Thomas et al. 2001a) because here the seasonal variations are occasionally overshadowed by the interannual ENSO cycles, whereas at low latitudes (300 km), altimeter and drifter data show two regions of maximum eddy kinetic energy, one along the Peru coast and another centred at ~30°S off Chile. This is consistent with patterns of surface temperature and satellitemeasured ocean colour that show cold waters with high pigment concentrations associated with the upwelling extending furthest offshore off Peru and off central Chile (Montecino et al. 2005), while being restricted to an extremely narrow coastal band off northern Chile (Morales et al. 1999, Thomas et al. 2001b). CTWs propagate poleward along the entire Peru and Chile coastlines, traceable to wind fluctuations in equatorial regions (Hormazábal et al. 2001). CTWs raise and lower the pycnocline/ nutricline, influencing the effectiveness of upwelling, with dominant frequencies of days to weeks off Peru (Enfield et al. 1987) and ~50-day periods off northern and central Chile (Shaffer et al. 1997, Hormazábal et al. 2001, Rutllant et al. 2004b). Ramos et al. (2006) show that variability of equatorial origin at both annual and semi-annual periods impose strong modulation on isotherm depth along the coast. CTWs appear especially energetic during EN periods and weaker during LN periods and austral winter (Shaffer et al. 1999). At the shorter time-/space scales, diurnal cycles in wind stress are important contributors to forcing along the arid northern Chilean coast, especially in summer (Rutllant et al. 1998), but become less important with increasing latitude, where

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storm-mediated variability on 3- to 7-day cycles increases (Strub et al. 1998), with a maximum in austral winter.

Coastal oceanography Coastal waters have been defined in many different ways depending upon the reason (e.g., scientific, geopolitical, international conventions, etc.) for their usage. Bowden (1983) defines them, from an oceanographic perspective, as “those on the continental shelf and the adjoining seas”. If such a definition would be used for the HCS, then the coastal strip would be rather narrow. Bottom depth in some areas of the HCS (i.e., Antofagasta) goes from less than 100 m at distances of 10 km from shore to more than 1000 m at 30 km. However, many characteristics of the plankton assemblages at those distances show that they are still largely influenced by the proximity of the coast. Thus, for the purposes of this analysis an alternative definition for coastal ocean is used: that part of the ocean where the proximity of the continents affects the circulation and ecological processes. This definition then includes coastal waters as defined by Bowden and the coastal transition zone defined by Hormazábal et al. (2004). The analysis of any oceanographic system is scale dependent (Haury et al. 1978, Rutllant & Montecino 2002). Thus, if space–time is considered as a continuum, then isolating parts of it is an observer decision, which may be influenced by its epistemological background (Ramírez 2005a) and the characteristics of the ecosystem to be studied (Marín 2000). Haury et al. (1978), in a classical article on scales of analysis in oceanography, propose two terms (mesoscale and gross scale) to refer to those scales where the effects of flow structures such as filaments, squirts, meanders and eddies (1–102 km) are dominant and where ecosystem patterns are advective (influenced by the physics of the system) and biological. Coastal upwelling is one of the main mesoscale oceanographic processes affecting the dynamics and the spatial and temporal structure of coastal ecosystems along the EBCs (Strub et al. 1998, Montecino et al. 2005). Upwelling flow structures such as filaments (Sobarzo & Figueroa 2001), squirts (Marín et al. 2003a, Marín & Delgado 2007) and shadows (Castilla et al. 2002a, Marín et al. 2003b) have been described for the Chilean coast. Herein, information about those structures is succinctly reviewed and the potential effects on the ecology of the coastal ocean in the HCS are discussed. The HCS, from the standpoint of coastal wind forcing, can be divided in two latitudinal areas near 26°S (Figueroa 2002). From 26°S to the north, meridional, upwelling-favourable winds are rather constant throughout the year; south of this latitude greater seasonality is observed. One important upwelling focus in the northern zone is the Mejillones Peninsula (23°S). Observational (Marín et al. 1993, Escribano et al. 2000, Marín et al. 2001, Olivares 2001, Sobarzo & Figueroa 2001, Escribano et al. 2002, Rojas et al. 2002, Marín et al. 2003b) and modelling studies (Escribano et al. 2004b) have shown that the dynamics of the coastal ecosystems in that area largely depend on the generation of upwelling filaments at Mejillones Peninsula. Indeed, the generation of filaments in the northern tip of the peninsula (Punta Angamos) has been identified as the main mechanism of nutrient enrichment in the surface layers (Marín & Olivares 1999). Furthermore, ‘upwelling shadows’ within Mejillones Bay, an equator-facing bay located in the northern end of the peninsula, have been dynamically linked to the generation of bifurcated filaments at Punta Angamos (Marín et al. 2003b). This shadow is an important physical structure within the bay, affecting PP (Marín et al. 2003b) and the retention of planktonic organisms (Olivares 2001). An alternative mechanism, described as an ‘upwelling trap’ by Castilla et al. (2002a) and also related to the coastal upwelling dynamics, generates higher temperatures inside Antofagasta Bay, a pole-facing bay at the southern end of the peninsula. In this case also the physically generated structure contributes to the retention of planktonic organisms. Thus, mesoscale flow features (upwelling shadows and upwelling traps),

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associated with cold-water filaments, seem to play an important role not only in relation to the biological productivity of coastal upwelling regions but also as mechanisms for the retention of coastal planktonic species. If account is taken of the fact that Mejillones Peninsula is located in the area where upwelling-favourable winds occur year-round, then those areas may constitute recurrent retention zones. Farther south, within the HCS area where upwelling is more seasonal, the Coquimbo Bay System (30°S) is another important coastal upwelling centre. This system is located within two coastal points (Punta Lengua de Vaca and Punta Pájaros) and is the site of intensive fisheries and eco-tourism (Ramírez 2005b). Filaments, including bifurcated upwelling filaments (Moraga et al. 2001), are known to generate at Punta Lengua de Vaca, contributing with cold, nutrient-high waters to the coastal system (Montecino & Quiroz 2000, Montecino et al. 2005). A recent Lagrangian study conducted within the bay (Marín & Delgado 2007) shows that the dominant equatorward flow is modulated at quasi-inertial frequencies, enhancing the coastal retention times. Furthermore, Marín et al. (2003a) have shown that cold-water squirts generate at the northern end of the bay system (near Punta Pájaros). Squirts seem to be geographically anchored to locations meeting the requirements for their generation (Strub et al. 1991). Thus, it is highly likely that indeed the squirts observed in the vicinity of Punta Pájaros are normally generated within the same area, making them a recurrent mesoscale flow feature. As a test of this idea a simple model was built using the ROMS model (Shchepetkin & McWilliams 2005) and initialised with the ROMSTOOLS package (Penven 2003). The study area was defined by longitudinal boundaries set at 74°W and 70°W and latitudinal boundaries at 32°S and 24°S (see also Figure 3). This model has been run for 2 yr in other eastern boundary systems such as the California Current System (Marchesiello et al. 2003) to obtain a statistical equilibrium condition. However, in the present test case a squirt was generated just after 17 days of integration, which closely resembles the squirt observed in an SST image obtained in January 2005 in shape, size and location (Figure 3). The interesting, and indeed serendipitous, observation resulting from this numerical exercise is that transient modes developed as perturbations (e.g., baroclinic instabilities) from climatological mean conditions generate coastal flow structures (squirts) which are recurrently found within the HCS. Satellite observations and Lagrangian drifter data (Marín & Delgado 2007) have in fact shown that this squirt is a recurrent feature in the area, reaching distances on the order of 140 km offshore. Squirt speeds, estimated both through feature-tracking analysis (Marín et al. 2003a) and Lagrangian drifters (Marín & Delgado 2007), range between 0.2 and 0.3 m s−1. Thus, considering that the lifetime of a single squirt is related to the active period of equatorward wind events, which for the area range between 3 and 7 days (Rutllant et al. 2004a), coastal organisms trapped within the squirt are likely to reach 100–200 km offshore in a period of less than a week. The conclusions that may be offered as a result of this brief, and highly condensed, analysis of the prevailing mesoscale coastal features of the HCS are nevertheless far reaching. In the first place, if these features are intrinsic to the HCS (generated as a result of coastline geometry, bottom topography and prevailing flow conditions including perturbations) and not dependent on largescale, low-frequency forcing (e.g., ENSO), then a whole new array of multiscale (nested) models are necessary to generate predictable bio-oceanographic coastal patterns for the HCS. Second, coastal upwelling areas can be described as a dynamic mosaic of nearshore retention/offshore expatriation patches or sectors. Thus, for a planktonic organism, remaining in a specific location (i.e., local populations) within the HCS coastal zone becomes a probabilistic process. If, for example, larvae are entrained within an upwelling shadow then there is a high chance that they will remain in the same sector for a period close to a week. If, on the other hand, the larvae are entrained within a squirt, then in a period close to 24 h they will be expatriated offshore. However, since the seascape is dynamic and depends upon the upwelling condition, it is not possible to ensure that a given geographic location will always act as a retention or expatriation locality. 206



26°S

73°

28°S

30°S

73°W 72°W 71°W

72°

71°

28°

28°

29°

29°

30°

30°

73° 60

0

72° 60

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71° 180 Kilometers

Figure 3 (See also Colour Figure 3 in the insert following page 344.) Result of the squirt simulation by means of the Regional Oceanic Modeling System (coloured image) for 17 January in comparison to the thermal squirt (NOAA satellite, black-and-white image) observed on 15 January 2005. Temperature coding of the coloured image goes from blue (lower temperature) to red (higher temperature) while for the NOAA image it goes from lighter to darker tones of grey. White areas in the NOAA image correspond to clouds. The study area in the model was defined by longitudinal boundaries set at 74°W and 70°W and latitudinal boundaries at 32°S and 24°S. The resolution was 1/10°, approximately 10 km, with 20 vertical levels, a minimum depth of 50 m and a smoothed bathymetry. The baroclinic time step was set to 900 s. Surface forcing (wind stress, heat fluxes, freshwater flux, SST, sea-surface salinity, and short-wave radiation) was derived from the monthly climatology found in the COADS dataset (Da Silva et al. 1994) and interpolated into the model grid. The temperature and salinity initial conditions for the model were obtained for the month of January from the ‘World Ocean Atlas 2001’ (Conkright et al. 2002). The Ekman and geostrophic velocities at the boundary were obtained combining these data with COADS winds, following Marchesiello et al. (2001), with level of no motion defined at 500 m (Marchesiello et al. 2003).

The water column chemistry The northern Chilean ocean margin (~18–30°S) presents distinctive characteristics in topography, climatic and oceanographic conditions, which modulate PP and water column chemistry. It features a comparatively low PP (for the HCS), and despite semi-permanent wind-driven upwelling some areas are considered as high-nutrient low-chlorophyll (HNLC) environments (Torres 1995, Daneri et al. 2000). These observations are in contrast to the localised prominent upwelling cells with relatively high PP, such as off Iquique (21°S), Antofagasta (23°S) and Coquimbo (30°S) (0.5–9.3 g C m−2 d−1; González et al. 1998, Daneri et al. 2000, Thomas et al. 2001a). Studies carried out in the main upwelling centres indicate that highest PP foci occur close to the coast over the narrow continental shelf and fuel major fisheries, with catches that represent 207


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40% of the annual landings of the HCS (Escribano et al. 2003 and references therein). This production also constitutes an important way of sequestering CO2 and supports a high rate of particulate organic matter (POM) exported to depth (González et al. 2000a, Pantoja et al. 2004). This material, which is partly remineralised in the water column, strengthens the oxygen-minimum zone (OMZ) and promotes biogeochemical anaerobic processes. In this sense, a sequence of mechanisms that are determined by the oceanographic conditions is regulating the chemistry of the water column and the sea bottom. These processes gain relevance in several aspects of material exchange (i.e., gas fluxes as CO2 and N2O), implying that this area could play a key role in the main global cycles (i.e., oceanic productivity, global warming, authigenic carbonatic and phosphorite mineral formation, etc.).

Oxygen distribution and relevance in organic carbon remineralisation Most research efforts have focused on specific areas where high productivity cells are recognised, but the interactions with biogeochemical processes are still poorly understood. Along the northern margin of Chile, a well-developed OMZ is observed between 100 and 500 m water depth (Blanco et al. 2001). This zone is basically a mid-water feature since the topography of the margin precludes the OMZ to impinge on a large area of the bottom, except off Antofagasta (Mejillones) where it touches the bottom from 50 m to 300 m depth. Normally the shelf is extremely narrow from southern Peru to northern Chile (10–15 km) in comparison with central Peru and southern-central Chile (40–60 km; Strub et al. 1998), where the OMZ extends over a wide area of the shelf, promoting distinct biogeochemical processes (Gutiérrez 2000, Neira et al. 2001). This characteristic of northern areas could thus affect pathways (i.e., aerobic or anaerobic) associated with organic matter (OM) degradation in the sediments, which is an important source of regenerated nutrients to the water column. Off Mejillones, a high percentage (86%) of photosynthetically produced particulated protein is being degraded within the upper 30 m of the water column (Pantoja et al. 2004), coinciding with oxygenated waters. In consequence, OM reaching the sediments at greater depths is depleted of proteins. Over the shallower shelf sediments, where also preserved fish debris and bones are found (Milessi et al. 2005), high pigment concentrations have been reported (42–100 µg g−1 of chloroplastic pigment equivalents in surface sediments; P. Muñoz et al. 2004a, 2005). Thus, there is a narrow band of inshore sediments that are enriched in fresh OM coming from the water column. The remineralisation of this material could generate an important flux of nutrients contributing to fertilisation of the water column. Similar predictions can be made for other areas of high PP along the coast of northern Chile, where upwelled waters containing preformed nutrients are enriched with recycled nutrients derived from the OM degradation in shelf sediments. The relevance of the sea floor in water column fertilisation and biological productivity as well as its relevance in the global carbon cycle has not been well examined. Some information is available for the role of sediments near the main upwelling centres, but nothing is known about the biogeochemical processes along the large extent of the margin between the main upwelling centres mentioned here. The high OM degradation rates result in CO2 supersaturated waters that, in combination with the CO2 input from upwelled waters, favour the CO2 flux from the ocean to the atmosphere. Some studies off Antofagasta (23–24°S) and Coquimbo (30°S) have measured a saturation of >200% in upwelled waters, an f CO2 up to 1000 µatm and CO2 flux ~3.9–4.0 mol C m−2 yr−1 (Torres et al. 2003 and references therein). These authors concluded that CO2 outflux is a highly variable shortterm process, depending on the intensity of the wind-driven upwelling, the depth of the upwelled waters, the OM degradation (positively correlated with the apparent oxygen utilisation, AOU) and the biological uptake. Furthermore, the high degradation rates of organic carbon by the microplankton community (dissolved organic carbon-, DOC; 1.1–21.6 µM h−1; G. Daneri unpublished data) and high respiration rates (81–481 mmol O2 m−2 d−1; R.A. Quiñones unpublished data) indicate 208



that nutrient recycling in the water column is an important process in the CO2 production, resulting in high concentrations of CO2 and outgassing when upwelled waters are warming up at the sea surface. A substantial proportion of the carbon assimilated by primary producers is reaching the bottoms via biogenic CaCO3 flux, as has been observed off Coquimbo (30°S) where sediment traps located at 2300 m water depth (~180 km off the coast) revealed that almost 40–90% of carbonate flux is associated with some species of foraminiferans characteristic of upwelling areas (H.E. González et al. 2004a, Marchant et al. 2004). These authors suggest that biogenic CaCO3 is the main pathway by which carbon is removed from the upper ocean, controlled by autochthonous and allochthonous foraminiferans, that is, large-size organisms with high sinking rates (1.5 days) and smaller organisms that are laterally advected. Therefore, part of this carbon is exported offshore, sequestered from the water column and preserved in the sediments (Hebbeln et al. 2000a,b), but it has not been clearly established what percentage of the total CO2 assimilated by primary producers this CaCO3 flux represents.

Macro- and micronutrient distribution The distribution of nutrients shows a high variability in the water column associated with upwelling pulses and mixing. High concentrations occur inshore and usually decrease in the offshore direction, followed by decreasing pigment concentrations (Escribano et al. 2003, Marín et al. 2003a, 2004a, Peñalver 2004). Off 30°S, high surface concentrations of nitrate, phosphate and silicate have been reported (~5–15, 0.5–1 and 5–8 µM, respectively; Peñalver 2004). The nutrient distribution at this latitude is also affected by the topography of the area, where several islands reduce the circulation and mixing of the water column (Peñalver 2004). In northern Chile, between 20°S and 22°S, high concentrations of nitrate were also observed near shore at 2 µM; Morales et al. 1996) and coincident with low oxygen concentrations (1000 mm yr−1; www.meteochile.cl). In general, rivers play an important role in the fluxes of trace metals, nutrient and particulate matter to coastal waters, and some of these components are considered to be important factors determining PP in the water column. For example, Fe has been proposed as a factor limiting PP in waters enriched with other macronutrients but low in pigment concentrations (Martin & Gordon 1988, Martin et al. 1993, Coale et al. 1998). Some recent studies suggest that this element should be relevant in PP off northern Chile, where low dissolved Fe concentrations have been measured (0.6–1.4 nM; R. Torres unpublished data). Additionally, other elements such as Cd and Co may also play an important role in biological processes. The concentration of dissolved Co in the water column shows a similar vertical distribution as macronutrients in an upwelling region off Peru (8–10°S), apparently controlled by biological uptake and remineralisation (Saito 2005). In the same sense, dissolved Cd in the water column for Mejillones Bay shows a typical micronutrient-like distribution (23°S; J. Valdés unpublished data). Low values of dissolved Cd (~0.4–1.6 nM) were found in the water column, while very high concentrations were measured in surface sediments (~60 µg g–1; Valdés et al. 2003). In other coastal waters of northern-central Chile, high Cd values in sediments are probably associated with biological uptake and subsequent deposition in the 209


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sediments (~30°S; 20 µm (Morales et al. 1996, Iriarte & González 2004), and high microphytoplankton abundances in the sediment record correlate positively with intense/more frequent upwelling events (higher PP) in waters of Mejillones Bay (Ortlieb et al. 2000). In contrast, small-size phytoplankton (80% of the total diatom abundance during upwelling events (González et al. 1987), and in correspondence with this, the highest biomass was concentrated in the microphytoplankton fraction (>20 µm) from winter through spring (González et al. 1989). The first studies that considered several trophic levels of the pelagic system (Peterson et al. 1988) and trophic models of carbon flux (Bernal et al. 1989) were conducted in the coastal area off Concepción during the late 1980s. These original studies supported the classical view that the upwelling areas are characterised by short food chains dominated by large chain-forming diatoms and few small clupeiform fish species or the ‘traditional food chain’ (Ryther 1969). This view has recently been challenged, highlighting the relevance of the microbial loop (Troncoso et al. 2003, 210



Vargas & González 2004) and gelatinous food web (H.E. González et al. 2004b). These trophic flows are important throughout the whole year in oceanic areas and are highly relevant during the non-productive periods (including EN events) in coastal upwelling areas. The carbon budget of the photosynthetically generated OM in the coastal areas of the HCS has been under debate for many years (Bernal et al. 1989, González et al. 1998). It is accepted nowadays that the fraction of the PP removed from the photic zone, which is highly variable on an annual basis (Hebbeln et al. 2000a), strongly depends on the various biological (internal metabolism), physical (stratification/mixing), and chemical (nutrient rich/poor waters) processes involved as well as the time of year. However, the sources of this variability, both in space and time, have been poorly analysed until recently (Morales & Lange 2004), mainly because of the lack of long-term time-series studies. In Figure 4 the main pathways of the photosynthetically generated organic carbon are depicted. Very high bacterial secondary production (BSP) has been reported in the coastal area of northern-central Chile (Troncoso et al. 2003, Cuevas et al. 2004), suggesting a tight coupling between PP and BSP. The pivotal role of bacteria is supported by the exceptionally high

Primary production 5700 Microplankton respiration

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Figure 4 Photosynthetically generated carbon and its flows (rate estimates in mgC m–2 d–1) through the more relevant biological processes in the upwelling system off Antofagasta (23°S) and Concepción (37°S).

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degradation rates of dissolved organic carbon in Concepción Bay (1–21 µM h−1; G. Daneri unpublished data) and the high BSP rates (range of 1100–2300 mg C m−2 d−1 or 19–50% of PP). Microbial community respiration rates (Eissler & Quiñones 1999) are also very high along the HCS, reaching average values of 1450 mg C m−2 d−1 (~28% of PP). Finally, both zooplankton grazing and export production (González et al. 2000b, Grünewald et al. 2002) gave values between 100 and 500 mg C m−2 d−1 (or mean values between 2% and 10% of PP). These carbon flows are more representative of the coastal upwelling systems of Antofagasta and Concepción because they are the most studied areas (from an oceanographic point of view) along the Chilean coast. Estimations of PP for the HCS along the Chilean coast are similar to those of the Peru (4000 mg C m−2 d−1; Walsh 1981) and about 2-fold higher than those of the California (1000–2500 mg C m−2 d−1; Olivieri & Chavez 2000) upwelling systems. In addition, typical upwelling values for the flow of OM through bacteria in the HCS (19–50% of PP) are well within the range (3–55% of PP) of those described for other upwelling systems in the world oceans (Ducklow 2000).

Processes affecting primary production and export processes In coastal areas of central and southern Chile there is a distinctive seasonal pattern involving the development of maxima (spring) and minima (winter) in phytoplankton biomass and PP during an annual cycle (Ahumada 1989, González et al. 1989). In contrast, high and more constant phytoplanktonic biomass and PP has been observed during an annual cycle along the northern coast off Chile (Marín et al. 1993, Rodríguez et al. 1996, Marín & Olivares 1999). Among the factors that might control the PP, light limitation (P. Montero & G. Daneri unpublished data) and Fe availability (Hutchins et al. 2002, R. Torres unpublished data) have been suggested for the Concepción and Coquimbo upwelling systems, respectively. In addition, high microzooplankton grazing might control the PP during non-upwelling conditions (i.e., winter) in Concepción Bay upwelling (Böttjer & Morales 2005). The understanding of the factors that regulate the magnitude and the variability of phytoplankton PP is quite complex in the HCS due to the geographically distinctive upwelling areas (wind stress, topography), seasonal changes (winter vs. spring) and coastal–oceanic gradients. Integrated values in Table 1 indicate that the variation in chlorophyll concentration is coherent with changes in PP; that is, estimates are lowest in the Coquimbo upwelling system and highest in Antofagasta and Concepción coastal upwelling areas. Higher chlorophyll-specific productivity at Coquimbo and Antofagasta during EN 1997–1998 (2.5–2.8 vs. 1.0–2.0) indicates that the increase in specific productivity did not result solely from a biomass decrease, but from a change in the phytoplankton size distribution (therefore in species composition), from the larger size class (microphytoplankton) to smaller size classes (pico- and nanoplankton). The intrusion of oligotrophic oceanic waters into the coastal area off Coquimbo (Shaffer et al. 1995) and Antofagasta (Iriarte & González 2004) during an EN could be a possible explanation for the low productivity and the dominance of smaller phytoplankton size fractions and their large contribution to total PP. This feature suggests that biological and physiological shifts occur at the phytoplankton species level in order to counteract the change in prevailing physical and chemical conditions in those areas (Montecino & Quiroz 2000, Pizarro et al. 2002). On the other hand, in the permanent/seasonal coastal upwelling of cold nutrient-rich waters in enclosed areas like Concepción Bay and Mejillones Bay, PP estimates increase up to 12 g C m−2 d−1. In the above-mentioned areas, the microphytoplankton fraction accounted for >50% for the total autotrophic biomass and PP. In general, the range of specific rate of productivity in the three upwelling areas could indicate that physiological factors such as nutrient supply and/or light availability may regulate the seasonal signal of productivity in those areas, whereas top-down processes such as grazing and production export might be important in removing a fraction of the generated photosynthetic carbon (Figure 4, Table 1). Furthermore, high biological 212

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Las Cruces (33°S) 1999–2000 Concepción Bay and Gulf of Arauco (36–37°S) Total: 200–21,000 Winter: 481–1600 Spring: 1770–16,670 Spring: 208–544

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Antofagasta (22–2°S) El Niño 1997-1998

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1.0

2.8

PB (mmol C mg Chl. a–1 d–1)

Winter: 48% nanoplankton Spring: 84% microphytoplankton

Spring: 65% microphytoplankton

Winter: >50%

Winter: 59% nanoplankton

60–86% microphytoplankton

48–68% nanoplankton

Dominant size class (as % of the total Chl. a)

Thalassiosira spp., Detonula pumila, Chaetoceros socialis, Chaetoceros curvisetus, Skeletonema costatum, Leptocylindrus danicus, Guinardia delicatula

Coast: Pseudo-nitzschia pseudoseriata, Chaetoceros compressus, Leptocylindrus danicus, Rhizosolenia imbricata, Ceratium tripos, Diplopsalis lenticula Oceanic: Chaetoceros coarctatus, Ch. dadayi, Ceratium contortum, C. gibberum, C. macroceros, Dinophysis rapa, Ornithocercus magnificus

Detonula pumila, Leptocylindrus danicus, Pseudo-nitzschia pseudoseriata, Prorocentrum micans

Chaetoceros spp., Thalassiosira spp., Rhisozolenia spp., Detonula pumila, Guinardia delicatula, Eucampia cornuta

Gymnodinium sp., Pseudo-nitzschia cf. delicatissima Autotrophic flagellates

Main phytoplankton taxa

González et al. 1989 Iriarte & Bernal 1990 Ahumada 1989 Daneri et al. 2000 Troncoso et al. 2003 Montecino et al. 2004 Cuevas et al. 2004

Narvaez et al. 2004

Montecino et al. 1996 Montecino & Pizarro 1995 Avaria & Muñoz 1982 Montecino & Quiroz 2000 Troncoso et al. 2003 Morales et al. 1996 Avaria et al. 1982

Pizarro et al. 2002 Iriarte et al. 2000 Ulloa et al. 2001 Troncoso et al. 2003 Iriarte & González 2004

Reference

Table 1 Range of primary productivity, chlorophyll, chlorophyll specific primary productivity rate and main phytoplankton taxa between the 22° to 37°S Humboldt Current System. Primary productivity and chlorophyll estimates are integrated to the 1% light penetration depth




MARTIN THIEL ET AL.

productivity observed along the northern and central coast of the HCS (23–37°S) is probably fertilising oceanic and oligotrophic areas through the formation of eddies and filaments that transport coastal water offshore (Marín et al. 2003a, Hormazábal et al. 2004, Letelier et al. 2004). Finally, much of what is known about upwelling dynamics, phytoplankton ecology and their role in biogeochemical cycles along the Chilean coast has been mainly gathered from selected upwelling areas (i.e., Antofagasta, Coquimbo, Concepción, Valparaíso) with scarce or no information from areas in-between.

Zooplankton consumers Zooplankton consumers are considered the link between primary producers and higher trophic levels in the pelagic realm of the world ocean. This function also implies a role to regulate the rate at which phytoplankton-C can be transferred through the food web or retained as zooplankton biomass. This rate greatly depends on population turnover rates, which vary widely among zooplankton taxa. Thus, the knowledge of the population biology of dominant species, the community structure and its variation, become key issues in understanding the ecological and biogeochemical role of zooplankton. In the HCS substantial progress has been made in the last decades about some of these issues. In the following sections we summarise the major advances in understanding population, community and ecosystem processes involving zooplankton in the upwelling region off northern Chile. Much less progress is being made in molecular, genomic and genetic biology of zooplankton in this region. However, the need for a better understanding of zooplankton ecology in the region will most likely motivate the use of molecular tools in the near future.

Biogeographic and biodiversity issues In the upwelling region off northern Chile most species or species assemblages are considered to be part of the subantarctic fauna. This fauna originates in the Austral region, but it is then advected northward. In northern Chile and off Peru, it mixes with some species of tropical and equatorial origin. The most studied taxa are Copepoda (Heinrich 1973, Vidal 1976, Hidalgo & Escribano 2001) and Euphausiacea (Antezana 1978, D. Fernández et al. 2002). Nearly 60 species of copepods have been identified, of which the dominant in the coastal zone are Calanus chilensis, Centropages brachiatus, Paracalanus parvus, Acartia tonsa, Eucalanus inermis, Oithona similis, Oncaea conifera and Corycaeus typicus. Calanus chilensis is an endemic species of the HCS (Marín et al. 1994), whereas Centropages brachiatus, Paracalanus parvus, Acartia tonsa and Oithona similis are cosmopolites and widely distributed along the Chilean coast. Eucalanus inermis is a typical tropical species, although the genus Eucalanus may contain about six species, which have not yet been clearly defined. Among euphausiids, the most abundant and endemic species of the HCS is Euphausia mucronata (Antezana 1978). This species is closely associated with upwelling centres off northern Chile (Escribano et al. 2000) and performs an extensive vertical migration into the OMZ. Euphausia eximia is another abundant species, which increases in number during EN (Antezana 1978, González et al. 2000b). Gelatinous zooplankton, on the other hand, has received much less attention, although it may form dense aggregations at times of the year in the coastal upwelling zone (Pagès et al. 2001), and it can exert a strong predation pressure on copepods (Giesecke & González 2004, H.E. González et al. 2004b, Pavez et al. 2006) and possibly on fish larvae. Species replacements seem to occur in the zooplankton associated with EN versus LN regimes (Hidalgo & Escribano 2001). Dominant species alternate between copepods (mainly Calanus and Eucalanus genera) and euphausiids during upwelling conditions and cyclopoid copepods, Paracalanus 214



and Acartia during EN events (González et al. 2002). Warm conditions during an EN may also cause reduced size of copepods at maturity (Ulloa et al. 2001). All these changes in structure of the pelagic system may have profound implications on the functioning and productivity of this region (see, e.g., Alheit & Niquen 2004) and should also be considered in future ecosystem studies.

Zooplankton and coastal upwelling in northern Chile In the coastal zone off northern Chile, wind-driven upwelling is the principal driver of biological productivity. Pelagic organisms benefit by high productivity in upwelling sites. However, upwelling zones comprise strongly variable environments and zooplankton must cope with changing conditions in this zone. The understanding of these physical and biological interactions in the water column of upwelling systems may give much insight into the key processes that control production and biological diversity of pelagic assemblages. In northern Chile, examples of physical–biological interactions taking place during coastal upwelling are restricted to particular conditions and sites (e.g., Escribano & Hidalgo 2000a, Escribano et al. 2001, 2002, Giraldo et al. 2002). Other studies (Marín et al. 2001) have shown some relevant findings that may help understanding mechanisms through which pelagic populations are able to maintain their populations in the food-rich coastal zone. Off Mejillones Peninsula (23°S), the interaction between a poleward flow and cold upwelling plumes may generate large eddies by which non-migrant plankton can be maintained nearshore. This type of circulation may act as a retention mechanism to avoid offshore advection of zooplankton (Giraldo et al. 2002). Thus, most species maintain their population in the food-rich upwelling centres being recirculated by near-surface currents. In other upwelling systems, such as the Benguela Current, zooplankton is maintained nearshore by vertical migration behaviour (Verheye et al. 1992, Roy 1998). Organisms advected offshore migrate to deep water and then are returned to shore by a compensatory flow. In northern Chile, however, vertical migration does not seem to be an important or widely used behaviour. In fact, most dominant species appear strongly constrained to the upper layer (100 m water depth) suggests higher macrofaunal standing stocks and abundances also at the shelf off central and southern areas, both for macro- (body size 0.3–2 cm) and megafauna (body size >2 cm), compared with northern Chile (Figure 11). Palma et al. (2005) reported macrofaunal biomass values for three transects covering a depth range from about 100 m to 2000 m at Antofagasta, Concepción and Chiloé (about 22°S, 36°S and 42°S, respectively). Though biomass trends with depth, beyond the OMZ, are in general unimodal at all transects, with higher values at intermediate depths and lowest within the 232



Megafaunal abundance Macrofaunal abundance

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Antofagasta (∼22°S)

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0 160

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900

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Depth (m)

Figure 11 Macro- and megafauna depth-related patterns for three transects (22°S, 36°S and 42°S) across the shelf and upper bathyal zone of the Chilean margin. Data for macrofaunal abundance, biomass and species number from Palma et al. (2005) and for megafaunal abundance and species number from E. Quiroga unpublished data).

OMZ and beyond 1350 m, average biomasses are lower off Antofagasta. Maximum values for this transect (6.9 g wet wt m−2) are reported at 518 m water depth, while values about an order of magnitude higher (60.7 g wet wt m−2) are reported off Concepción at 784 m depth. For southern Chile (~42°S) intermediate values (39.2 g wet wt m−2) are reported for a station located at 1250 m depth. This also indicates a deepening of macrofaunal biomass maxima with latitude (Figure 11). For the megafauna observed at the same three transects, though biomass values are not available, abundances in general exhibited a similar pattern to that previously explained for macrofaunal 233


MARTIN THIEL ET AL.

biomass (Palma et al. 2005). On average, pooling the data for the three transects, abundance (~500 ind. m−2) and species number (~25) peaks were located between 1000 and 1500 m depth.

Temporal patterns of variability in shelf communities As explained in previous sections of this review, the coastal zone off northern and central Chile is strongly influenced by seasonal wind-driven upwelling, giving rise to one of the areas with the highest PP rates known worldwide (Fossing et al. 1995, Daneri et al. 2000). Since water column oxygen conditions and sediment organic loadings fluctuate at both intra- (i.e., seasonal) and interannual scales (i.e., ENSO cycle), some degree of coupling with the dynamics of benthic communities is expected (Tomicic 1985, Arntz & Fahrbach 1991). This has been demonstrated for southern-central Chile (~36°S), where seasonal and interannual changes in upwelling intensity can lead to changes in bottom-water dissolved oxygen concentration, in the amount of OM reaching the bottom (Gutiérrez et al. 2000), in the quality and lability of deposited OM (Neira et al. 2001, Sellanes & Neira 2006) and in the sediment nitrogen fluxes (P. Muñoz et al. 2004b). During the last strong EN event (1997–1998), important insights were gained by examining the effects of changing environmental conditions on local bacterial, meiofaunal (Neira et al. 2001, Sellanes et al. 2003) and macrofaunal (Gutiérrez et al. 2000) communities off central Chile. The largest biomasses of the bacterial component have been observed after several years of upwelling-favourable conditions, which are associated with cold LN phases of the ENSO (V.A. Gallardo et al. unpublished data), while the bacterial biomass is effectively depressed during warm EN phases. A decreasing trend in macrofaunal density, as well as the presence of deeper-burrowing infauna, evolved toward the end of EN 1997–1998, mainly due to the decrease of the polychaete Paraprionospio pinnata (Gutiérrez et al. 2000). It appears that more oxygenated bottom waters and oxidised sediment during EN caused P. pinnata to fail in its summer recruitment. In addition, it is probable that increased competition and predation by other species have contributed to its decline. Indeed, it has been reported that during EN, many subtropical predators invade the coastal areas (Arntz et al. 1991), negatively affecting the surface-feeding polychaetes (Tarazona et al. 1996). In central Chile, P. pinnata recovered its numerical dominance only in summer 2003, i.e., 5 yr after the end of EN. Severe hypoxic and sulphidic conditions that developed during summer 2003 probably eliminated or precluded possible competitors and/or predators, triggering the explosive increase of the P. pinnata population during this period (Sellanes et al. 2007). In northern Chile (20–30°S) few datasets extending over at least 12 months are available from shallow benthic communities (20–80 m). For the time period 1990–1995, relatively high abundances of polychaetes have been reported from several stations in northern Chile (23°50′S) at water depths of 50–60 m (Carrasco 1997). This author remarked on the absence of a clear seasonal signal in abundance changes of the main polychaete species, and he suggested that the observed variations rather reflected long-term patterns. At a long-term monitoring station near 28°S, abundances of polychaetes were high in 1995 (Figure 12), comparable to those found by Carrasco (1997). High abundance, biomass and species diversity at 28°S were associated with the warm period 1993–1995 and followed by a strong decline in 1996, coincident with LN conditions, which continued during the EN 1997–1998 (Lancellotti 2002, D. Lancellotti & W. Stotz unpublished data). Gradual disappearance of spionids, as observed in Huasco in 1996 during LN (Figure 12), was also observed during the same time period in Iquique (20°S) between 9 and 30 m depth (Quiroga et al. 1999). In northern-central Chile, during EN events, increased wave activity and freshwater runoff are frequent, in contrast to calmer periods recorded during LN events. Turbulence and runoff, in a zone where rains and strong storms are uncommon, probably oxygenate the water column, resuspend OM and/or provide terrigenous material, thus favouring reproduction and settlement of macrobenthic species living below the zone of direct wave and sediment deposition impact (>20 m 234



6000

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93 94 95 96 97 98 99 2000 01 02 03 04 05 Years

Species richness

Abundance (ind m−2)

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0

Figure 12 Temporal variations in the abundance of soft-bottom macrofauna in the upper sublittoral zone (northern Chile, 28°S) and the lower sublittoral zone (central Chile, 36°S); average abundance and species richness (S) are given; data from the upper littoral zone are taken by Smith-MacIntyre grab (1995–1996) or with sediment cores by divers (1997–2005) (D.A. Lancellotti & W. Stotz unpublished data); data from the lower sublittoral zone are taken with a multicorer, and several samples from each year were pooled, thus not allowing intra-annual variation to be seen (J. Sellanes unpublished data).

water depths). Effects of EN (and other) events on the temporal variability of benthic soft-bottom communities at present are difficult to evaluate because very few long-term datasets from benthic habitats are available from the HCS along the Chilean coast. It is herein suggested that long-term monitoring programmes should be implemented, sampling on a seasonal or bimontly basis, following examples in Peru (Tarazona et al. 2003, Arntz et al. 2006, Peña et al. 2006) and the Northern Hemisphere (Frid et al. 1996, Kroencke et al. 1998, Salen-Picard et al. 2002).

Intertidal and subtidal hard-bottom communities Hard bottoms along the coast of northern-central Chile are generally restricted to a narrow fringe extending from the intertidal zone to shallow sublittoral waters. The rock substratum is composed of rock of volcanic, granitic or sedimentary origin (Fariña et al. in press). The extensive rocky shores between 18°S and 40°S are mostly exposed to strong wave action, and they are only interrupted by short stretches of sandy beaches, which increase in extent toward the south (see Sandy beaches, p. 227, Figure 9), thereby leading to a wider separation of neighbouring hardbottom environments. Communities on intertidal and subtidal hard bottoms are dominated by macroalgae and suspension-feeding animals that form large patches (occasionally extending over to neighbouring soft bottoms) or belt-like stretches (running parallel to the shore at a certain tidal level). Most intertidal and subtidal hard bottoms are covered by one or a few dominating habitatforming organisms. Patches may persist for many years at a given location (Durán & Castilla 1989, Fernández et al. 2000, Vega et al. 2005), and they offer abundant microhabitat and food for associated organisms (Moreno & Jara 1984, Vásquez et al. 1984, Núñez & Vásquez 1987, Buschmann 1990, Vásquez 1993a, López & Stotz 1997, Sepúlveda et al. 2003a,b). Here these habitat-forming 235


MARTIN THIEL ET AL.

species, their spatial and geographic extent, their temporal dynamics and their role as ecosystem engineers (EEs) will be described. Further, it will be briefly discussed how the spatiotemporal distribution of these EEs may influence local biodiversity, population dynamics and trophic interactions in hard-bottom communities along the HCS.

Habitat-forming species on hard bottoms Large kelps (Lessonia nigrescens, L. trabeculata, Macrocystis integrifolia, M. pyrifera and Durvillaea antarctica), which grow abundantly in the low intertidal and shallow subtidal zone of the HCS, have compact and complex holdfasts that offer abundant and diverse microhabitats on the rock substratum itself. Their stipes and blades reach lengths of 2.5 m (Lessonia trabeculata) up to 30 m (Macrocystis pyrifera), and they have an important effect on local hydrodynamics because they act as wave breakers and slow down currents (Graham 2004). Smaller macroalgae with shorther thalli of 5–50 cm (such as Halopteris funicularis, Glossophora kunthii, Asparagopsis armata, Corallina officinalis and Gelidium chilense) form dense carpets (turfs) that offer primary and secondary microhabitats because they act as sediment traps retaining sand and shell fragments between their thalli and stolons (López & Stotz 1997, Kelaher & Castilla 2005). A diverse group of suspension-feeding EEs on hard bottoms include polychaetes (Phragmatopoma moerchi), barnacles (Austromegabalanus psittacus), bivalves (Perumytilus purpuratus, Semimytilus algosus, Choromytilus chorus and Aulacomya ater), and ascidians (Pyura chilensis and P. praeputialis). Their matrices reach heights of 2–40 cm, offering abundant space between living individuals (Cerda & Castilla 2001) or in remaining tubes or shells of dead individuals. Matrices of these suspension feeders may also retain considerable amounts of sediments (Prado & Castilla 2006), thereby providing secondary substratum for associated organisms. Habitat-forming species compete among themselves for space on hard-bottom substrata. Several studies indicate that mussels are competitively superior over barnacles (Navarrete & Castilla 1990, 2003, Tokeshi & Romero 1995a) and can also overgrow turf algae (Wieters 2005), ascidians may outcompete mussels (Castilla et al. 2004a) or barnacles (Valdivia et al. 2005), and large kelp may recruit in and then overgrow turf algae (Camus 1994a). Superior competitors themselves may be suppressed by predators (e.g., mussels and ascidians by gastropod and seastar predators; Paine et al. 1985, Castilla 1999, Castilla et al. 2004b) and grazers (Vásquez & Buschmann 1997, Buschmann et al. 2004a). Humans, acting as top predators by removing intermediate consumers, also strongly influence the structure of hard-bottom communities in northern and central Chile (Moreno et al. 1986, Castilla 1999). Furthermore, recruitment and growth of habitat-forming species are controlled by a variety of processes (e.g., upwelling) that drive larval and food supply (Navarrete et al. 2002, Nielsen & Navarrete 2004, Wieters 2005). Following disturbances and detachment, open space on hard bottoms is quickly recolonised, starting with ephemeral algae, which subsequently are replaced by large and long-lived turf or kelp algae and suspension feeders (Durán & Castilla 1989; Valdivia et al. 2005).

Spatial and temporal dynamics of ecosystem engineers The geographic range of most macroalgae and suspension-feeding EEs extends throughout northerncentral Chile (into Peru), but not all of them have a continuous latitudinal distribution (Table 3). All EEs from hard bottoms have pelagic dispersal stages, but in the case of the macroalgae the planktonic phase is of short duration (minutes to hours). Kelps of the genera Lessonia and Macrocystis extend from southern Chile to north of 18°S, but EN events may provoke large-scale extinctions in northern Chile (see also Kelp forests, 236

237

0–15 0–30 0 0–8 0–8 0–8 0–8 0–2 0 0 0 10–20 0–20 1–20 0–10

Macrocystis integrifolia

Macrocystis pyrifera

Durvillaea antarctica

Glossophora kunthii

Halopteris funicularis

Asparagopsis armata Corallina officinalis Gelidium chilense Phragmatopoma moerchi Perumytilus purpuratus

Semimytilus algosus

Aulacomya ater

Austromegabalanus psittacus

Pyura chilensis

Pyura praeputialis

E,S,P

E,S,P

E,S,P

E,S

E

E

S,P E,S,P E,S,P E E

E,S,P

E,S,P

E

S,P

S,P

E,S

2.5

0.6

0.2

0.2

0.2

0.1

0.12 0.15 0.06 0.2 0.05

0.1

0.25

15

30

10

6

1000 x

x

x

x

x

x

x

1000

Patch persistence (years)

10

Ecosystem engineer

38,39

Table 3 Main ecosystem engineer species from intertidal and subtidal hard bottoms along the Humboldt Current System of northern and central Chile




MARTIN THIEL ET AL.

p. 240ff.). Durvillaea antarctica does not extend further north than 32°S. Patches or belts formed in the low intertidal zone by Lessonia nigrescens and Durvillaea antarctica generally have extents of several square metres up to >100 m2. Subtidal kelp forests of Lessonia trabeculata and Macrocystis spp. may extend over >1000 m2, comprising some of the largest habitat patches formed by EEs. Distances between neighbouring patches are small in the case of Lessonia spp. and Durvillaea antarctica but individual forests of Macrocystis spp. can be separated by several hundred kilometres (see also Kelp forests, p. 240ff.). Persistence of patches over time may be favoured by recruitment of new sporophytes into existing kelp patches (Santelices & Ojeda 1984). Turf algae generally form smaller patches, with individual patches rarely exceeding an area of a few square metres. Corallina officinalis and Gelidium chilense (and other turf algae, e.g., Montemaria horridula, Rhodymenia skottsbergii) form long-lived patches in the intertidal zone (López & Stotz 1997, Vásquez & Vega 2004a, Wieters 2005), and distances between neighbouring patches are relatively small (Table 3). Glossophora kunthii, Halopteris funicularis, Asparagopsis armata (and others, including Corallina officinalis) occur mainly on shallow subtidal hard bottoms where they form patches of several square metres and, while thalli may disappear during the winter, the stolons persist over several years (Vásquez et al. 2001a). Patches may extend their size or renew thalli via asexual proliferation. The polychaete Phragmatopoma moerchi forms patches of several square metres in extent in the low intertidal and shallow subtidal zone in areas with a high supply of sand and shell fragments (Sepúlveda et al. 2003b). These patches persist over several years, but disappear if renewal is reduced, either due to low larval supply or high postsettlement mortality (Zamorano et al. 1995). The barnacle Austromegabalanus psittacus forms aggregations in the low intertidal and shallow subtidal zone; patches generally are small, rarely exceeding more than a few square metres in area. This species occurs all along the coast of northern and central Chile and little is known about the temporal dynamics of individual patches. Bivalves form extensive patches of a few square metres up to >1000 m2 in area in the mid-intertidal (Perumytilus purpuratus), low intertidal (Semimytilus algosus) and subtidal zones (Choromytilus chorus, Aulacomya ater). In the absence of predators patches can persist over many years (Durán & Castilla 1989), facilitated by regular recruitment into adult patches (Alvarado & Castilla 1996). Most bivalve species have a wide latitudinal distribution, but Perumytilus purpuratus and Aulacomya ater are almost entirely absent over an extensive area in northern Chile between 23°S and 32°S (Fernández et al. 2000, personal observations), which appears to be mainly due to limited larval supply in that region (for Perumytilus purpuratus see Navarrete et al. 2005). The ascidian Pyura chilensis occurs in small patches in the shallow subtidal zone (e.g., Vásquez & Vega 2004b), while the congener P. praeputialis forms extensive belts in the low intertidal zone (Table 3). Patches of P. chilensis persist over many years at the same location (personal observations), but little is known about the population dynamics within patches. In P. praeputialis, recruitment may be most successful in the vicinity of adults (Clarke et al. 1999), thereby favouring the long-term persistence of patches. While P. chilensis has a wide geographic distribution, P. praeputialis is restricted to a small range of 70 km along the Bay of Antofagasta (23°S) in northern Chile (Castilla et al. 2000).

Macrofauna associated with ecosystem engineers on hard bottoms A wide diversity of organisms is associated with habitat-forming species on hard bottoms of northern and central Chile. Highest species richness is found in the kelp holdfasts, intermediate numbers of associated species are reported from ascidian and bivalve reefs, and turf algae harbour fewest species of associated macrofauna (Figure 13A). This relationship appears to be related to

238


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140

120

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100

100

80

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60 60 40

40

Pyrua praeputialis

Perumytilus purpuratus

Perumytilus purpuratus

Turf algae

Phragmatopoma moerchi

Gelidium chilense

Corallina officinalis

Halopteris funicularis

Durvillaea antarctica

Kelps


0


0

Lessonia nigrescens

20

Lessonia trabeculata

20

1 to 10 10 to 1000 >1000 Patch size (m2) B

Suspension feeders

A

Figure 13 (A) Species richness of macroinvertebrates associated with habitat-forming macroalgae or suspension feeders from intertidal and subtidal hard bottoms of the northern and central coast of Chile; for reasons of comparability only studies that reported at least seven phyla of associated macrofauna were considered. (B) Average species richness in biotic habitats of different patch sizes; information obtained from López & Stotz 1997, Gelcich 1999, Godoy 2000, Thiel & Vásquez 2000, Cáceres 2001, Cerda & Castilla 2001, Hernández et al. 2001, Vásquez et al. 2001b, Thiel & Ullrich 2002, Sepúlveda et al. 2003a,b, Prado & Castilla 2006.

the fact that kelp beds and ascidian and bivalve reefs have a comparatively large spatial extent while patches of turf algae rarely cover more than a few square metres (Figure 13B). A positive relationship between patch size and number of associated species has been revealed for most habitatforming species (Vásquez & Santelices 1984, Villouta & Santelices 1984, Thiel & Vásquez 2000, Hernández et al. 2001, Sepúlveda et al. 2003a,b). Several macrofauna species have been reported from a variety of different biotic habitats. Of 251 species identified from biotic substrata (see references in Figure 13), 11.6% have been found in all three types of main biotic habitats (kelps, turf algae and suspension feeder reefs), 23.5% have been found in two types and 64.9% are only reported from one type of habitat. It must be emphasised that so far no single study has compared the associated fauna among the three main types of EEs, and there is little indication that there are habitat specialists that only occur in one type of biotic substratum. For example, Hernández et al. (2001) emphasise that several of the polychaetes found in patches of the barnacle Austromegabalanus psittacus also occur in other habitats. Similarly, Sepúlveda et al. (2003b) mention that many species from surrounding habitats associate with the reef-building polychaetes Phragmatopoma moerchi. They also emphasise that these biotic substrata may serve as recruitment habitat for some organisms. Similar observations led López & Stotz (1997), who found juvenile stages of many crustaceans and molluscs in Corallina officinalis, to

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speak of a ‘transitory fauna’ in biotic substrata (see also Vásquez & Santelices 1984, Cerda & Castilla 2001). EEs may thus favour many mobile organisms that temporarily find shelter in these habitats (e.g., Vásquez et al. 2001b). In this context, Castilla et al. (2004b) reported that the intertidal ascidian Pyura praeputialis facilitates the extension of mobile macrofauna from the subtidal into the mid-intertidal zone, thereby enhancing local species richness. The main functional groups of the organisms associated with biotic habitats are suspension feeders (32.4% of all species), grazers (25.2%) and predators (23.4%) (H. Bastias & M. Thiel unpublished data). Vásquez et al. (2001b) found very similar proportions of functional groups both in kelp holdfasts and on the surrounding hard bottoms. By offering structural protection, EEs are considered to mediate species interactions and buffer the effect of physical stress, often favouring suspension feeders (Wieters 2005, Valdivia & Thiel 2006). While the role of EEs in sustaining and promoting local biodiversity on intertidal and subtidal hard bottoms has been elucidated in numerous studies during the past decades (Vásquez & Santelices 1984, Villouta & Santelices 1984, Cerda & Castilla 2001, Sepúlveda et al. 2003a,b, Castilla et al. 2004b, Prado & Castilla 2006), relatively little is known about their trophic role on exposed rocky shores of northern-central Chile. Several studies have underlined the role of kelp forests as contributors of algal biomass to neighbouring habitats (Rodríguez 2003) and as feeding grounds for fish predators that consume understory algae and kelps (Angel & Ojeda 2001) or associated fauna (Núñez & Vásquez 1987, Palma & Ojeda 2002). Fish consumers are known to play an important role in kelp food webs of northern-central Chile (Angel & Ojeda 2001, Fariña et al. in press) but little is known about the food webs in other EEs. While most studies acknowledge the importance of EEs as habitat for associated organisms, their trophic efficiency (uptake of nutrients and suspended matter, release of dissolved and particulate organic matter) and the role of associated macrofauna in the tropho-dynamics of communities on intertidal and subtidal hard bottoms have not been thoroughly studied (see also Graham et al. 2007). The high biomass and diverse assemblage of associated consumers suggest that EEs are energetic power plants that concentrate and convert food resources in a similar way to kelp, seagrass or suspension feeder reefs in other parts of the world (e.g., Asmus & Asmus 1991, Lemmens et al. 1996, Wild et al. 2004).

Kelp forests Giant kelp dominate shallow, subtidal rocky-bottom areas in temperate and cold seas down to a depth of ~40 m (Dayton et al. 1984, Harrold & Pearse 1987, Vásquez 1992, Graham et al. 2007). Kelp distribution from south Peru to central Chile is as follows: (1) intertidal rocky areas are dominated by Lessonia nigrescens, which forms belts along exposed coasts; (2) rocky subtidal environments are dominated by Lessonia trabeculata until 40 m in depth; (3) Macrocystis integrifolia forms shallow kelp beds from the intertidal zone to depths of about 15 m. In southern-central Chile, these species are gradually replaced by Durvillaea antarctica, which dominates the intertidal zone in wave-exposed areas (Hoffmann & Santelices 1997), and in subtidal areas by Macrocystis pyrifera, which occurs in both wave-exposed and –protected habitats (Buschmann et al. 2006a). While the two species from the genus Lessonia have an almost-continuous distribution along the entire Chilean continental coast, Macrocystis integrifolia has a fragmented distribution, forming patchy populations in northern Chile. In this zone Lessonia trabeculata and Macrocystis integrifolia coexist, but mixed kelp populations have segregated patterns of bathymetric distribution, M. integrifolia being more abundant in shallow areas (Vega et al. 2005). Local populations may vary from hundreds of metres to hundreds of kilometres in extent. The observed distribution patterns are the result of complex life-history strategies that interact with environmental factors such as spatial and temporal variation in water movement, nutrient availability and temperature (Buschmann et al. 2004b, V. Muñoz et al. 2004, Vega et al. 2005). 240



The kelp forest community Kelp communities are highly productive (Dayton 1985), and they harbour a high diversity and abundance of invertebrates and fishes. Kelps, especially their holdfasts, constitute feeding areas, refuges against predation and bottom currents, spawning, settlement areas and nursery sites (Vásquez & Santelices 1984, Vásquez et al. 2001c, Vásquez & Vega 2005). Below the kelp canopy a wide diversity of turf algae exists, including several Corallinales, Asparagopsis armata, Halopteris paniculata and Gelidium spp.; several species of barnacles and other sessile invertebrates (Pyura chilensis, Phragmatopoma moerchi, Aulacomya ater) are also part of the associated species sheltered by the kelp canopy (Vásquez et al. 2001b,c, Vásquez & Vega 2004a). In contrast to the Northern Hemisphere, no large predators have been reported for southeastern Pacific kelp beds (Graham et al. 2007). Instead, invertebrate predators such as the muricid snail Concholepas concholepas, seastars (Meyenaster gelatinosus, Stichaster striatus, Heliaster helianthus and Luidia magellanica), and intermediate-size coastal fishes (Cheilodactylus variegatus, Semicossyphus maculatus and Pinguipes chilensis) dominate the predator guild in kelp forests from northern-central Chile (Vásquez 1993b, Vásquez et al. 1998, 2006). These predators feed on a diverse guild of herbivores, including sea urchins (Tetrapygus niger and Loxechinus albus), gastropods (Tegula spp. and Fissurella spp.), as well as fishes (Aplodactylus punctatus, Girella laevifrons and Kyphosus analogus) (e.g., Medina et al. 2004). These herbivore species graze on kelp and associated algae, regulating their abundance and distribution (Vásquez & Buschmann 1997, Vega et al. 2005, Vásquez et al. 2006). Marine mammals widely distributed in the coastal zone of the HCS, such as sea lions Otaria flavescens and sea otters Lontra felina, also use kelp beds as feeding areas.

Population dynamics and spatial distribution of kelps in northern-central Chile The kelp species from northern and central Chile belong to the Laminariales, which have a complex life cycle with two morphologically different stages: one conspicuous stage, recognisable as kelp that produces spores (the sporophytes), and a microscopic stage comprising independent female and male plants (the gametophytes) that lead a hidden life in the benthos. Sporophytes are the product of gametic reproduction, which is triggered by environmental factors (temperature, irradiance, photoperiod, and nutrient concentrations). The sporophytes themselves are reproductive year-round, but peak spore release has been observed during winter. Since spore survival in Laminariales is short, the dispersal range of kelps is generally assumed to be quite reduced (Graham et al. 2007); if spores do not settle within a relatively short period they die (Santelices 1990a). However, spores may survive in the guts of different herbivores (Santelices & Correa 1985, Santelices & Payá 1989) or as filaments in darkness (Santelices et al. 2002). Furthermore, fertile floating plants may act as spore carriers and thereby contribute to dispersal (Macaya et al. 2005). Juvenile sporophytes of Lessonia recruit onto hard-bottom substrata during late winter–spring, and in the field are capable of producing spores after 6–8 months (Santelices & Ojeda 1984; see review by Edding et al. 1994). Interference by adult plants inhibits the intertidal recruitment of juvenile L. nigrescens in exposed habitats. Nevertheless, water movements produce whiplash effects (sensu Dayton et al. 1984) that gives protection against grazers, thereby promoting successful recruitment of sporophytes (Santelices & Ojeda 1984). In subtidal habitats abundance of grazers, currents and reproductive behaviour of two species of elasmobranchs (Schroederichthys chilensis and Psammobatis scobina) affect Lessonia trabeculata populations. Grazing modifies algal morphology, producing two morphotypes: shrub-like and tree-like morphs. Water movement affects these differentially and generates higher mortality on tree-like morphs (Vásquez 1992). Short distances between plants (or high densities) reduce the access of grazers to the holdfasts. The 241


MARTIN THIEL ET AL.

whiplash effect of fronds and stipes pushes herbivores away from the plants, reducing grazing pressure. On the other hand, spawning of egg cases of elasmobranchs on L. trabeculata ties the stipes together, thereby reducing the whiplash effect and thus permitting grazers to approach kelp plants. Additionally, this ‘tie effect’ modifies plant shape toward the tree-like morph, and plants are more easily dislodged by water movement (Vásquez 1992). Longevity of kelps from northern Chile in the field is not well known since they do not show any evident age-related structure. Nevertheless, individual Lessonia plants can survive in the field for as long as 5 yr (J.A. Vásquez personal observations), and Macrocystis integrifolia has been reported as a perennial species in northern Chile (Buschmann et al. 2004b). Several factors generate significant biomass loss in the field: grazing pressure, wave impact, and spore release, which takes place mainly during summer (Santelices & Ojeda 1984, Edding et al. 1994). Lessonia and Macrocystis populations in northern-central Chile grow throughout the year but exhibit growth peaks during spring–summer (Buschmann et al. 2004b, Tala et al. 2004). Growth patterns are modified by wave impact, quantity and quality of light, water temperature and nutrient concentration (Buschmann et al. 2004b, Vega 2005). Local factors such as intraspecific interactions (Santelices & Ojeda 1984), herbivory (Vásquez & Buschmann 1997, Vásquez et al. 2006) and coastal upwelling events (González et al. 1998, Vásquez et al. 1998) can modify seasonal patterns of abundance and distribution (see also Graham et al. 2007). Large-scale phenomena such as ENSO produce interannual variability in abundance and could eventually generate local extinctions, as observed after the EN events of 1982–1983 and 1997–1998 (Soto 1985, Tomicic 1985, Vega 2005, Vásquez et al. 2006). Major impacts of EN were observed in kelp beds from lower latitudes (18–21°S). For example, a kelp bed occupying an area of ~40 ha at 18°S during the 1970s (IFOP 1977) disappeared as a consequence of EN 1982–1983 (Soto 1985) and has not recovered since. Similarly, during EN 1997–1998, the density of adult sporophytes on subtidal hard bottoms at 21°S decreased rapidly and linearly with increasing positive thermal anomalies (Figure 14). Six months later the site remained completely devoid of adult sporophytes, and no recolonisation occurred in the subtidal zone during the study period. In areas south of 23°S positive thermal anomalies registered during EN 1997–1998 had only limited effects on kelp beds (Figure 14). As a result, the spatiotemporal abundance patterns of M. integrifolia sporophytes in northern-central Chile is highly variable (Figure 14).

Kelp conservation and human activities Many of the diverse kelp-associated species have significant socioeconomic importance for human populations along the coast in north-central Chile and have been subject to harvesting by local human communities since pre-Columbian times (Jerardino et al. 1992, Vásquez et al. 1996). Spatial and temporal dynamics of kelp beds are significantly affected by anthropogenic impacts produced by both intense harvesting and severe pollution with organic as well as mining waste (Faugeron et al. 2005; Vásquez & Vega 2005). Lessonia nigrescens, L. trabeculata and Macrocystis integrifolia are commercially exploited between 18°S and 32°S. These species account for >95% of landings of macroalgae and basically are used for alginic acid extraction. Until 2002, collected biomass in dry weight (dry wt) amounted to ~200,000 t, almost exclusively based on stranded kelps, resulting from natural mortality of plants with holdfasts that are weakened by grazing and then detached by strong bottom currents and waves. Since 2003, however, due to international needs for raw (dry) materials and also due to increasing demands for fresh algae (to sustain aquaculture of herbivorous invertebrates in northern Chile), harvesting of natural kelp increased to ~300,000 t dry wt per year. This has led to the recent implementation of new administrative rules in order to mitigate the impact on natural kelp populations. Regulations aim at the establishment of a sustainable kelp fishery, applying the following strategies: (1) harvest management (Vásquez 1995, 1999, 2006), (2) stock 242


Intertidal kelp

5

San Marcos (21°S)

4

Camarones (18°S)

3

1998 1

2

−1

1

−3

0 5

−5 5

4

1996

1997

Constitución (23°S)

1

2

−1

1

−3

0 5

−5 5

Playa Blanca (28°S)

3

3

1

2

−1

1

−3

0

−5 5

5 4

Los Choros (29°S)

3

3

1

2

−1

1

−3

0

−5 5

5 4

San Lorenzo (30°S)

3

3

1

2

−1

1

−3

0

−5 5

5 4

2000

3

3

4

1999

5 3 1 −1 −3 −5

Thermal anomalies (°C)

Abundance of adult sporophytes (ind m−2)

5 4 3 2 1 0

Thermal anomalies (°C)

Subtidal kelp

Abundance of adult sporophytes (ind m−2)


Los Vilos (32°S)

3

3

1

2

−1

1

−3 −5

0 1996

1997

1998

1999

2000

Figure 14 Temporal variation (between 1996 and 2000) of abundances of adult sporophytes of M. integrifolia (●) and thermal anomalies estimated in situ (line) over a latitudinal gradient in northern Chile. Note: At San Marcos, an intertidal kelp population appeared after the El Niño event (▫), while the subtidal kelp bed did not recover. At Camarones (top) no sporophytes were observed during the study period. (Modified from Vega 2005)

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MARTIN THIEL ET AL.

enhancement (Vásquez & Tala 1995), (3) cultivation (Edding & Tala 2003, Westermeier et al. 2006, Gutierrez et al. 2006), and (4) conservation programmes including Marine Protected Areas (MPAs; CONAMA 2006). Considering the high variability of kelp populations in northern Chile, the limited dispersal capability of Lessonia species, and in particular the patchy distribution of beds of Macrocystis integrifolia, sustainable exploitation of natural kelp forests requires integrated management plans with continuous monitoring of standing stocks.

Export and import processes within the HCS Kelp forests, as other EEs, strongly influence trophic fluxes in benthic environments (e.g., Graham et al. 2007). In the HCS, some of the most important trophic connections occur in the vertical direction, such as dissolved nutrients released from sediments into the water column, upwelling of nutrient-rich waters from subsurface layers to the sea surface, or POM sinking from the euphotic zone toward deeper water layers and finally to the sea floor. In addition, there exist numerous types of horizontal transfer of particulate or dissolved components between the marine and the terrestrial realm, between the benthic and the pelagic environment, or between benthic habitats. In the following section the importance, intensity and frequency of these exchange processes in the HCS of northern-central Chile are considered, with a focus on coastal habitats.

Exchange between realms Exchange between the marine and the terrestrial environment occurs in both directions. Flux of materials toward the sea is via rivers, which (due to limited freshwater flow) is usually only of minor importance between 18°S and 30°S. In some areas in northern-central Chile, dissolved and solid components from human activities are continuously supplied to the marine environment, impacting local intertidal and subtidal communities (Vásquez et al. 1999, Gutiérrez et al. 2000, Lancellotti & Stotz 2004). During certain time periods (EN events or summer rains in the Andes highlands), river flow increases dramatically, transporting mainly sediments but also large quantities of terrestrial vegetation to nearshore coastal waters (Vásquez et al. 2001a). Shallow subtidal habitats along the coast of northern-central Chile are infrequently impacted by these mud flows (Miranda 2001, Vásquez et al. 2001a, Lancellotti & Stotz 2004), and it is considered likely that these impacts occur in parallel over a large geographic range. In the reverse direction, several natural exchange mechanisms are relevant in the HCS, namely, energy transfer by seabirds and marine mammals from offshore waters to coastal habitats, which occurs in a highly concentrated manner on breeding or roosting sites. Additionally, some terrestrial vertebrates (rodents, lizards, songbirds) from coastal environments forage along the drift line or in the intertidal zone (Navarrete & Castilla 1993, Fariña et al. 2003a, Sabat et al. 2003). Most of these organisms maintain relatively stable territories along the coastline, and thus material transfer from the intertidal to the upper supralittoral zone is dispersed in space, but relatively continuous in time. The same process has been reported from coastal habitats in California, where populations of insects, spiders, lizards, rodents and coyotes are mainly maintained and modulated by the food subsidy from the marine environment (Polis & Hurd 1995, 1996, Polis et al. 1997). Under certain conditions, marine resources are also transported toward the shore without the aid of biotic agents (invertebrate or vertebrate consumers defecating on land). Mortality of seabirds and marine mammals during EN events results in large numbers of animal carcasses accumulating on local beaches (e.g., Guppy 1906, Arntz & Fahrbach 1991). Similarly, during storm events, algae or benthic invertebrates are detached and cast onto the shore (González et al. 2001). These food bounties attract large numbers of terrestrial vertebrate scavengers but since supply is highly infrequent (e.g., Moore 2002), no quantitative estimates of material transfer during these events are available. 244



Fisheries also contribute to the transfer of OM from the marine toward the terrestrial realm. This not only includes direct (extraction) but also indirect forms of transfer, such as by scavenging seabirds around fishing vessels at sea (Weichler et al. 2004) or in fishing ports (Ludynia et al. 2005). Populations of kelp gulls (Larus dominicanus) near main population centres in northern-central Chile depend to a large extent on these human-derived food sources, and they then distribute remains in terrestrial environments (Ludynia et al. 2005).

Exchange between environments There is a wide range of exchange processes between the pelagic and the benthic environments. This includes, for example, supply of POM from the water column to soft bottoms where microand macroorganisms remineralise this POM, returning dissolved materials to the water column (Graf 1989, Marcus & Boero 1998, Dunton et al. 2005). Suspension feeders are important agents, which aid in transfer of suspended material (e.g., phytoplankton and kelp detritus) from the water column to the benthic system (Wolff & Alarcón 1993). In some of the bays of northern-central Chile dense stocks of natural or culture beds may significantly affect these fluxes (Uribe & Blanco 2001, Avendaño & Cantillánez 2005). The intensity and direction of transfer can be affected by regional discontinuities in the oceanographic conditions (e.g., distance from upwelling areas), which influence the transport and flux of POM and nutrients (Graco et al. 2006). These processes are also exposed to large-scale temporal variations in oceanographic conditions (e.g., ENSO cycles) (Farías et al. 2004, P. Muñoz et al. 2004b). Fisheries in small fishing ports contribute POM in the form of fish remains to soft bottoms (Sahli 2006). On hard bottoms, macroalgae and suspension feeders take up nutrients and suspended POM from the water column, returning algal remains, repackaged faeces or dissolved excretions to the water column. Most large kelps are continuously shedding senescent parts (e.g., Tala & Edding 2005). These authors estimated that annual export of shed plant detritus from a kelp forest of Lessonia trabeculata may amount to 18 kg wet wt m−2. What proportion of this detritus remains suspended in the water column or sinks immediately to the bottom is not known at present. Kelp productivity shows some seasonal variation, but kelp detritus is supplied throughout the year, at least in northern-central Chile (Tala & Edding 2005). This suspended kelp detritus may also sustain the large proportion of suspension-feeding organisms on intertidal and subtidal hard bottoms (see also Intertidal and subtidal hard-bottom communities, p. 235ff. and Bustamante & Branch 1996). Little is known about the role of DOC released by kelp forests. It enhances bacterial populations (Delille et al. 1997) and contributes to foam lines at the sea surface, which are thought to play a role in propagule dispersal and survival (Meneses 1993, Shanks et al. 2003a). Foam lines are frequently observed along the coast of northern-central Chile. Fish consumers also have an important role in energy transfer from benthic toward pelagic environments (Angel & Ojeda 2001). This transfer includes all feeding guilds of fishes from herbivores to omnivores and carnivores (Angel & Ojeda 2001). Studies of trophic coupling between hard bottoms and the water column have mainly focused on kelp forests and little is known about these trophic interactions in other hard-bottom communities (see also Intertidal and subtidal hardbottom communities, p. 235ff.).

Exchange between benthic habitats Exchange between neighbouring communities (NCs) occurs throughout the shallow subtidal zone. The form of materials exchanged between NCs and the direction of transport can be highly variable. Algal detritus exported from kelp forests contributes an important food source for animal communities on intertidal hard bottoms (Bustamante et al. 1995, Bustamante & Branch 1996, Rodríguez-Graña 245


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Table 4 Different types of material transfer between communities within the HCS of northern Chile. Distances of transport increase with increasing length of line, intensity of transfer increases with increasing size, and frequency augments with increasing number of dots. Material type

Agent

Distance

Intensity

Frequency

●

●●●●●

Lancellotti & Stotz 2004

●

●●●●●

Vásquez et al. 1999 Vásquez et al. 2000

●

●

●

●●●●●

●

●●●●●

●

●●●

Uribe & Blanco 2001

●

●●●

Bustamante & Branch 1996 Tala & Edding 2005

●

●●●●●

●

●●●●●

Between Realms TERRESTRIAL (T) –MARINE (M) Particulate inorganic matter River (T to M) (mining discharge) Dissolved metals River (T to M) (mining discharge) Particulate inorganic matter River (T to M) (river floods) Organic matter Currents (carcasses and algae) Organic matter Terrestrial vertebrates (M to T) Nutrients and dead biomass Seabirds (food) (M to T) Between Environments PELAGIC-BENTHIC POM & phytoplankton Suspension feeders POM (algal detritus)

Currents

Between Benthic Habitats NEIGHBOURING COMMUNITIES POM (detached algae) Currents Shell remains

Waves and currents

Reference

Miranda 2001 Guppy 1906 Arntz & Fahrbach 1991 Navarrete & Castilla 1993 Fariña et al. 2003a Sabat et al. 2003 Sanchez-Pinero & Polis 2000 Ludynia et al. 2005

Rodríguez 2003 Tala & Edding 2005 Bomkamp et al. 2004 Personal observations

Note: POM = particulate organic matter.

& Castro 2003). The transfer of large amounts of algal fragments from subtidal kelp forests toward the shore has been considered as a principal food source, structuring and maintaining macrofauna communities on sandy beaches (Colombini et al. 2000, Dugan et al. 2003). Transport of detached kelp plants or parts to aggregations of sea urchins in tide pools is considered to be an important trophic subsidy for these grazers (Rodríguez 2003). Arrival of kelp in the intertidal zone of northerncentral Chile continues throughout the year, but highest quantities arrive from late spring until early autumn, also depending on the proximity to source habitats (Rodríguez 2003). The importance of kelp transfer to deeper subtidal habitats (for the Californian coast see, e.g., Kim 1992, Harrold et al. 1998, Vetter & Dayton 1998, 1999) or to the rocky subtidal zone has not been evaluated in the HCS, but given that the main kelp species are non-buoyant (Lessonia spp.), it is assumed that large fractions of detached kelp may be accumulating on deeper or wave-sheltered subtidal bottoms. In addition to kelp detritus, hard-bottom communities also export large quantities of shell remains to NCs (Bomkamp et al. 2004). Along the coast of northern-central Chile, shell gravel is relatively common near exposed headlands (Ramorino & Muñiz 1970). These sediments are mainly composed of shell fragments from barnacles, sea urchins and bivalves, but source habitats, fluxes of these materials from hard bottoms to sediments and the relevance of local hydrography have not been examined.

246



Frequency and intensity of exchange processes Exchange of nutrients or particulate matter among marine communities in northern-central Chile varies in frequency and intensity (Table 4). Some of the most important and frequent exchange processes in northern Chile occur in the vertical direction (sedimentation of POM, release of nutrients into the water column, upwelling of nutrient-rich waters). Horizontal transfer processes appear to be most intense and frequent in coastal habitats, such as, for example, supply of kelp detritus to NCs. In contrast to this relatively continuous exchange of material, depositions of terrestrial sediments to coastal waters or of dead plants and animals to local beaches appear to be some of the least frequent and unpredictable transfer events. When these events occur, their intensity is often so high (e.g., Arntz 1986) that they exceed the escape or ingestion capacity of the organisms in the receiving habitats. This can result in the destruction of local communities and the incorporation of materials to deeper sediment layers. Transfer of marine-derived materials in colonies of seabirds and sea lions is also very intense (and frequent), but in northern-central Chile cannot be utilised by terrestrial organisms due to lack of water. A similar effect is observed in the water column and sediments of the OMZ where recycling processes are suppressed due to the lack of oxygen (Graco et al. 2006). Thus the intensity of the fluxes, which overcome the recycling capacity of receiving communities, may favour the long-term storage of POM not only in shelf sediments (H.E. González et al. 2004a), but also in terrestrial, intertidal and subtidal habitats of the HCS (islands with seabird and sea lion colonies, sandy beaches, subtidal kelp accumulations). It appears to be important to estimate carbon and nutrient export (and storage) not only to shelf sediments but also to terrestrial soils and intertidal and subtidal bottoms along the HCS.

Propagule supply, dispersal and recruitment variability Exchange of biological information (i.e., gene flow) depends on the dispersal ability of the organisms in question. Dispersal of individuals determines the scale at which species interact with the physical environment, the nature and consequences of the interaction with other species, the way in which they respond to perturbations and ultimately the selective forces and rates to evolve, speciate or go extinct. Because in most benthic habitats there is a predominance of species with complex life cycles, which include a free-swimming larval stage (Thorson 1950, Strathmann 1990), high dispersal capabilities are intuitively associated with most marine organisms. However, this is not a rule since coexisting with species with planktonic larvae there always exists a myriad of species with very limited dispersal potential, such as most macroalgae and direct developers or brooding species (Reed et al. 1992, Kinlan & Gaines 2003, Shanks et al. 2003b). Moreover and to further complicate things, many species use rafting as an alternative method of long-distance dispersal (Santelices 1990a, Thiel & Gutow 2005). Despite the early realisation of the high diversity of life cycles found in every marine habitat, the ecological consequences of such diversity on species interactions and on the structure and dynamics of benthic communities are only beginning to be unveiled (Kinlan & Gaines 2003, Leibold et al. 2004, Velázquez et al. 2005).

Methodological approaches to the study of dispersal The study of dispersal in the ocean is fraught with methodological problems imposed by the difficulty of following the usually microscopic propagules over extended time. Indirect methods to estimate aspects of dispersal have been developed. For instance, the use of highly variable neutral DNA markers offers an unprecedented opportunity to estimate realised dispersal distances (Palumbi

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1995, Kinlan & Gaines 2003, Palumbi 2003, Sotka & Palumbi 2006). Similarly, trace element microchemistry (elemental fingerprinting) (Swearer et al. 1999, DiBacco & Levin 2000, Zacherl et al. 2003) or stable isotope ratios (Herzka et al. 2002, Levin 2006) hold promise for identifying larval origin under specific environmental conditions. However, these techniques do not yet provide a quantitative measure of the fate of all propagules released from a focal location (i.e., the ‘dispersal kernel’). Therefore, spatially explicit connectivity among local populations, the type of information needed regarding the location of MPAs (Botsford et al. 1994, Lockwood et al. 2002, Kaplan 2006), remains tractable only through the combination of biophysical models linking larval attributes with advection-diffusion physical processes (Marín & Moreno 2002, Largier 2003, Siegel et al. 2003, Guizien et al. 2006, Kaplan 2006, Levin 2006, Aiken et al. 2007).

Studies of dispersal in HCS Studies of dispersal within the HCS are scarce at best. Santelices (1990a) provides a review of dispersal in marine seaweeds and points out that most information comes from laboratory studies conducted under idealised hydrographic conditions or from rather anecdotal evidence of colonisation of new habitats. Studies conducted by incubating seawater samples have demonstrated the existence of a multispecific ‘spore cloud’ which is present year-round in coastal waters of central Chile (Hoffmann & Ugarte 1985, Hoffmann 1987). These studies showed the patchy and temporally variable nature of the spore cloud, but the dispersal distances and mechanisms involved are unclear. Considering the small size of spores (5–150 µm) and their short duration (a few hours, but it can be up to few days; Santelices, 1990a), stochastic turbulence diffusion probably plays a major role in shaping the dispersal kernels in algal dispersal. Nearshore advective currents within the dispersal scale of spores (e.g., tidal currents, breaking waves, internal tidal bores) cannot be ruled out, however. Recent studies by Bobadilla & Santelices (2005) conducted by sampling the water column with a semi-automated sampling device (Bobadilla & Santelices 2004), illustrate the great temporal variability in multispecific dispersal kernels for major algal groups and dispersal distances exceeding 100 m. The most direct studies of dispersal of invertebrates in the HCS are restricted to species with short larval duration, such as tunicates (Castilla et al. 2002a,b, 2004a). These studies sampled larval distribution at distances from a unique adult population source. Quantitative aspects of dispersal for species with long-lived larval stages are virtually unknown for any invertebrate or coastal fish species in the HCS. Several physical processes that can increase offshore and alongshore advection, or instead increase retention of larvae near shore, have been described for the coast of Chile, such as upwelling filaments, cyclonic circulation in embayments, topographically controlled eddies and upwelling shadows and traps (Vargas et al. 1997, Marín et al. 2001, Castilla et al. 2002a, Escribano et al. 2002, Wieters et al. 2003, Narváez et al. 2004). These features undoubtedly influence dispersal of coastal species, potentially increasing self-recruitment (Swearer et al. 2002), but their effect on connectivity among adult populations of any species is hard to demonstrate. A few high-resolution 3-dimensional numerical models of currents in the coastal ocean have been developed and tested against physical data for different sections of the coast of Chile (Mesías et al. 2001, Aiken et al. 2007). Coupled with Lagrangian larval-tracking techniques these biophysical models can generate testable hypotheses about dispersal and connectivity in real biological systems (e.g., Aiken et al. 2007). There is an urgent requirement to develop highly variable, neutral molecular markers such as microsatellites for algal and invertebrate species inhabiting the HCS system to improve the ability to infer dispersal distances over ecological timescales and to test hypotheses about connectivity (see also Population connectivity, p. 252ff.) derived from theoretical models.

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Settlement studies in the HSC A more tractable and directly related problem is the settlement and recruitment of species to a given habitat. Settlement is the process through which a spore or larva makes permanent contact with the benthic habitat (Keough & Downes 1982). Since most adults of benthic organisms are sessile or have limited movement, settlement marks the end of the effective dispersal phase. For brooding organisms with mobile adults or those that use rafting as a secondary dispersal mechanism this is of course not the case (Thiel & Haye 2006). Recruitment is the input of new individuals to the benthic population measured at some arbitrary time after settlement. Therefore, while settlement is expected to reflect the arrival or supply of propagules, recruitment can be substantially modified by postsettlement mortality. Of the considerable number of studies of supply ecology conducted in the HCS over the past decades, very few have come close to measuring settlement (Hoffmann & Ugarte 1985, Hoffmann 1987, Moreno et al. 1993a, 1998, Martínez & Navarrete 2002, Vargas et al. 2004, Lagos et al. 2005, Narváez et al. 2006), and most have actually examined recruitment at varying time intervals after settlement. The paucity of studies of settlement of marine algae, due largely to the enormous difficulties of identifying sporelings to species level (Hoffmann 1987, Santelices 1990a), has not permitted the identification of mechanisms involved in algal settlement patterns. On the other hand, using intertidal barnacles, mussels and several gastropod species as model organisms, a few larval transport mechanisms have been demonstrated in the central and southern HCS. At the locality of Las Cruces in central Chile, which has been characterised as an upwelling shadow (Kaplan et al. 2003, Wieters et al. 2003, Narváez et al. 2004), the onshore daily settlement of several invertebrate species is associated with conditions that favour the occurrence of internal tidal bores, which appear to be common in the area when the water column is well stratified (Vargas et al. 2004). These results suggest that, for a number of intertidal invertebrates, internal tidal bores (Pineda 1991, 1994a) can be an important mechanism of onshore transport. In contrast with results at some sites in the California upwelling ecosystem (e.g., Farrell et al. 1991, Wing et al. 1995a), studies at Las Cruces showed that settlement of invertebrates was not directly associated with upwelling-relaxation events, which occur throughout spring and summer over synoptic timescales. The suggestion here is not that the upwelling-relaxation transport model (e.g., Roughgarden et al. 1988, 1991, Wing et al. 1995a,b) plays no role in settlement and recruitment in central Chile, but rather that at Las Cruces larval transport toward the shore does not seem to be dominated by these mechanisms. Indeed, spatially intensive studies over a region of about 120 km around Las Cruces showed a clear mesoscale spatial pattern in barnacle settlement, apparently imposed by the topographic variability in upwelling intensity (Lagos et al. 2005). Thus, the spatial variability in upwelling intensity typical of central Chile (Strub et al. 1998, Broitman et al. 2001, Halpin et al. 2004) might influence the spatial position of the larval pool and probably affects the scales of dispersal of invertebrates, as suggested also by the study of spatial synchrony in recruitment of species with contrasting dispersal potential (Lagos et al. in review). Topographic effects on patterns of settlement and recruitment have also been reported for a variety of brachyuran crab species (A.T. Palma et al. 2006). Buoyancy fronts produced by river plumes, common from about 30°S to the south in the HCS, in conjunction with wind stress can also play a role in delivering larvae to shore (Vargas et al. 2006c). Narváez et al. (2006) also report on the effect of what they called ‘large warming events’, which occurred a few times in spring–summer in association with downwelling-favourable (northerly) winds. During these specific large warming events these authors observed significant synchrony in recruitment of several invertebrate taxa (decapods, gastropods, polychaetes, mussels and sea urchins), suggesting that larvae could be entrained in these advective fronts and delivered onshore. A roughly similar phenomenon has been observed around Valdivia in southern Chile, where southward and onshore movement of warm waters produced by winter

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storms delivered larvae of several gastropod species to the shore (Marín & Moreno 2002, C.A. Moreno in a personal communication to S.A. Navarrete). Few studies have directly and simultaneously examined the distribution of larvae in the plankton, physical processes and settlement onshore in the HCS. Even fewer have examined larval behaviour under field or laboratory conditions (Poulin et al. 2002a,b, Manríquez et al. 2004, Vargas et al. 2006a).

Patterns of recruitment and benthic communities Systematic studies of recruitment of species in the HCS have focused on (1) characterising spatial and temporal variation in the arrival of new individuals for intertidal and a few subtidal species (e.g., Jara & Moreno 1983, Hoffmann & Santelices 1991, Stotz et al. 1991a, Camus & Lagos 1996, Martínez & Navarrete 2002); (2) relating these patterns with large-scale oceanographic anomalies, such as El Niño events (e.g., Moreno et al. 1998, Navarrete et al. 2002); (3) examining the effects of recruitment variability on population dynamics and recovery of local populations from physical, biological or human-induced disturbance (e.g., Santelices & Ojeda 1984, Moreno et al. 1993b, Duarte et al. 1996, Alvarado et al. 2001); (4) determining the consequences of recruitment variation on the nature and intensity of species interactions in the adult habitat (e.g., Navarrete & Castilla 1990, Moreno 1995, Navarrete et al. 2005, Wieters 2005); and (5) characterising the effects on the processes that regulate the dynamics of entire intertidal communities over large spatial scales (Navarrete et al. 2005). The far-reaching ramifications of persistent variation in recruitment have been most amply demonstrated in recent studies that quantify patterns of recruitment over large temporal (years) and spatial (tens to hundreds of kilometres) scales. These studies are starting to shed light on, and find recurrent patterns in, the causes of the typically large, baffling and usually ‘unpredictable’ variation in coastal ecosystems. Studies along the California coast have found large-scale regularities in patterns of recruitment of sessile species that can help reconcile odd experimental results (Menge et al. 1994, Connolly et al. 2001, Menge et al. 2003). Studies in the HCS conducted by Navarrete et al. (2002, 2005) have evaluated the effects of variation in wind-driven upwelling on community regulation along 900 km of coastline between 29°S and 35°S during 72 months. Sharp discontinuities in upwelling regimes around 30–32°S produced abrupt and persistent breaks in the dynamics of benthic and pelagic communities over hundreds of kilometres (regional scales) (Figure 15A,B). Rates of mussel and barnacle recruitment changed sharply at 32°S, determining a geographic break in adult abundance of these competitively dominant species. Analyses of satellite images also corroborate the existence of regional-scale discontinuities in dynamics and concentration of offshore surface chl-a that had been previously described at coarser resolution (Thomas 1999, Thomas et al. 2001b). Intertidal field experiments showed that the paradigm of top-down control of intertidal benthic communities (Castilla & Durán 1985, Paine et al. 1985, Castilla 1999, Navarrete & Castilla 2003) holds only south of this geographic discontinuity. To the north, populations seem recruitment limited, and predators have negligible effects, despite attaining similarly high abundances. Thus, geographically discontinuous oceanographic regimes set bounds to the strength of species interactions and define distinct regions for the design and implementation of sustainable management and conservation policies along the HCS. Further ecological studies using molecular markers are needed to define the consequences of this variation for the genetic population structure of mussels and barnacles, as well as for other components of intertidal communities, many of which do not experience such a discontinuity in recruitment, despite having similar life histories and general biology (Figure 15C,D).

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Figure 15 Average recruitment of intertidal invertebrates along the coast of central Chile at sites ordered from north to south, from ~29°S to 35°S. Data correspond to long-term (3–7 yr) averages per site of individuals found in replicated collectors that replaced monthly. The arrow in panels (A) and (B) indicates approximate position of regional discontinuity in intertidal chthamalid barnacles in the high intertidal zone and the dominant mussel Perumytilus purpuratus in the mid-zone. See Navarrete et al. (2005) for further details.

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Population connectivity Connectivity can be defined as the extent to which populations in different parts of a species’ range are linked by exchange of larvae, recruits, juveniles or adults (Palumbi 2003) and determines the degree of cohesion of its genetic pool and the geographic structure of its genetic diversity. The intensity and geographical scale of connectivity within a species is given by the realised dispersal through active and passive mechanisms, which depend on species life-history traits and environmental characteristics. Among the numerous dispersal mechanisms reported in marine organisms, active swimming/crawling and planktonic larval transport, together with rafting and anthropogenic dispersal, are considered as the most relevant to achieve connectivity among local populations (Thiel & Haye 2006). Of particular relevance for connectivity of marine species are the temporal and spatial oceanographic characteristics such as currents, upwelling, water masses and gyres. For example, even though a species may have long-lived planktonic larvae, in a particular bay the larvae may not effectively disperse due to local larval retention (e.g., Swearer et al. 1999, Poulin et al. 2002a, Baums et al. 2005). In contrast, benthic species lacking dispersive larval stages can achieve long-distance dispersal by rafting or anthropogenic transport (Thiel & Haye 2006). Indeed, a recent study demonstrated that biogeographic patterns along the coast of South Africa are reflected in the genetic population structure of littoral organisms regardless of their dispersal stages (i.e., with or without planktonic larvae) (Teske et al. 2006). It seems particularly interesting to pursue this avenue in the HCS of northern-central Chile where no distinct biogeographic barriers but rather taxondependent breaks exist (see section on biogeography). Moreover, the unique characteristics of the HCS make it an interesting system to study the genetic connectivity of marine populations. Important characteristics of the system for genetic connectivity are its wide geographic extent and the oceanographic cyclic variations that lead to temporal and spatial changes in population size and distribution. It is expected that both life-history traits and oceanography play crucial roles in determining the realised dispersal of marine populations and thus their connectivity and the extent of their geographic ranges. Few population genetic studies have been published on marine species of the HCS, although there are several currently being developed on pelagic fishes, marine invertebrates and algae. Nevertheless, some predictions may be formulated and, where possible, validated through existing examples. The pattern of genetic connectivity among local populations of a species determines the geographic structure of its genetic diversity (Figure 16). The frequency, intensity and geographical scale of dispersal within a species shape the resulting gene flow that counteracts the action of genetic drift and local selection. In this context, the intensity of genetic drift is principally determined by population dynamic processes such as population size variation, local extinction, recolonisation and founder effects, all intimately related to connectivity among populations. Therefore, acting both on gene flow and genetic drift, the different patterns of connectivity should result in different geographic structuring of the genetic diversity. Very high gene flow (at the scale of the geographic distribution of the species) leads to genetic homogeneity among local populations independently of the geographic distance. With lower levels of gene flow, different patterns of population genetic structure may result depending on the association between gene flow and geographic distance. If the magnitude of the gene flow is associated with geographic distance, which may be the case for many organisms with planktonic larvae, a genetic cline may form through the range of distribution, characterised by genetic differentiation being proportional to geographic distance, a pattern known as isolation by distance (IBD). This pattern will also be influenced by the direction and strength of the currents. If the magnitude of the gene flow is not strongly associated with geographic distance, which may be the case for organisms that disperse through passive mechanisms, the resulting pattern may be chaotic patchiness. So far two parameters have been

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Figure 16 Summary of the patterns of genetic connectivity that may result from the interaction among intensity of the gene flow, its association with geographic distance and its geographic and temporal continuity.

considered: amount of gene flow and association with geographic distance. A third relevant parameter is the geographic and temporal continuity of the gene flow, from very continuous to a highly discontinuous gene flow that will lead to a break in the geographic structure of the genetic diversity. However, because the amount of time required to reach equilibrium between migration and drift is at least hundreds of times the generation time of a species, the genetic structuring may also reflect historic connectivity. Moreover, such equilibrium cannot be reached under high temporal variability in the pattern of connectivity.

Population connectivity studies in the HCS A first and very general prediction for the HCS is associated with the long and continuous extent of the southeastern Pacific coast, without apparent geographic breaks. In this context, IBD and genetic homogeneity should be the prevalent patterns of geographic structure of the genetic diversity, particularly for organisms that achieve high gene flow through long-lived planktonic larvae or frequent rafting routes. The hairy edible crab Cancer setosus, which is of commercial interest, may represent an example of the above scenario. Gomez-Uchida et al. (2003), using allozymes and AFLPs (amplified fragment length polymorphisms), show genetic homogeneity over 2500 km of the Chilean coast for this species. The authors propose that this pattern may reflect the long-lived larvae (60 days) of C. setosus, the absence of geographic barriers and the oceanographic conditions (north and southward currents) of the area that allow effective mixing of larvae. A similar pattern of genetic homogeneity has been observed in pelagic fishes such as Chilean hake (Merluccius gayi gayi) between 29°S and 41°S (Galleguillos et al. 2000) and Chilean jack mackerel (Trachurus murphyi) between 20°S and 40°S (E. Poulin unpublished data). It is predicted that species with high connectivity and extensive geographic ranges may appear less affected by the oceanographic cyclic variations of the HCS, either because they suffer less population reduction or because they have a relatively rapid recovery after a disruptive event. Overall, these taxa may lose less genetic variability and show a faster ecological recovery after ENSO events.

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Conversely, it is predicted that taxa with low dispersal potential will exhibit pronounced genetic structure and will be the most affected by the oceanographic variations. Genetic studies of the macroalga Lessonia nigrescens show that gene flow is limited among nearby populations (Martínez et al. 2003, Faugeron et al. 2005). Additionally, these authors found that 20 yr after the EN 1982–1983 event, which caused a massive mortality of L. nigrescens on 600 km of the coastline, northward recolonisation had only advanced 60 km (Martínez et al. 2003). Lessonia nigrescens is a good example of a species very vulnerable to oceanographic changes, specifically EN, and that may be continuously recovering from drastic population reductions and local extinction, never reaching migration–drift equilibrium. The genetic structure found for L. nigrescens corresponds to chaotic patchiness at a small geographical scale (tens of metres), reflecting recent population dynamic processes (years to tens of years) and life-history traits such as very low distance dispersal of propagules (Faugeron et al. 2005). Other species that show genetic differentiation at a small spatial scale are the alga Mazzaella laminarioides (Faugeron et al. 2001) and the edible and overexploited snail Chorus giganteus that has a low larval dispersal potential (Gajardo et al. 2002). For the edible and also overexploited scallop Argopecten purpuratus, Moragat et al. (2001) found both genetic and morphological differentiation between populations at the two protected sides of the Mejillones Peninsula (50 km apart) and discuss that it is probably due to currents that restrict the gene flow between the two localities. It can further be predicted that biogeographic breaks will reflect strong barriers to dispersal and thus gene flow for species with low dispersal potential, leading to breaks in the geographic structure of the genetic diversity of species. It has been shown that along the northern Chilean coasts, habitat discontinuities can cause gene flow interruptions (e.g., Faugeron et al. 2001, 2005). Species with lower dispersal potential will be more vulnerable to breaks, while species with high potential of dispersal may not show evidence of a genetic break associated with a biogeographic break, as is the case of Cancer setosus (Gomez-Uchida et al. 2003). Even though for the HCS it has not yet been demonstrated that recognised biogeographic breaks correspond with the geographic distribution of the genetic diversity, it has been shown to be the case for other biogeographic regions such as Point Conception in the California Current System (e.g., Burton 1998, Wares et al. 2001). Rafting may be a very advantageous dispersal mechanism for populations that suffer recurrent extinctions and recolonisations, mostly in the extent of the HCS where macroalgae with high floatability are very abundant. Once organisms are in a raft that has the potential to be in voyage for weeks or months, the rafting-mediated gene flow resulting may not be strictly associated with geographic distance and the resulting pattern of connectivity will depend on the intensity of gene flow, that is, if the rafting route is frequent, intermittent or episodic (see Thiel & Haye 2006). We predict that given the abundance of floating macroalgae, rafting routes along the Chilean coast may be intermittent or frequent, leading to patterns of genetic diversity ranging from chaotic patchiness to homogeneity. Ongoing studies of the isopod Limnoria chilensis may contribute to the validation of this prediction. These organisms have the potential to persist in rafts for long periods of time because they are brooders, have local recruitment and feed on the raft. It is interesting to mention that even though Lessonia nigrescens shows high genetic differentiation even at small spatial scales, the geographic distribution of the genetic diversity does not follow an IBD pattern, suggesting that some long-distance dispersal may occur, although it is not known whether it could be via freeliving spores or on drifting fragments of mature thalli (Faugeron et al. 2005). The HCS appears to be an interesting model for studying marine connectivity patterns in variable environments. Despite the general lack of such studies, recent and still unpublished results on pelagic fishes such as anchovies and sardines, and commercially exploited benthic marine invertebrates like the gastropod Concholepas concholepas and the bivalve Mesodesma donacium, support the existence of genetic homogeneity at large geographical scales as a consequence of the absence of contemporary biogeographical barriers along the HCS for species with high dispersal 254



potential. In general, it is expected that further studies of different biological systems will show that all the patterns of connectivity (Figure 16) are present along the HCS as a result of the interaction of present and past environmental conditions with species life-history traits.

Biogeography Large-scale patterns in the HCS The pioneer work by S.P. Woodward in 1856, which is probably the earliest biogeographical classification involving the southeast Pacific (Camus 2001, Harzhauser et al. 2002), was followed by a series of foundational studies (e.g., Dall 1909, Ekman 1953, Stuardo 1964, Viviani 1979, Santelices 1980, Brattström & Johanssen 1983, among others) that provided a consistent view of the major biogeographic features of the HCS temperate area (south of the tropical Panamian Province), based on physical gradients and patterns of endemism, richness and spatial turnover of species, and supported by subsequent studies (see reviews by Fernández et al. 2000 and Camus 2001). Overall, two main biotic replacements along the coast differentiate three biogeographical units (see Brattström & Johanssen 1983 and Camus 2001 for reviews on available classifications): (1) a warm-temperate biota extending from northern Peru (4–6°S) toward a variable, taxondependent limit in northern Chile (usually 30–36°S), often designated as Peruvian Province, and dominated by subtropical and temperate species; (2) a cold-temperate biota (also present in southern Argentina) extending along the fragmented coast of the Chilean archipelago from 54°S to about 41–43°S, corresponding to the Magellanic Province dominated by subantarctic and temperate species, exhibiting reduced wave exposure and an estuarine condition due to the dilution caused by high rainfall levels, glaciers and rivers (Ahumada et al. 2000); and (3) a transition zone between both provinces, characterised by strong numerical reduction of subtropical and subantarctic species at its southern and northern borders, respectively, rather than by diffusive overlap of biotas. However, many species occurring throughout this transition zone have a subantarctic affinity and a wide distribution in Chile (e.g., Menzies 1962, Castillo 1968, Alveal et al. 1973, Santelices 1980), probably facilitated by the HCS transporting cool water masses toward the north, which is also considered to be the main reason why the area lacks a definite biogeographic character. Traditionally, the important physical changes around 42°S are considered to be external forcings that act as effective filters for dispersal, and with few exceptions, this zone represents the steepest induced transition along the HCS coast. Contrastingly, the northern limit of the transition zone is remarkably diffuse for the whole coastal biota (Figure 17) and highly variable depending on the taxon examined (Camus 2001), which has been attributed so far to the apparent absence of major physical discontinuities between northern Peru and Chiloé Island (e.g., Brattström & Johanssen 1983, Jaramillo 1987). Such variation mirrors a typical pattern of transitions (Brown & Lomolino 1998), due to differential attenuation rates among taxa related with their different dispersal ability and physiological tolerance. In fact, some particular taxa (e.g., peracarid crustaceans; Thiel 2002) show a well-defined overlap of northern and southern species with a gradual replacement pattern. On a wider taxonomic basis, however, the breaking points for different taxa do exhibit some latitudinal scattering throughout northern Chile, but they are significantly concentrated around 30°S and 33°S (see comparative analyses of animal and macroalgal taxa in Brattström & Johanssen 1983, Lancellotti & Vásquez 2000, Meneses & Santelices 2000, Santelices & Meneses 2000, Camus 2001). Notably, these multiphyletic breaks include southern and northern limits of species with very different life forms and ecological requirements, even involving pelagic groups (e.g., Antezana 1981, Hinojosa et al. 2006). This information strongly suggests that such breaks are not a passive outcome of dispersal and tracking of key environmental variables. For instance, recent work shows that latitudinal patterns of SST (the main causal factor invoked in most studies) fail to explain 255


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Figure 17 Break points of distribution along the north-central Chilean coast documented in 29 biogeographical classifications published between 1951 and 2006 (breaks outside the range 17–38°S are not shown). The figure includes a wide spectrum of animal and macroalgal taxa of varying hierarchical levels, life forms and habitats, and some classifications may partially overlap in one or more of these categories. Each classification is represented along an imaginary vertical axis in correspondence with the numbers at the bottom, where dashed lines denote the zones in which break points were found (each one marked by a thick horizontal dash), ranging from one to three break points per study. Horizontal lines highlight the latitudes 30°S and 33°S where break points tend to concentrate. Classifications (for numbers 1–10 and 11–22, see references in compilations by Brattström & Johanssen 1983 and Camus 2001, respectively): 1, Mollusca (Carcelles & Williamson 1951); 2, Foraminifera (Boltovskoy 1964); 3, Mollusca (Marincovich 1973); 4, several animal taxa (Knox 1960); 5, eulittoral organisms (Hartmann-Schröder & Hartmann 1962); 6, intertidal Mollusca (Dell 1971); 7, benthic animals (Semenov 1977); 8, several animal taxa (Viviani 1979); 9, Anthozoa (Sebens & Paine 1979); 10, several animal taxa (Brattström & Johanssen 1983); 11, benthic macroalgae (Santelices 1980); 12, Bryozoa (Moyano 1991); 13, fishes (Mann 1954); 14, Isopoda (Menzies 1962); 15, several animal taxa (Ekman 1953); 16, several animal taxa (Balech 1954); 17, planktonic Euphausiids (Antezana 1981); 18, Asteroidea (Madsen 1956); 19, macroalgae (Alveal et al. 1973); 20, sandy beach Isopoda (Jaramillo 1982); 21, Porifera Demospongiae (Desqueyroux & Moyano 1987); 22, several invertebrate taxa (Lancellotti & Vásquez 1999); 23, demersal fishes (Sielfeld & Vargas 1996); 24, littoral Teleostei (Ojeda et al. 2000); 25, Phaeophyta (Meneses & Santelices 2000); 26, Rhodophyta (Meneses & Santelices 2000); 27, several animal and algal taxa (Camus 2001); 28, Peracarid crustaceans (Thiel 2002); 29, benthic Polychaeta (Hernández et al. 2005); 30, pelagic barnacles (Hinojosa et al. 2006).

variations of mollusc diversity along the HCS, which would be determined by shelf area (Valdovinos et al. 2003), while the distribution of some littoral species appears more related to regional variations in temperature trends (Rivadeneira & Fernández 2005). Thus, the transitions in northern Chile are

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not readily explained simply by the contact between warm and cold biotas, and proper explanations will require a multivariate, integrative approach and an exploration of possible external forcings.

The role of past and present processes in northern-central Chile (18–35°S) Different lines of historical and ecological evidence suggest that northern-central Chile constitutes a very complex and dynamical biogeographic scenario. Clearly, present-day patterns are not divorced from physical changes related to the origin and installation of the cold HCS during the Tertiary or from subsequent Quaternary fluctuations (e.g., see Villagrán 1995, Hinojosa & Villagrán 1997, Villa-Martínez & Villagrán 1997, Maldonado & Villagrán 2002). The establishment of the HCS had involved both the northward advance of the subantarctic biota and the northward retraction of a former tropical/subtropical biota (Brattström & Johanssen 1983, Camus 2001), with consequences that may persist until the present, reflected in the heterogeneous character of the northern biota. For instance, 10 of the nowadays most common bivalves in northern Chile exhibit upper thermal tolerances exceeding the highest temperatures recorded during EN events in the past century (Urban 1994), which is unexpected for species evolving in a cold upwelling system. At least one of them (Argopecten purpuratus) is thought to be a relict of the Miocene tropical/subtropical fauna (Wolff 1987). Such physiological ‘anomalies’ suggest the presence of an inertial faunistic component within the modern northern biota (i.e., remnants of the former warm fauna that escaped the forced retraction to lower latitudes, and maintained their warm-water characteristics facilitated by recurrent post-Miocene warming events such as EN) (Wolff 1987). In comparison, the marine flora appears more homogeneous and dominated mainly by subantarctic species, while tropical/subtropical species are virtually absent (Santelices & Meneses 2000). For instance, some common and ecologically important kelp species are not only more sensitive to warming episodes but also their upper thermal tolerance varies in accordance with the thermal latitudinal gradient (Martínez 1999). In this regard, the northern fauna underwent repeated distributional alterations in the past associated with climatic fluctuations. Some of them involved the simultaneous range retraction and expansion of different species (e.g., Ortlieb et al. 1994), but more often the occurrence of tropical/ subtropical fauna related to warming (EN-like) events in the Pleistocene (e.g., Ortlieb 1995) and Holocene (e.g., Llagostera 1979). Moreover, Neogene processes related to the establishment of the modern upwelling system in the HCS (e.g., shallowing of the OMZ) provoked a mass extinction of bivalve molluscs (>75% of species), with lasting impacts on their current distribution patterns and biological characteristics (Rivadeneira 2005), and similar effects have been suggested for the polychaete fauna (R.A. Moreno et al. 2006a). Notwithstanding, the causal relationships between historical events and current distribution patterns in northern Chilean waters remain largely unexplored, although their importance may be overwhelming. On the other hand, modern processes also have strong influences in northern Chile, particularly interannual fluctuations related to ENSO, which, however, should be looked at retrospectively, acknowledging the frequency and importance of EN-like events throughout the Holocene (e.g, see Maldonado & Villagrán 2002). Strong EN events can modify the taxonomic composition of littoral communities (e.g., Arntz 1986, Castilla & Camus 1992, Camus et al. 1994, Vega et al. 2005, Vásquez et al. 2006) and the geographical occurrence of many species including key structural components such as intertidal and subtidal kelps (e.g., Lessonia nigrescens, L. trabeculata and Macrocystis pyrifera). Short-term modifications of community composition during EN occur through local extinctions and invasions, depending also on local conditions, which either favour or prevent their occurrence (Arntz 1986, Camus et al. 1994, Martínez et al. 2003, Castilla et al. 2005a). Moreover, the impacts on key species may scale up to produce long-term biogeographic changes, as exemplified by the dramatic effects of the 1982–1983 EN event on the intertidal kelp Lessonia

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nigrescens (for similar cases in subtidal kelp beds see Vega et al. 2005, Vásquez et al. 2006 and the discussion of EN effects here). Lessonia nigrescens plays a key role in Chilean rocky shores (e.g., Ojeda & Santelices 1984, Castilla 1988, Santelices 1990b), and its presence/absence has direct effects on community organisation and diversity. The kelp suffered a regionally correlated local extinction (also involving the loss of its rich holdfast fauna) along 600 km of coastline, which left a few and highly isolated patches, provoking a strong alteration of its geographical population structure (Camus 1994b, Martínez et al. 2003). The regional recovery process of L. nigrescens was slow, more effective toward higher latitudes, and only partial as it failed to re-establish populations in northernmost Chile (Castilla & Camus 1992). Twenty years later, northward recolonisation advanced less than 60 km, and some recovered populations lost >50% of their genetic diversity exhibiting significant IBD (Martínez et al. 2003). These extinctions also lead to transient changes in biotic interactions within the community (see El Niño section), which had negative effects on local kelp recruitment and contributed to its slow recovery (Camus 1994a). Additionally, Scurria scurra, a limpet living on the stipes and holdfasts of Lessonia nigrescens (Muñoz & Santelices 1989), suffered a concomitant extinction. In southernmost localities, Scurria scurra recolonised and re-established its association with the kelp within 1 yr following the EN event, but in some northernmost localities it failed to do so for at least 11 yr (Camus 1994b) (and remains absent in some places until now; P.A. Camus unpublished data). Overall, the ecological, biogeographical and evolutionary consequences derived from the recurrent extinction-recolonisation dynamics undergone by different species in northern Chile are not yet fully understood. However, it may be argued that they promote changes in spatial patterns of genetic diversity and gene flow, increase betweencommunity diversity, and affect the dynamics of endpoints of distribution, leading to unstable biogeographical limits (Camus 2001, Thiel 2002). This last effect can be reinforced by the transient or permanent invasion of warm-water species favoured by EN episodes (e.g., Soto 1985, Tomicic 1985, Arntz 1986, Castilla et al. 2005a, Coloma et al. 2005, Arntz et al. 2006), thus contributing to the mixed biogeographic character of the northern Chilean biota. However, while some general conclusions can be drawn at this level, a proper understanding of large-scale patterns will need to distinguish their historical and ecological components and consider the physical-biological coupling generating differential responses among taxa. In this regard, the factors affecting dispersal and recruitment deserve special attention. Even though EN is known to be related in varied ways to the recruitment of coastal species (e.g., Soto 1985, Glynn 1988, Vega et al. 2005), in northern Chile its effects on dominant littoral species may be negligible or highly specific, with no clear association with interannual variations (Navarrete et al. 2002). Both mesoscale and regional factors related to the spatial structure of upwelling dynamics seem promising to explain such recruitment variations (e.g., Lagos et al. 2005, Navarrete et al. 2005). Additionally, the spatiotemporal dynamics of the OMZ (e.g., Morales et al. 1999, Palma et al. 2005) and the mesoscale eddy activity bounding coastal ecosystems (Hormazábal et al. 2004) may both play a significant role in understanding the dynamic connection between oceanographic processes and biogeographic patterns.

El Niño-La Niña in coastal marine communities The El Niño Southern Oscillation (ENSO) is the largest modern source of interannual variability in the ocean–atmosphere system (e.g., Wang et al. 1999) and, even though its effects are stronger in the tropics, it significantly affects marine life in northern Chile. The ENSO cycle has been a crucial factor in the global climate for at least the past 130,000 yr (Cane 2005), showing continuous, although variable, activity during the last 12,000 yr (Moy et al. 2002); regionally, it has had a major influence on the Chilean coast since the Holocene (e.g., see Ortlieb et al. 2000, Maldonado & Villagrán 2002; see also Biogeography, p. 255ff.). This suggests that coastal communities in northern Chile have continuously been shaped by impacts of EN events (Camus 1990, 2001). 258



Properties of coastal waters in northern Chile are primarily driven by remote equatorial forcing, which can provoke strong changes in PP due to the availability of nutrients and essential trace elements, corresponding to ENSO cycles (Takesue et al. 2004). Such alterations can trigger a complex chain of biological effects derived from bottom-up controls and physiological constraints, which may involve several levels of biological organisation at different spatial scales, during and between EN. The dramatic and widespread impacts of EN 1982–1983 on coastal communities of northern Chile allowed the identification of biological changes associated with ENSO such as bathymetric or latitudinal migrations, invasion by exotic species, behavioural alterations, reproductive and recruitment failures, increasing population abundance, population decrease due to mass mortality, and in the most severe cases local population extinctions (e.g., see Soto 1985, Tomicic 1985, Arntz 1986, Glynn 1988, Camus 1990, Castilla & Camus 1992, Sielfeld et al. 2002, Vega et al. 2005, Arntz et al. 2006, Vásquez et al. 2006). As a whole, these impacts affect all kinds of taxa and environments, although with clear species- or site-specific components (e.g., in northern Chile none of these effects involves either an entire taxonomic group or the whole suite of species from a given place; e.g., see Soto 1985, Tomicic 1985). Moreover, the type and magnitude of impacts, as well as the range of affected taxa, may vary from one event to another, depending both on the strength of the event and the type of physical-biological couplings that may take place (e.g., see Navarrete et al. 2005). On the other hand, some biotic modifications may occur, or have simultaneous effects, at both local and regional scales (Camus 1994a, 2001, Vega et al. 2005, Vásquez et al. 2006), as observed also in the northeast Pacific (Edwards 2004). Additionally, from a socioeconomic perspective, the increase or decrease in abundance or diversity of fisheries resources at some places may be certainly interpreted as positive or negative effects, respectively (e.g., Arntz 1986). However, from an ecological point of view, it would be as yet uncertain to qualify such changes in the same terms, even for species with no recognisable importance with variations that may have unknown or unpredictable consequences for the community. Thus, simple generalisations on the ecological impacts of EN in northern Chile may still be inappropriate, except at very specific levels. This situation is mainly because of (1) the lack of long-term and systematic biological observations encompassing several events, preventing robust comparisons before, during and after EN conditions, and (2) the irregularity of ENSO itself (e.g., Wang et al. 1999) and the lack of correlation between EN and LN in their strength and duration (e.g., Kerr 1999). Nonetheless, ENSO impacts are indisputably relevant in the ecology of coastal communities in northern Chile. One of the key aspects needed to understand EN effects is its recurrent impact on ‘engineer species’ (sensu Jones et al. 1994) such as kelps, which play a crucial role for the diversity, complexity, structure and functioning of coastal communities along the southeast Pacific (Graham 2004, Vega et al. 2005, Vásquez et al. 2006). Local extinction of kelps is frequent during strong EN events (Camus 1994a,b) such as the 1982–1983 episode, when intertidal populations of Lessonia nigrescens and Macrocystis integrifolia disappeared from the area between 10°S and 21°S and so did the invertebrate community associated with their holdfasts (Soto 1985; see also Biogeography, p. 225 ff.). Concurrent and dramatic impacts were reported during the same event (Soto 1985), affecting ecologically important species of ascidians (e.g., Pyura chilensis), seastars (e.g., Stichaster striatus, Heliaster helianthus) and several fish species, most of them associated with kelp beds. However, the implications of these impacts, both at population and community levels, remain largely unknown. A long-term series of subtidal community dynamics during variable ENSO conditions (1996–2005) has been recently published (Vásquez et al. 2006). Although the EN 1997–1998 was catastrophic and produced local kelp extinctions on the coasts of Chile and Peru (Fernández et al. 1999, Godoy 2000, Martínez et al. 2003), site-dependent conditions allowed the persistence of 259


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some kelp assemblages of Macrocystis integrifolia and Lessonia trabeculata around 24°S (Martínez et al. 2003, Vega et al. 2005). These effects would be related to the frequency and intensity of local coastal upwelling (González et al. 1998, Lagos et al. 2002), which minimised the impact of warming and retained high concentrations of nutrients within the coastal environment (Takesue et al. 2004). A long-term analysis of the structure and organisation of kelp communities in northern Chile (Vásquez & Vega 2004b, Vásquez et al. 2006), which included EN and LN events, showed that the abundance of Macrocystis integrifolia (1) increased significantly during EN 1997–1998, (2) decreased during LN 1999–2001, dropping nearly to zero in 2000, and (3) became reestablished and recovered during a period of positive thermal anomalies in 2002–2003 (Figure 18). This pattern

Figure 18 (A) Upwelling index (Offshore Ekman Transport OET), (B) ENSO index (Southern Oscillation Index SOI-grey line and Multivariate ENSO Index MEI-black line), (C) temporal variability of Macrocystis integrifolia, (D) Lessonia trabeculata, (E) benthic grazers, and (F) benthic predator densities, including El Niño 1997–1998 and La Niña 1999–2000.

260



was different from that recorded on the California coast, where the rapid recovery of M. pyrifera following EN 1997–1998 was favoured by the establishment of a cold period (1998–2000) and the survival of sporophytes in deep environments (Ladah et al. 1999; Edwards 2004). In northern Chile, the recolonisation rate of kelp assemblages occurred comparatively slowly (Martínez et al. 2003; see also Population connectivity, p. 252ff. and Biogeography, p. 255ff.), even though cold conditions prevailing during 1998–2000 enhanced the upwelling effect. In this regard, the slow recovery of Lessonia nigrescens after EN 1982–1983 (Castilla & Camus 1992) appeared more related to biotic constraints: recruitment was strongly reduced by a combination of postsettlement grazing and inhibition by encrusting coralline algae, while erect coralline algae played a key role as facilitators, allowing the kelp some escape from grazers and space competitors (Camus 1994a). On the other hand, the decreased abundance of Macrocystis integrifolia was caused by a significant reduction in the adult plant population and the lack of recruitment of juvenile sporophytes (Figure 18). Thus, the disappearance of the M. integrifolia population occurred 2 yr after EN 1997–1998 and was inversely correlated with a population increase of the sea urchin Tetrapygus niger (Figure 18). In contrast, information from other areas of the southeastern Pacific during EN 1997–1998 showed that superficial warming decreased the abundance of kelp on shallow bottoms, inducing migrations of grazers to deeper zones (Fernández et al. 1999, Godoy 2000, Lleellish et al. 2001). In northern Chile, during EN 1997–1998 and LN 1998–2000, different events favoured the increase of sea urchin populations during the cold phase, including (1) induction of mass spawning due to increases in SST and persistence of upwelling events, (2) reduction in density of adult seastars, and (3) changes in the feeding behaviour of the seastar Heliaster helianthus, one of the most important benthic predators on Chilean and Peruvian coasts (Tokeshi & Romero 1995b, Vásquez et al. 2006) (Figure 18). Thus, the long-term study of subtidal communities suggests that different bottom-up and top-down factors might control ecosystem changes in northern Chile, including (1) the intensity and frequency of upwelling, which may buffer the positive thermal anomalies of SST and maintain high nutrient levels, favouring kelp persistence during EN events; (2) site-dependent oceanographic conditions, which may generate optimal conditions for spawning, larval development, and recruitment of echinoderms during and/or after EN events; (3) an overall abundance increase of carnivores which is correlated with an abundance decline of the most conspicuous grazers; (4) population dynamics of adult seastars and sea urchins which may become decoupled during EN events; (5) species-specific population dynamics of some predator species (e.g., Luidia magellanica), and changes in dietary composition in others (e.g., H. helianthus), which may promote population increase of its prey, the urchin Tetrapygus niger, during EN events; and (6) changes in abundance of T. niger, which might be a key factor controlling the development of two alternate states: environments dominated by kelp beds versus barren ground areas. In a wider context involving both subtidal and intertidal environments, EN impacts can be summarised as a large-scale bottom-up effect influencing various (and as yet difficult-to-predict) levels of marine food webs. However, this is just the initial path for most impacts, and top-down effects should not be neglected (e.g., see Nielsen & Navarrete 2004). Future research on EN impacts could consider at least five aspects related to the variability of biological effects, which may serve as guidelines or study framework: (1) the southward intensity attenuation of EN signals produces a latitudinal impact gradient, with reduced effects toward higher latitudes (e.g., Castilla & Camus 1992, Martínez et al. 2003); (2) in the spatiotemporal context, many effects are episodic and/or local (e.g., abundance variability), and some others may propagate their effects to larger spatial scales (e.g., distribution changes, local extinctions), being highly persistent over time (e.g., see Camus et al. 1994); (3) on a taxonomic basis, some taxa are recurrently affected (e.g., kelps), others exhibit no significant impacts (e.g., chlorophytes), and some taxa can be more affected in their reproduction while others in their recruitment (e.g., Camus 1994a, Navarrete et al. 2005, Vásquez et al. 2006); (4) the genetic and evolutionary consequences of recurrent phenomena such as mass 261


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mortalities, extinction-recolonisation processes, and variations in population connectivity are presumably of critical significance, but they are just beginning to be explored (e.g., see Martínez et al. 2003; see also Population connectivity on p. 252ff. and Biogeography, p. 255ff.); and (5) interannual variations related to EN and LN may be strongly related to both small-scale processes such as the Madden-Julian Oscillation (Madden & Julian 1971) and large-scale processes such as the Pacific Decadal Oscillation (PDO; e.g., Trenberth & Hurrel 1994, Zhang et al. 1997) or the Antarctic Oscillation (e.g., Gong & Wang 1999), with possible biological implications for benthic community dynamics that are virtually unknown. Up to now, the ecological knowledge of EN impacts on the marine communities from northern Chile continues to be mainly descriptive, focused on a reduced number of species and places. Nonetheless, prior studies have shed some light on the wide biological scope and geographical extent of such impacts and the need for comparative, multiscale and long-term approaches to obtain meaningful results.

Physiological adaptations of marine invertebrates Physiological variation is the result of genetic, developmental and/or environmental influences (Spicer & Gaston 1999). Thus, physiological diversity and adaptations are linked to environmental characteristics and variability. The understanding of how living organisms function (i.e., their physiology) is aided by comparing the way different animals deal with environmental constraints (Schmidt-Nielsen 1997). Major environmental factors affecting the animal’s physiology are temperature, oxygen and energy (food) availability. The HCS in northern-central Chile is an interesting scenario for running physiological studies due to changing environmental characteristics, such as (1) the occurrence of a latitudinal temperature gradient, (2) extended zones with permanent and/or seasonal upwelling (cold seawater temperature and low oxygen content), (3) some closed bays with relatively high temperatures (e.g., Antofagasta Bay) compared with the surrounding areas and (4) the occurrence of thermal anomalies like ENSO. The HCS is characterised also by the occurrence of oxygenminimum waters, where physiological adaptations of organisms should be expected, even of species from shallower (10–50 m) waters, which may occasionally be confronted with low oxygen concentrations (when there is upwelling of oxygen-deficient waters). Surprisingly few studies are available on physiological adaptations to hypoxic conditions of benthic organisms from the HCS. One of these studies was the characterisation of the pyruvate oxidoreductase enzymes involved in the biochemical adaptation to low oxygen conditions in nine benthic polychaetes from the HCS (González & Quiñones 2000). Pyruvate oxidoreductase enzymes permit the metabolism to produce adenosine triphosphate (ATP) at high rates under environmental or physiological hypoxic conditions (Livingstone 1983). Interestingly, these enzymes were found to be more numerous and with different pyruvate consumption rates in the most abundant and worldwide distributed polychaete species (Paraprionospio pinnata) (González & Quiñones 2000). Another study of biochemical adaptations to hypoxic conditions in the HCS was done on two key species (in terms of trophodynamics and abundance) of this system, the euphausiid Euphausia mucronata and the copepod Calanus chilensis (González & Quiñones 2002). The enzyme lactate dehydrogenase (LDH, a key enzyme of the anaerobic pathway) from Euphausia mucronata was two orders of magnitude higher than that of Calanus chilensis. Higher activities of the LDH indicate higher anaerobic capacities, and this may enable Euphausia mucronata to conduct daily vertical migration through the oxygen-minimum layer (see also Zooplankton consumers, p. 214ff). In contrast, low LDH activities restrict Calanus chilensis to oxygenated waters (González & Quiñones 2002). The importance of the interaction between oxygen and temperature has been explored in recent studies of the brooding behaviour of decapod crabs (Baeza & Fernández 2002, Fernández et al. 2006b). 262



Closed bays with relatively high temperatures and productivity in the HCS offer suitable conditions for the permanence and reproduction of the scallop Argopecten purpuratus, a species more characteristic of warm waters. An increment of 2.5°C in bottom temperatures (normally 15.5°C) during EN 1982–1983 in Tongoy Bay (30°S) augmented dramatically gonad mass and spawning, and as a consequence spat (juvenile) collection exceeded levels from previous years by 300% (Illanes et al. 1985). However, total gonadal levels of lipids and proteins increased markedly in A. purpuratus conditioned for reproduction at 16°C, but these increases were less pronounced at 20°C (Martínez et al. 2000). Moreover, during gonad maturation muscle carbohydrate levels dropped considerably, as well as the activity of a pyruvate oxidoreductase, the enzyme octopine dehydrogenase (Martínez et al. 2000). Muscle carbohydrate (i.e., glycogen) and glycolytic enzymes have been shown to decrease greatly in other scallop species such as Chlamys islandica and Euvola ziczac (Brokordt et al. 2000a,b). This leads to a decrease in muscle metabolic capacity and thus in escape capacities, which is facilitated by muscle contractions. A reduction of escape capacities during reproduction has been observed in Argopecten purpuratus as well as in Chlamys islandica and Euvola ziczac (Brokordt et al. 2000a,b, 2006). In the intertidal and shallow subtidal zones of the HCS, temperature is the main variable changing over various spatial and temporal scales, with unpredictable interannual patterns. Under normal conditions, physical environmental conditions are relatively stable in the shallow subtidal between 18°S and 35°S (HCS), where salinity typically ranges between 34 and 35 and temperature may vary from 12°C to 22°C. However, due to terrestrial influence, the temperature conditions in the intertidal are different along this latitudinal gradient. For example, the range of mean temperatures registered in high intertidal pools during the summer is ~13–33°C in Antofagasta (23°S), ~13–30°C in Carrizal Bajo (28°S), and ~11–25°C in Las Cruces (33°S) (Pulgar et al. 2006). During EN, these differences in thermal conditions may be enhanced. To evaluate phenotypic plasticity or evolutionary responses of organisms to different habitat temperatures, comparative studies have typically focused on species distributed along latitudinal gradients (Vernberg 1962, Graves & Somero 1982, Stillman & Somero 2000, Pulgar et al. 2006). However, local thermal gradients (TGRs) can be formed by fine-scale variation in, for example, the marine intertidal vertical zones. The intertidal zone is characterised by important spatial and temporal gradients of temperatures, which may be equivalent to those found over a large latitudinal gradient. Intertidal organisms have evolved physiological tolerance adaptations that are important in determining the upper vertical distribution of the species. Studies of congeners or conspecifics allow adaptive variation to be clearly demarcated, independent of effects of phylogeny (Stillman & Somero 2000). Crabs of the genus Petrolisthes (Anomura: Porcellanidae) are widely distributed not only along the intertidal zone of the HCS, but also worldwide, covering huge latitudinal gradients. One of the few studies of physiological adaptations of marine invertebrates in the HSC (Las Cruces, 33°S) was done in five species of the genus Petrolisthes (P. granulosus, P. laevigatus, P. violaceus, P. tuberculatus and P. tuberculosus) (Stillman & Somero 2000). Each species is found at different vertical levels, from the low (P. tuberculosus), mid-low (P. tuberculatus), middle (P. violaceus), mid-high (P. laevigatus) to the high (P. granulosus) intertidal. The limits of thermal tolerance (LT50) were strongly correlated with the vertical position of the species in the intertidal zone (y = 36.02 − 1.88x, r 2 = 0.97) and with the maximal habitat temperature (Table 5) (Stillman & Somero 2000). Thus, species have adapted their upper thermal tolerance limits to coincide with microhabitat conditions. Interestingly, mid-high and high intertidal species (P. laevigatus and P. granulosus, respectively) live near their limits of thermal tolerance. While these LT50 values offer some hints, it may be extremely interesting to explore at which temperatures these organisms enter suboptimal ranges, that is, where they may be able to survive but where growth and reproduction may be compromised. Petrolisthes laevigatus from southern-central Chile dramatically reduces oxygen consumption between 18°C and 20°C (maximal average temperature range found in its 263


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Table 5 Thermal tolerance limits (LT50), and spring maximal habitat temperature of Petrolisthes along intertidal vertical gradient in a locality at the HCS (33°S) Species of Petrolisthes P. P. P. P. P.

tuberculosus tuberculatus violaceus laevigatus granulosus

Vertical position in the intertidal

~ Limit of thermal tolerance (LT50, °C)

~ Maximal habitat temperature (°C)

Low Mid-low Middle Mid-high High

27.5 28.5 30.5 31.5 35.0

14 18 18 28 33

Note: Data from Stillman & Somero (2000).

habitat), which suggests the beginning of the organism’s decompensation, or the commencement of another homeostatic mechanism independent of oxygen consumption (Yaikin et al. 2002). Higher thermal stress or having thermal limits close to actual maximal habitat temperatures might increase the ‘cost of living’ of the species in the upper intertidal (Somero 2002, Stillman 2002). This higher cost of living would be associated with the cost of repairing thermal damage (heat-shock proteins, Hsp) and adapting systems through acclimatisation (Somero 2002). Although tropical Petrolisthes species have higher thermal limits than HCS species, the latter show a wider range of thermal tolerance between low and high intertidal species (Stillman & Somero 2000). Moreover, low intertidal HCS species show greater phenotypic plasticity in their thermal tolerance than high intertidal species (Stillman & Somero 2000). Considering global warming, species inhabiting the upper intertidal zone, living at the ‘edge’ of their thermal limits, would be more affected than species from the lower intertidal, which live far from their thermal limit and with a greater thermal phenotypic plasticity. These physiological traits may have an important effect on the borders in the latitudinal distribution of a species and consequently also on biogeographic limits. Thermal effects that occur outside the normal physiological range involve deleterious changes at the cellular level, especially in systems involved with oxygen uptake, delivery and utilisation (Stillman 2002), such as the cardiac system (Frederich & Pörtner 2000). The heart of a crab species living in the upper intertidal has an Arrhenius break temperature (ABT, temperature at which a break occurs in the slope in an Arrhenius plot, i.e., log rate vs. reciprocal of absolute temperature, K) that is 5°C higher than in a crab species from the low intertidal (Stillman & Somero 1996). These differences were associated with the Na+ K+ ATPase (adenosine triphosphatase) activity, which is necessary for the establishment of the membrane action potential that permits the heartbeat (Stillman 2002). Similar to the rate of heartbeat, oxygen consumption by isolated mitochondria exhibits a ‘break’ at some high temperatures (Dahlhoff & Somero 1993). Phenotypic plasticity in the ABT of mitochondrial respiration has been observed in abalone Haliotis congeners from different thermal habitats (latitude and vertical positions along subtidal-to-intertidal gradient) (Dahlhoff & Somero 1993). Protein synthesis and heat-shock response has been shown to change spatially among gastropod (Tegula) congeners from the temperate subtidal to the low intertidal zones (Tomanek 2001) and seasonally in the bivalve Mytilus trossulus (Hofmann & Somero 1995). Hsp have the function of refolding and ‘rescuing’ proteins damaged by thermal denaturation (Becker & Craig 1994, Hofmann & Somero 1995). Intertidal species of Tegula showed greater expression of Hsp70 than the subtidal species when temperature increases (Tomanek 2001). The energy cost associated with replacing damaged proteins and maintaining Hsp may be an important proportion of cellular energy demands (Hofmann & Somero 1995, Somero 2002). Intertidal and shallow subtidal invertebrates present a ‘cascade’ of physiological responses that enable them to adapt and finally survive changes in environmental conditions. Despite the important 264



latitudinal and vertical environmental conditions gradient of the HCS, there are few studies of physiological responses of invertebrates living in this ecosystem. The range of temperatures registered in the intertidal (rocky pools) along this latitudinal gradient (Pulgar et al. 2006) is very similar to the TGR from the low-to-high intertidal zone observed in central Chile (Table 5). Therefore, it could be expected to find similar thermal physiological variability and local adaptations of congeners or conspecifics in the latitudinal gradient to those found in the intertidal vertical gradient. Because in this area the pattern of environmental variability shifts from a relatively predictable seasonal pattern to a more unpredictable pattern of high interannual variability (i.e., ENSO), physiological characterisation of congeners or conspecific organisms inhabiting the intertidal and shallow subtidal zones along the latitudinal gradient of the HCS would be particularly interesting. Moreover, since algal availability increases from northern to southern Chile (Santelices & Marquet 1998), and physiological compensation associated with environmental stress increases cost of living (Somero 2002, Stillman 2002), latitudinal changes in food availability should also be considered in future studies. It appears particularly interesting to examine how increased costs of living near the distribution limit of a species influences its reproductive potential.

Reproductive patterns of selected marine invertebrates in the HCS Some of the factors that vary with upwelling intensity and persistence, such as temperature and PP, are known to critically affect per capita reproductive investment of marine invertebrates (e.g., MacDonald & Thompson 1985, 1988, Clarke 1987, Brey 1995, Phillips 2002). In the northeastern Pacific, reproductive hot spots coincide with regions exhibiting high PP, which suggests not only that bottom-up processes play a central role in explaining reproductive output, but also that spatial heterogeneity in reproduction needs to be considered in conservation and management plans (Leslie et al. 2005). The clear break in eddy kinetic energy and equatorward wind stress reported at 30°S in the southeastern Pacific (Hormazábal et al. 2004) coincides with two contrasting regimes in chl-a concentration both in coastal areas and offshore (Yuras et al. 2005). Chl-a concentration is negatively correlated with seawater temperature along the HCS (Strub et al. 1991, Thomas et al. 2001b). The effects of small- and large-scale variation in environmental conditions related to upwelling persistence and strength on reproductive patterns along the HCS have recently been analysed.

Primary productivity and gonad production Gonad and egg production of marine invertebrates do not exhibit a clear latitudinal cline in investment in reproduction along the HCS in central Chile (Fernández et al. 2007). However, gonad production shows variable patterns throughout the study region, which extends from 28°S to 36°S, and this variability appears related to the trophic level of the species being investigated and the proximity of the study sites to upwelling centres. Carnivores such as Concholepas concholepas and Acanthina monodon do not show variation in gonad investment between 28°S and 36°S (Figure 19). In contrast, some of the dominant intertidal suspension-feeding species, Perumytilus purpuratus and Nothochthamalus scabrosus, and one of the most abundant herbivores, Chiton granosus, exhibit strong variation in gonad production among sites, even between adjacent locations (Fernández et al. 2007). Investment in reproduction of suspension-feeding and herbivore species is constantly higher in some sites (e.g., Los Molles 32°24′S, Consistorial 33°49′S) and lower in others (e.g., Montemar 32°96′S, Matanzas 33°96′S). Furthermore, there is a positive and significant correlation between investment in gonads among Perumytilus purpuratus, Nothochthamalus scabrosus and Chiton granosus (p always < 0.05; P. purpuratus-N. scabrosus: R = 0.89, P. purpuratus-Ch. granosus: 265


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30°S

28°S 0.15

32°S

34°S

36°S

A

0.1 0.05 0 0.03

B

0.01

D

Pichilemu

El Quisco

Consistorial Matanza

Montemar Quintay

Los Molles

Guanaqueros

Arrayan

Totoralillo

Temblador

Cta. Hornos

E

Huasco

400 300 200 100 0

C

Pta. Choros

0 0.125 0.02 0.075 0.05 0.025 0 0.5 0.4 0.3 0.2 0.1 0 500

Brood size n/basal area

Reproductive output

0.02

Figure 19 Reproductive output of (A) Concholepas concholepas, (B) Acanthina monodon, (C) Chiton granosus, (D) Perumytilus purpuratus, and (E) brood size (number of embryos per unit of basal area) of Nothochthamalus scabrosus, along the study site which ranges from 28°S to 36°S in the eastern South Pacific Ocean. The white bars are used for carnivore species, the black bars for herbivore or filter-feeding species. The white arrows indicate upwelling centers, the black arrows point to sites not influenced by upwelling along the HCS.

R = 0.64, N. scabrosus-Ch. granosus: R = 0.73; Fernández et al. 2007). Investment in gonads of carnivore species does not show any correlation with the reproductive patterns exhibited by herbivores or suspension-feeding species. The patterns reported were consistent over the 3-yr study (coefficient of variation = 1.3%). These results suggest that the small-scale variation in environmental conditions along the upwelling-favourable region of the HCS seems to affect gonad production and consequently larval production, and therefore that some locations potentially serve as source populations of propagules. Transplant experiments support the hypothesis that local environmental conditions determine reproductive output. Suspension feeders (e.g., Perumytilus purpuratus) 266



and herbivores (e.g., Chiton granosus) transplanted between the most and least suitable sites for reproduction in central Chile showed that reproductive output was strongly determined by the site to which the animals were transplanted while the site of origin showed a negligible effect (Fernández et al. 2007). Organisms transplanted to Los Molles showed high reproductive output and organisms transplanted to Matanzas showed low reproductive output, regardless of the site of origin (Fernández et al. 2007). These contrasting results suggest that environmental variables, such as PP, may affect investment in gonads of lower trophic level benthic invertebrates. These environmental conditions seem to be related to the spatial variation of upwelling conditions. The central coast of Chile is dominated by seasonal wind-driven upwelling that forces cold, nutrient-rich water into the upper water column (Wieters et al. 2003). However, it is remarkable that the well-documented relationship between cold upwelled water and high chl-a concentration over large spatial scales off the coasts of Chile and California (Strub et al. 1991, Thomas et al. 2001a) is not observed at smaller spatial scales (Wieters et al. 2003). Between 28°S and 36°S, the lowest chl-a concentrations are associated with coldest upwelled waters (Wieters et al. 2003). Matanzas and Montemar are sites influenced by upwelling centres, in contrast with the lower influence of upwelling in areas such as Los Molles or Consistorial (Broitman et al. 2001, Wieters et al. 2003). Upwelling centres also show higher growth rate of corticated algae, which are not consumed by herbivores, and low growth of ephemeral algae (Nielsen & Navarrete 2004). Although temperature decreased from 28°S to 36°S and is lower in upwelling centres, gonad production is not correlated with seawater temperature (Fernández et al. 2006a). Most likely, the low chl-a concentration associated with upwelling conditions (Wieters et al. 2003) and low abundance of benthic ephemeral algae (Nielsen & Navarrete 2004) are the main factors affecting reproductive output of lower trophic level, intertidal species along the HCS. Evidence from other geographic regions supports the hypothesis that patterns of PP associated with upwelling conditions determine reproductive output (Leslie et al. 2005). However, more information is necessary to clearly identify the set of environmental variables affecting reproductive investment.

Temperature and brooding requirements Although the effect of temperature on gonad production is not evident along the upwelling region associated with the HCS between 28°S and 36°S, studies conducted over larger spatial scales (and wider temperature ranges) suggest that temperature can affect egg production in species that aggregate embryos. Oxygen is a limiting factor in embryo aggregations of marine invertebrates (Cohen & Strathmann 1996, Lee & Strathmann 1998, Fernández et al. 2003) and temperature affects oxygen availability in different types of embryo packing (Brante et al. 2003, Fernández et al. 2006a). Studies of the brachyuran crab Cancer setosus show that brooding females respond to the increased oxygen demand of the embryos at higher temperatures by increasing abdominal flapping frequency, a behaviour that supplies oxygen to the brood (Brante et al. 2003). Similar patterns of female response to embryo oxygen demand have been reported throughout embryo development in other crab species (Baeza & Fernández 2002, M. Fernández et al. 2002). The change in female brooding behaviour (abdominal flapping frequency) produces a higher rate of oxygen supply, dramatically affecting the costs of brooding. Between 10°C and 18°C, a 45% increase in brooding costs was estimated for large-size crabs, such as C. setosus (Brante et al. 2003). This substantial increase in brooding behaviour and cost seems to affect egg production and survival (Fernández et al. 2003). Field data showed that gonad investment (or reproductive output) in C. setosus is lower in northern Chile (approximately 20°S) than in central (30–33°S) and southern (40°S) Chile. Lardies & Castilla (2001) reported a similar latitudinal trend in reproductive output for Pinnaxodes chilensis. Moreover, embryo loss in Cancer setosus increases with temperature (Brante et al. 2003), suggesting that temperature also affects embryo survival throughout the brooding period. Reproductive output of most brachyuran crab species does not vary south of 30°S 267


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(Brante et al. 2004). The contrasting pattern in reproduction between populations of brachyuran crabs north and south of 30°S seems also to occur in other species that aggregate embryos. It is interesting that despite the lack of difference in reproductive output of Concholepas concholepas between 28°S and 36°S (Figure 19), embryo packing shows a clear break that coincides with the break exhibited by brachyuran crabs (north and south of 29–30°S). The mean number of embryos per unit of capsule area was significantly lower in capsules from sites north of 29°S than south thereof (20.6 vs. 31.8, respectively), over a range of study sites located between 30°S and 42°S. Although there is no dramatic break in temperature at 30°S (Broitman et al. 2001), the pattern of embryo packing is explained by the mean temperature at the study site shortly before egg deposition. The evidence accumulated to date suggests that the higher brooding costs north of 30°S seem to affect reproductive output and that the effect is consistent across the brooding modes and taxa studied so far. This finding suggests that environmental conditions are less favourable for brooding species in the northern part of the HCS. In fact, large-scale studies of the patterns of species distribution in relation to larval developmental mode support the hypothesis that the spatial distribution of brooding species is explained by the cost of brooding, which is associated with the cost of oxygen provision (Astorga et al. 2003, Marquet et al. 2004). It is less clear if the impact of the cost of brooding on reproductive output may change with adult body size. Recent studies have shown that small crab species perform the same active brooding behaviours as large species (e.g., Pisoides edwardsi, Acanthocyclus gayi; Fernández et al. 2006b). However, mean oxygen consumption of brooding females, which is a proxy of brooding cost, is not significantly different from oxygen consumption of non-brooding females (Fernández et al. 2006b). These results contrast with previous studies of large species, showing a 2- to 3-fold increase in oxygen consumption by females carrying later-stage embryos over non-brooding females (e.g., Cancer setosus: Baeza & Fernández 2002; Homalaspis plana: Ruiz-Tagle et al. 2002; Ovalipes trimaculatus: Fernández & Brante 2003). Therefore, the patterns described for brooding species may be dependent on adult body size. In fact, the small crab species Acanthocyclus gayi and A. hassleri do not show any pattern in reproductive output along the region extending from 28°S to 36°S (Espinoza 2006). Although existing evidence on reproductive output along the HCS suggests that the behaviour of populations north and south of 30°S depend on the cost of oxygen provision in larger-size species, more studies are needed in order to generalise the findings on the effects of temperature on brooding costs, the link between adult size and brooding mode, and the consequences on species distribution, especially in regions influenced by upwelling where low oxygen conditions cover extended regions of the ocean (Fuenzalida et al. accepted).

Larval life in the HCS Oceanographic conditions in the HCS ecosystem expose planktonic larval forms to environmental conditions in which they do not behave simply as a passive particle. Therefore, morphological, physiological and behavioural characteristics present in fish and benthic invertebrate larval forms inhabiting any particular habitat can be interpreted as effective local adaptations evolved to face this unique ecosystem. Herein, the aim is to give a brief overview of some behavioural and feeding traits together with transport processes described in larval forms of benthic invertebrates and fishes inhabiting the HCS.

Behavioural traits Larvae of many marine benthic invertebrates, differing from lecitotrophic short-lived larvae, spend extended periods in the plankton prior to settlement and metamorphosis (Thorson 1950, Pechenik 1986, 1999). Usually, during their pelagic phase, planktotrophic larvae (PL) of marine benthic 268



invertebrates must remain suspended, locate and gather food, avoid predators and unfavourable conditions, disperse to new areas and select sites for settlement (Strathmann 1974, Palmer & Strathmann 1981, Scheltema 1986). The importance of larval behaviour has been recognised as an important subject in marine ecology and the recent focus on the role of nearshore oceanographic processes controlling larval dispersal and recruitment of benthic organisms has revived interest in the behaviour of larvae in the plankton (reviewed in Le Fèvre & Bourget 1992). Although larval transport seems to be mainly controlled by hydrographic factors (Thorson 1950), larval behaviour may influence their final destination (e.g., Butman 1987, Pineda 1994a,b). In general, sources of mortality in the dispersal phase for PL are food limitation, extreme conditions of temperature and salinity, low dissolved oxygen, UV radiation and pollution (Pechenik 1986). However, because of the difficulties associated with larval tracking in the field, the relative importance of each potential source of larval mortality in nature is still largely unknown. Field and laboratory experiments have revealed a number of behavioural mechanisms that allow larvae to contend with these selective pressures on mortality. However, specific studies of characteristics displayed by PL throughout the HCS are largely absent. Few published studies regarding larval characteristics linked with ecological importance have been conducted in Chile (e.g., Manríquez & Castilla 2007). Field evidence of a DVM pattern in competent larvae has been described for the muricid gastropod Concholepas concholepas (Poulin et al. 2002a,b). It has been suggested that the behaviour of competent larvae of this species may help them to avoid offshore transport during upwelling events. Studies with several crustacean species have shown clear synchronisation between reproduction and the hydrodynamic processes promoting larval transport or retention (Yannicelli et al. 2006a). Similar studies with other PL in the HCS incorporating behavioural or physiological responses to environmental variables such as currents, temperature and salinity are urgently needed. However, along the HCS those studies are precluded by the absence of basic knowledge such as larval identification in many species of Chilean invertebrates. Work to investigate the scale of the source–sink dynamics of PL in the temporally heterogeneous HCS ecosystem is also largely absent. More recent techniques involving chemical tags in calcified structures of PL have been successfully used to identify potential larval source and sink areas, and larval trajectories in the HCS (Zacherl et al. 2003).

Feeding and larval food environment in the HCS Recruitment is recognised as the leading determinant of population dynamics of benthic invertebrate and fish species and it strongly adjusts the importance and intensity of species interactions in the HCS (see above). Since PL rely on food to grow and develop, availability and quality of food particles during larval development are important factors influencing larval recruitment (Vargas et al. 2006a). Despite the high variability in larval food composition, there are well-documented, persistent, temporal and spatial differences in plankton structure and abundance along the HCS. For instance, large and geographically persistent heterogeneity in chl-a has been documented along the Chilean coast (Thomas et al. 2001b, Yuras et al. 2005; and see Dominant primary producers and their role in the pelagic food web, p. 210ff.). The general pattern indicates that a large percentage of chl-a is found in the large phytoplankton fraction in permanent coastal upwelling areas off northern Chile (Iriarte & González 2004; see also Dominant primary producers and their role in the pelagic food web, p. 210ff.), as well as during spring/summer at seasonal upwelling sites in central Chile (González et al. 1989, Vargas et al. 2006b). Similarly, a large and geographically persistent heterogeneity in chl-a levels has been observed to the north of latitude 32°S (Yuras et al. 2005). The scarcity of available published data on larval fish and invertebrate feeding makes it difficult to evaluate potential larval survival and recruitment along the HCS. Currently, the largest body of literature is available for larval fish of commercially exploited species. Because of their seasonal 269


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Diatoms

Heterotrophic nanoflagellates Microprotozoa

Copepod stages

mid-size prey

Physical processes Affecting both food and larvae Turbulence Internal waves, tides River plume motion Frontal structures Filaments Upwelling/downwelling

large prey small prey

Diatoms

Microprotozoa

Autotrophic nanoflagellates

Heterotrophic nanoflagellates

Picoautotrophs

Bacteria

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+

Food size

Spatial/temporal heterogeneity

Autotrophic nanoflagellates

Bacteria

+

Low chlorophyll

Picoautotrophs

Food size

High chlorophyll

Heterotrophs

−

Autotrophs

−

Heterotrophs

Figure 20 Conceptual scheme of main pathways of interaction in coastal food webs involving invertebrate (i.e., veliger competent larvae and barnacle nauplii) and fish larvae under spatial/temporal variation in chlorophyll levels in the Humboldt Current System. The thickness of the arrows represents main predator-prey interactions. The sizes of the boxes or circles represent the dominance in terms of biomass of a specific food item (both autotrophic and heterotrophic prey) during each condition. Physical processes discussed in this chapter, which affect both larvae and food distribution, are also included. Arrows directed to food items at top or bottom boxes were included for convenience.

nature, in the coastal upwelling area off central Chile (36–37°S), the changes in the phytoplankton protozoan and microplankton community that constitute the food supply for larval fish also occur on a seasonal basis (Vargas et al. 2006b). Accordingly, larvae produced in the middle of winter in the HCS, when PP and seawater temperatures are low and wind-induced turbulence in the upper part of the water column is high, are probably not going to face the same prey as those produced in summer, when upwelling and PP are at maximum (Figure 20). Most of the studies of larval fish feeding have focused on changes in prey field, diet overlap, and their influence on larval feeding over short timescales, between adjacent areas, or have attempted to investigate whether evidence of starvation occurred in wild-collected larvae (Llanos et al. 1996, Balbontín et al. 1997, Pizarro et al. 1998, Llanos-Rivera et al. 2004). For the anchovy, Engraulis ringens, in northern Chile (20–21°S), scarce incidence of starvation was even observed during autumn, a season of reduced 270



plankton production (Pizarro et al. 1998). In northern Chile, the diet of the myctophids Diogenichthys laternatus and Triphoturus oculeus, which feed diurnally and share the upper water column (0–200 m depth), was shown to vary according to prey availability in the field. The diets of both species overlap in periods and areas where food is more abundant (e.g., copepods, copepodids, nauplii, invertebrate eggs and ostracods), but differ under conditions of low prey availability (RodríguezGraña et al. 2005). Other studies in central Chile have shown that either microplankton concentrations did not appear limiting in winter (Castro et al. 2000, Hernández & Castro 2000, Castro 2001) or larval diet overlap occurred among several species but during periods of high food abundance (Llanos et al. 1996, Balbontín et al. 1997, Llanos-Rivera et al. 2004). Hence, the few studies carried out on feeding of larval fish along the HCS in Chile suggest that, although the larval food abundance may vary among seasons and localities, starvation due to limiting food availability alone does not seem to be such a common feature, even in seasons of lower production (autumn and winter). For starvation to occur, other factors may be necessary, such as increased turbulence, which may play a concomitant role during the low productivity seasons, at least in southern Chile (Cury & Roy 1989, Castro et al. 2002). Scarce dietary overlap and local morphological and physiological adaptations (i.e., increased reserves in fish eggs and larger yolk size in fish larvae; Llanos-Rivera & Castro 2004, 2006) seem to exist in larval periods when ontogenetically food is highly necessary (i.e., onset of feeding) or when food becomes less abundant in the environment (Balbontín et al. 1997, Llanos-Rivera et al. 2004, Rodríguez-Graña et al. 2005). For benthic invertebrates, sharp spatial transition in phytoplankton biomass associated with upwelling dynamics has been assumed to have important effects on larval condition (Wieters et al. 2003). This assumption is based on the idea that phytoplankton is the most important food item for PL. A major impediment to understanding how food quantity and quality influence larval life under natural conditions is that virtually all published information comes from laboratory rearing studies in which larvae are offered a monospecific or controlled-mix algal culture. One of the first studies done in the HCS analysing feeding preferences in larval invertebrates, by Vargas et al. (2006a), found evidence that barnacle nauplii (Jehlius cirratus and Notobalanus flosculus) and veligers (Concholepas concholepas) exhibited high consumption of heterotrophs (i.e., ciliates and dinoflagellates), but the size spectrum of food particles removed by barnacle nauplii was in contrast with those for C. concholepas veligers. Barnacle nauplii preyed heavily on small picophytoplankton (20 µm). The ability of barnacle nauplii to feed on small prey hinges on the small spaces between their limb setules (Stone 1989). This important finding indicates that omnivorous larval feeding should be the norm in the pelagic ecosystem and might explain why larvae maintain positive growth in environments where phytoplankton is thought to be limiting (Crisp et al. 1985). Therefore, the scarce published information for the HCS suggests that the inference of patterns of larval condition and recruitment over large scales from chl-a biomass, now easily measured from satellite images (e.g., Thomas et al. 2001b), has to be regarded with caution. The scenario suggests that a large spectrum of food particles is available for larval feeding, and species may adapt their feeding preferences in relation to temporal/spatial food distribution along the HCS as well as their physiology and energetic reserves to counteract the spatial and temporal variations in food quality and quantity (Vargas et al. 2006b) (Figure 20).

Upwelling and larval transport processes The wind-driven seasonal upwelling, besides its paramount effect on PP in the coastal zone, induces profound changes in the dynamics of coastal waters that directly affect the distribution and abundance of organisms in the nearshore areas as well as over the continental shelf and slope. To reside 271


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in these hydrodynamically variable but highly productive environments, coastal and offshore organisms have developed a series of reproductive strategies that enable them to counteract the effects of offshore advection in the surface Ekman layer or take advantage of other oceanographic processes to return to the coastal zone (see also Coastal oceanography, p. 205ff.). For instance, on a seasonal timescale, small pelagic clupeiform fish tend to synchronise their reproductive timing with processes that induce shoreward transport and coastal retention. Probably the best-documented species in the HCS of central Chile are the anchovy (Engraulis ringens) and common sardine (Strangomera bentincki) that reproduce during winter, when the wind-driven Ekman layer is directed shoreward and thus eggs and larvae are retained in the coastal area (Castro et al. 2000, Castro 2001, Cubillos et al. 2001, Hernández-Miranda et al. 2003). Shelf-break, slope-demersal and mid-water fish species, such as Chilean hake (Merluccius gayi), big eye flounder (Hippoglossina macrops), and the mesopelagic Maurolicus parvipinnis, instead seem to prefer early spring reproduction when subsurface waters drive their eggs and larvae to the coast during the upwelling season, where they develop in the season of higher production (Vargas et al. 1997, Vargas & Castro 2001, Landaeta & Castro 2002, Landaeta et al. 2006). This strategy, originally described for some large offshore copepods such as Rhincalanus nasutus, and more recently observed also in the galatheid Pleuroncodes monodon and the majid Libidoclaea granaria, is currently accepted as a common feature among several types of organisms that have in common larvae inhabiting subsurface waters in central Chile (Castro et al. 1993, Yannicelli et al. 2006a,b). Other species of decapod crustaceans also synchronise their reproduction with the seasonal changes in hydrodynamics for their transport or retention (Yannicelli et al. 2006b). In this group, however, several species reproduce during the upwelling season in spring and summer and their horizontal distribution seems to be associated with their behavioural ability for vertical migration (i.e., Neotrypaea uncinata, Pagurus sp.). Other species with protracted larval periods such as Emerita analoga and Blepharipoda spinimana, which reproduce late in summer and early spring and then reside in the upper water column without signs of vertical migration, probably use more than a single retention process, as yet unknown, because their reproduction occurs over periods of contrasting hydrographic processes (Yannicelli et al. 2006b). Among molluscs, there is scarce information; for instance, the larval retention and shoreward transport mechanism of the gastropod Concholepas concholepas are the only information reported for the HCS. In this case, avoidance of the surface Ekman layer by competent larvae appears to be accomplished by an inverse vertical migration that reduces their time exposed to seaward flow and keeps the larvae between the coast and an upwelling front (Poulin et al. 2002a,b). Coastal oceanographic processes such as upwelling shadows have also been reported (Escribano et al. 2002) as larvae retention mechanisms and such might constitute an understudied coastal larval retention system along the HCS. Tidal transport of larvae, associated with frontal structures near the coast, the entrance of estuaries or large bays, has also been reported recently for central and southern Chile. Internal tidal bores associated with semi-diurnal temperature changes coincided with bivalve, gastropod and crustacean settlement, suggesting that coastward larval transport occurred during these events in summer (Vargas et al. 2004). In Corral Bay (~40°S), changes in larval fish distributions coincident with changes in the estuarine front position in different tidal phases have been reported by Vargas et al. (2003) as an indication of potential larval tidal transport. More recently, with the aid of fineresolution current profiles and stratified larval collections over 24-h cycles, semi-diurnal changes in water flow patterns and of decapod and larval fish fluxes in and out of the Gulf of Arauco have been estimated (Yannicelli et al. 2006a, R. Veas unpublished data). The overall larval fluxes were modified by their vertical position in the water column, the diurnal vertical migration patterns and the tidal cycle. Interestingly, although clearly associated with tidal phases, the larval fluxes experienced by these decapod and fish larvae in the Gulf of Arauco did not correspond exactly to selective tidal stream transport (STST) as it has been reported at the entrance of other bays and estuaries in other coasts of the world (Forward et al. 1998). 272



Seaward and alongshore exportation of larvae in river plumes, filaments and eddies seems to be a common feature along the HCS. In northern Chile, a mixture of coastal and offshore larval fishes, both in the coastal and adjacent offshore areas, has been repeatedly reported in different seasons and years (Loeb & Rojas 1988, Rojas et al. 2002, Rodríguez-Graña et al. 2005). Besides the existence of a narrow continental shelf, at least three oceanographic processes involving larval transport have been advocated to explain such patterns: (1) seaward surface Ekman transport and upwelling plumes, (2) the presence of filaments, and (3) warm-water intrusion during EN years, all well-documented processes in northern Chile (González et al. 2000b, Sobarzo & Figueroa 2001). In central Chile, surface Ekman transport and cold-water filaments have also been reported to export chl-a, meroplankton and ichthyoplankton from the coastal zone (Cáceres 1992, Morales et al. in review) and they have been used to explain part of the larval mortality rate estimations in the coastal zone (Castro & Hernández 2000, Landaeta & Castro 2006). Larval transport associated with river plumes, mesoscale processes not occurring in northern Chile, has also been proposed as a determinant of barnacle larval transport (Vargas et al. 2006c). Such plumes are also potential areas of increased larval food availability at the frontal area. Other still-unknown oceanographic processes capable of transporting larval forms alongshore probably occur in the HCS. In fact, it is worth noting that almost no reports exist on the role of the Humboldt Current itself or the Chile-Peru poleward undercurrent as a means of alongshore latitudinal larval transport (except for potential transport of larval Pleuroncodes monodon in subsurface waters off central Chile; Yannicelli 2005). In summary, a number of larval transport processes have been described along the central and southern part of HCS. Three of them seem particularly important when considered from an ecological perspective: (1) larval transport (seaward and shoreward) in the surface Ekman layer, (2) exportation from the coast in filaments (especially off the Mejillones Peninsula in northern Chile and off the Talcahuano area of central Chile) and (3) the coastward subsurface larval transport in upwelling waters in spring and summer with the concomitant mixing of offshore and coastal species near shore. Examples of important differences between northern and central Chile are the influence of oceanographic processes such as EN events (much stronger in northern Chile) and of river plumes (absent in northern Chile). Overall, populations of invertebrates and fishes along the HCS develop multiple strategies to cope with the intense periods of transport during early life stages. Timing the reproductive seasons with specific oceanographic events is most common. However, specific reproductive timing depends on the local sites of reproduction, capability of larvae to move vertically in the water column, length of larval life history/cycle and other not so well studied processes (i.e., retention in upwelling shadows) or those occurring during the adult life stage (seasonal growth, energy accumulation, oogenesis) that may finally affect the larval characteristics mentioned above.

Life-history adaptations of macroalgae The macroalgal flora along the HCS is characterised by the presence of endemic species (32%) mixed with species of different biogeographical affinities, including subantarctic (34%), widely distributed (23%), tropical (3.5%) and amphi-equatorial (7%) species (Santelices 1980). In general, this species composition, which is comparatively poorer than that reported for other regions and increases in number toward higher latitudes (Lüning 1990, Santelices & Marquet 1998), responds to the relative biogeographic isolation as a consequence of the predominant direction of the water circulation regime in the HCS (Santelices et al. 1980, Meneses & Santelices 2000). However, an increasing number of invasive species as a result of human activities (ship transport, aquaculture, etc.) has recently been reported with concern about their impact on indigenous species (Castilla et al. 2005a). For example, recent surveys indicate an increase of subtropical elements, a decrease in endemic species and two break points in species composition at 12°S and 42°S (Meneses & Santelices 2000). 273


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Environmental gradients and strategies of resistance, recovery and recolonisation Although benthic algae are not in direct contact with the main stream of the HCS, subsystems such as the Arica-Mejillones Current or Gunther Undercurrent are located close to the coast (13–15°C (Peters & Breeman 1993). Such thermal requirements suggest that they may survive moderate EN episodes, but probably a recovery through sexual reproduction may become limited. In the case of species restricted to the HCS (e.g., the red alga Chondracanthus chamissoi), the growth of both gametophytes and sporophytes increases at temperatures of 25°C, which is 5–6°C higher than water temperature in northern Chile (Bulboa & Macchiavello 2001). This is a known phenomenon in macroalgae and may reflect a reasonable safety limit to survive unpredictable increases of temperature for a long time (e.g., months during EN) or may represent a potential for shifting the distribution boundaries northward. It must be emphasised that most of the macroalgae from northern-central Chile studied so far show a capability of growth at very low temperatures (>10°C), indicating clearly an adaptation to the cooling effect of the HCS and reflecting their subantarctic affinities (Wiencke & tom Dieck 1990, Peters & Breeman 1993, Santelices & Marquet 1998). Physiological and morphofunctional adaptations In general, there are only a few ecophysiological studies addressing adaptations of the life history of macroalgae to varying light and nutrient conditions and they have normally been restricted to the genera of commercial importance (e.g., Lessonia, Gracilaria, Gelidium, Mazzaella and Porphyra) (Oliger & Santelices 1981, Hoffmann & Santelices 1982, Hoffmann et al. 1984, Correa et al. 1985, Hannach & Santelices 1985, Avila et al. 1986, Bulboa & Macchiavello 2001, Véliz et al. 2006). Although physiological performances (measured as growth or photosynthesis) are comparable to those of species from other biogeographical regions, much of the existing information is site-specific and has been gathered from individual species, indicating adaptations to narrow ranges of environmental variability. However, in genera such as Gracilaria and Porphyra, which are exposed to highly changing environmental conditions in enclosed bays, estuaries or upper littoral zones, broader ranges of environmental tolerance may be expected (Gómez et al. 2004, 2005a). Due to its physical configuration, the coast along the HCS is characterised by high energy, and hence physical perturbations such as wave action or sand erosion/accretion fluctuations are important and often govern the population dynamics of various infralittoral algae such as Lessonia, Mazzaella and Gymnogongrus. Thus, algae have developed a suite of morphofunctional adaptations such as alternation of crustose and erect morphs, large size and seasonal regulation of abundance (Santelices et al. 1980, Jara & Moreno 1984, Santelices & Ojeda 1984, Santelices & Norambuena 1987, Gómez & Westermeier 1991, Westermeier et al. 1994, Vega & Meneses 2001). In many cases, these adaptive strategies operate in the early development of the life cycle and in both isomorphic 276



and heteromorphic phase expressions. Processes such as coalescence of spores (Santelices et al. 1999, 2003), regrowth from crusts or vegetative propagation (Hannach & Santelices 1985, Gómez & Westermeier 1991, Macchiavello et al. 2003), selective mortality of early developmental stages (Martínez & Santelices 1998), differential phase ratios (Vega & Meneses 2001) and synchronisation of spore release (Edding et al. 1993, Tala et al. 2004) have been described in algae from northerncentral Chile, which are related to potential selection under changing environmental conditions. While the seasonal and latitudinal gradients in environmental conditions along the HCS are recognisable, the magnitude of their impact on the physiological and reproductive biogeography of benthic algae remains diffuse. Therefore, comparative studies focused on assemblages from different sites along the HCS are needed in order to define, for example, the environmental thresholds involved in reproductive fitness, physiological adaptation and recovery capacity. A special case: enhanced solar radiation Ozone depletion (with the concomitant increase of UV-B radiation) over the Antarctic region, which in spring can reach areas as far north as 36°S, has opened the debate about its consequences on the marine biota of cold and temperate regions (Madronich et al. 1995, Sobolev 2000). Short wavelengths (UV-B) affect photosynthesis in different ways and have detrimental effects on DNA and other key cell components (Bischof et al. 2006). Recent studies indicate a potential impact of current solar UV radiation on photosynthesis of intertidal macroalgae from the southern limit of HCS (39°S; Gómez et al. 2004, Huovinen et al. 2006). In northern Chile (30°S), zoospores, gametophytes and embryonic sporophytes of subtidal Lessonia trabeculata and the intertidal L. nigrescens show elevated sensitivity to UV exposure, leading to high spore mortalities and decreases in germination under current UV doses (Véliz et al. 2006). In general, UV sensibility of early stages correlates with the depth-distribution patterns of the parental sporophytes, suggesting that this factor may play a relevant role in depth zonation of benthic algae in this region (Gómez et al. 2005b). Recent surveys indicate that intertidal species display various photoprotective mechanisms, in particular the ‘dynamic photoinhibition’, regarded as a down-regulation of the photosynthetic apparatus to quench the impact of excess energy (Gómez et al. 2004). Some intertidal algae also have noticeably high contents of UV-absorbing substances (e.g., mycosporine-like amino acids) (Huovinen et al. 2004). Whether algae from different latitudes exhibit a differential susceptibility to UV radiation in the context of the HCS remains unclear and future work should give new insights into the ecophysiological strategies of macroalgal assemblages in scenarios of climate change.

A brief history of exploitation of natural resources in the HCS First coastal settlers Evidence of coastline occupation off Peru and Chile may date from as early as 11,000 yr ago (Llagostera 1979, Muñoz 1982, Báez et al. 1994, Keefer et al. 1998, Sandweiss et al. 1998, Méndez 2002, Méndez & Jackson 2004, Santoro et al. 2005). One of the first studies of the earliest human habitation along the coast of northern-central Chile (Arica to Coquimbo) was carried out by Junius Bird in 1941 during excavations of shell middens. This author (Bird 1943, 1946) showed that the earliest evidence of human settlements on this coast went back to ~6000 yr before present (BP) and people used highly specialised artifacts with an efficient maritime adaptation. According to Santoro et al. (2005) evidence of the earliest inhabitants (~11,000 yr BP) is difficult to find since most prehistoric sites are now on drowned landscapes. However, there is clear evidence that coastal populations in northern-central Chile used marine natural resources during the Holocene (Llagostera 1979). It is thought that early fishermen dived in subtidal waters to collect molluscs and they also 277


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speared large fish and marine mammals such as sea lions (Otaria juvata and O. flavescens) (Santoro et al. 2005). A report on the navigation off the west coast of South America carried out by Lothrop (1932) showed artifacts used by aboriginal people of northern Chile. Most remarkable was a raft composed of two cylinders of seal hides tied together to support a small platform on top, and while voyages out of sight of land were rarely attempted, it was not uncommon to remain at sea for 2 or 3 days (Lothrop 1932). Llagostera (1979), working between 21°S and 25°S, found evidence of groups of people with the ability to exploit natural resources from coastal waters ~10,000 yr BP, and he found remains of molluscan species such as Concholepas concholepas, Fissurella spp., Tegula atra, Choromytilus chorus, several species of fishes (e.g., Isacia conceptionis and Trachurus murphyi), and semifossilised bones of sea lions and dolphins. Reports from southern Peru suggest specific sites where mainly fish and seabird resources were exploited (Keefer et al. 1998), while at other (more ephemeral) sites they processed mainly molluscs (Sandweiss 2003). Studies from southern Peru and northern Chile demonstrated that the resources exploited by early coastal settlers of the HCS changed with time (Llagostera 1979, Sandweiss et al. 2001), which is taken as an indication of climate change and EN events, causing gradual or abrupt changes in available resources. Similar observations were made in central Chile near 32°S (Báez et al. 2004). For the late Holocene (4000–2000 BP) Méndez & Jackson (2004) reported a high degree of mobility of coastal people in central Chile, who apparently moved between different sites in a region, exploiting the accessible resources at a given site in an opportunistic manner. A systematic study of remains of marine invertebrate fauna from central Chile (Curaumilla, 33°S) defined the ecological role of early inhabitants as shellfish gatherers (Jerardino et al. 1992). According to these authors, they probably modified areas of the rocky intertidal, causing decreases in mean sizes of Concholepas and Fissurella. Interestingly, Llagostera (1979) suggested that the appearance of larger shells of Concholepas in shell heaps is indication of an increasing radius of action and the exploitation of new fishing grounds. This author also remarked that the appearance of some fish species (e.g., cusk eels, locally called ‘congrio’, from the genus Genypterus, which can only be fished at greater depths) around 3000 BP is indication that coastal fishermen started to venture farther out to sea during that time period. In general, prehistoric people used littoral resources in an opportunistic manner, and during the past millennia they increasingly widened their radius of action, improved their navigating skills (rafts) and developed their fishing techniques (fishing nets, hooks). Extraction of marine resources was not only for subsistence of local groups, but also for an intensive transfer of fish toward inland sites (Briones et al. 2005). Local people persisted and exploited marine resources until well after the appearance of the Spanish (Llagostera 1979). The resources collected and captured by prehistoric people are the same that still today play an important role in the fisheries of northern and central Chile.

Artisanal benthic fisheries The present artisanal fisheries in the HCS between 18°S and 35°S are very diverse, comprising ~13 species of algae, ~45 species of invertebrates (molluscs, crustaceans, echinoderms, tunicates) and ~68 fish species (SERNAPESCA 2005). Before the 1980s, fisheries in general were of low level and stable since products only went to local markets and the intensity of exploitation was comparatively moderate (Stotz 1997, González et al. 2006). During the mid-1980s the export of most of the resources was initiated, producing increases in captures, and once the accumulated biomass was used up, the resources remained at low and fluctuating levels (Figure 22). In general, these fluctuations have been interpreted as the classical signs of a badly regulated fishery, as

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described by Hilborn & Walters (1992): discovery of a stock, development of its fishery and subsequent overexploitation and eventually collapse. In order to improve the management status of benthic fisheries (mainly dive fisheries for invertebrates and algae), Chile has established a system of Territorial User Rights for Fisheries (TURF), called Areas de Manejo y Explotación de Recursos Bentónicos (AMERB or Management and Exploitation Areas for Benthic Resources MEABR, which are the diverse names and shortcuts under which they have been described in the literature) (Castilla 1997, Stotz 1997, Castilla & Fernández 1998, Aviléz & Jerez 1999, Bernal et al. 1999, Meltzoff et al. 2002, González et al. 2006). This management tool grants exclusive fishing rights over a defined coastal area to legally established organisations of local fishermen. These areas are exploited according to a management plan, developed by professionals, approved by authority, and then worked by fishermen under the permanent supervision of the administrative authority. This system was initiated in practice by fishermen at the beginning of the 1990s, but legally established in 1997 (Stotz 1997). In general, it has proven to be a good tool to increase stocks and recover depleted fisheries (Stotz 1997, Castilla

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& Fernández 1998, González et al. 2006), albeit it appears to be deficient in particular situations (e.g., Gelcich et al. 2006). Since resource stocks are continuously monitored in management areas, with captures well controlled and registered, it has been possible to begin to understand the underlying nature of fluctuating landings. Contrary to the above-described classical pattern of a badly managed fishery, it has often turned out to reflect the natural variability of the environment of the HCS. Global-scale phenomena, such as EN events, which produce an outburst of some resources and disappearance of others, together with more localised processes of upwelling and current systems, generate a complex, spatially and temporally changing, mosaic of conditions. Fluctuating fisheries are mainly the consequence of this, the fishermen following (and suffering) natural variations, but not always causing them, as generally assumed. The following description of the four most valuable resources fished by artisanal fishermen illustrate this.

Case studies Scallop fishery The scallop fishery can be considered as a ‘boom-and-bust’ fishery, where the ‘boom’ is caused mainly by EN events (Wolff 1987, Stotz 2000, Wolff & Mendo 2000, Stotz & Mendo 2001, von Brand et al. 2006). During EN, recruitment of Argopecten purpuratus is intense, and during following years, given the normal fast growth of the species (Stotz & González 1997), huge stocks of scallops build up, with fluctuations of several orders of magnitude between years (Figure 22A, stock increased with EN 1982–1983; Figure 23A stock increased with EN 1997–1998) (Stotz & Mendo 2001). However, following the increased scallop stocks, the development of similar predator populations and/or the shift of prey preference of predators in response to increased scallop abundance (Ortiz et al. 2003), together with the fishery, leads to an increasing mortality, which finally generates the ‘bust’ (Figure 23A) (Wolff 1987, Jesse & Stotz 2002, León & Stotz 2004). Fishermen are just able to take advantage of part of the EN production before the natural mortality, caused mainly by predation (e.g., by Octopus mimus), but also by mass strandings (González et al. 2001), takes most of the scallops away (Figure 23B). Normal (LN) years are characterised by small and, due to spatially (Aguilar & Stotz 2000) and temporally variable recruitment, fluctuating scallop stocks, which supply a low-level fishery (Figure 22A). Taking advantage mainly of natural recruitment while preventing predation, in northern Chile aquaculture has been able to build up stocks and harvest at levels several times above fishery production (Figure 22A) (Stotz 2000). Surf clam fishery The ‘macha’ Mesodesma donacium fishery also shows boom and bust fisheries along the coast (Figure 22B). In the past, on most sites a relatively stable low-level fishery existed, produced by fishermen working in the intertidal zone. During the mid-1970s fishermen learned also to dive for this resource, which means putting the boat behind the breakers (‘rompiente’), diving, loaded with at least 40–50 kg of lead, through the surge and then working underneath the breakers. This began in Region V (32–35°S). After having depleted the local stocks in Region V, some fishermen came over to Region IV (29–32°S), where the local fishermen quickly learned the same technique. This has generated a tradition of divers, mainly from Regions IV and V, working on the macha along the entire Chilean coast, making use of and depleting macha stocks throughout the country (Figure 22A). However, after the establishment of AMERBs for this resource, other reasons for the depletion became apparent. With the EN 1997–1998 all the macha beds between Arica and Coquimbo, managed conservatively within AMERBs, died off within a few days. The beds in Coquimbo were smothered by mud, washed into the bay by a river flood due to heavy rainfall (Miranda 2001). For 280


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Figure 23 (A) Stock size of the scallop Argopecten purpuratus in an AMERB, (B) the impact of the harvest on the decrease of the same scallop stock, and (C) of a Mesodesma donacium stock in another AMERB, ‘Area de Manejo y Explotación de Recursos Bentónicos’ (MEABR, Management and Exploitation Area for Benthic Resources).

Arica no specific reason was identified, but increased temperatures have been mentioned as a cause of mortality for the same species in Peru (Arntz et al. 1987, 1988). In the Coquimbo area, the only beds left were in Tongoy Bay, which were exploited according to what was considered then a very conservative strategy. However, despite dynamic (and apparently conservative) management, these beds also disappeared (Figure 23C) (Aburto & Stotz 2003). Captures were almost irrelevant compared with the natural decrease (Figure 23C), which was not renewed by recruitment, because larval supply is spatially (Ortiz & Stotz 1996) and temporally very irregular or absent during many years. A similar situation was shown for macha beds managed within management areas in Region VIII, where, due to the almost complete lack of recruitment, the stocks within the AMERBs collapsed by 2004 (Figure 22B) (Stotz et al. 2004). Following their disappearance during or after 281


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EN events, the recovery of local macha beds can be extremely slow, as observed in Peru (Arntz et al. 2006). The ‘loco’ fishery (sold as ‘Chilean abalone’ in international markets) The loco fishery makes use of the muricid gastropod, Concholepas concholepas, described as a top predator in the intertidal systems (Castilla & Paine 1987), but with its fished populations mainly occurring in the subtidal zone (Stotz 1997). In such areas and given that its main prey items are suspension feeders (barnacles, tunicates), the species appears more as a browser, through its prey taking advantage of high PP in the water column near upwelling areas (Stotz 1997, Stotz et al. 2003). Thus, production and consequently fishery landings vary greatly along the coast, with main landing sites coinciding with the most important upwelling areas (Figure 24) (Stotz 1997). Given

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Figure 24 Concholepas concholepas: Variability of larval retention and recruitment along the coast of Regions III and IV, production (average of the period 1985–1995) along the coast of Region IV and harvests of three AMERBs located at coastal areas differing in production (note the scale of y-axis). (Figure adapted from J. González et al. (2004) and Stotz (1997)) Harvest data of AMERB obtained from SERNAPESCA (www. sernapesca.cl); stars show for Region IV the approximate locations of the AMERBs, ‘Area de Manejo y Explotación de Recursos Bentónicos’ (MEABR, Management and Exploitation Area for Benthic Resources).

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its complex reproductive biology, which starts with reproductive aggregations for copula, the laying of egg capsules in which the larvae develop for about 30 days, then followed by a 3-month period of pelagic life, the result is high recruitment variability at temporal and spatial scales, depending on coastal oceanography and topography, and the potential for retention areas (Stotz 1997, J. González et al. 2004) (Figure 24). Larvae settle and metamorphose mainly on adult barnacles covering adult shells (Manríquez et al. 2004) or in association with recently settled barnacles (Stotz et al. 1991b), which may vary greatly between years and sites (Stotz et al. 1991a). This produces a very complex metapopulation structure (J. González et al. 2004, 2006). However, given its long life, with individual growth varying greatly (depending on food availability), fluctuations become partly attenuated when the individuals finally recruit to fisheries at an age between 3 and 4 yr with a size of 10 cm of peristomal opening length (Pérez & Stotz 1992, Stotz & Pérez 1992, Stotz 1997). Thus, while spatial variability of the fisheries is great, the temporal variability at each site becomes partially attenuated (Figure 22C after 1996), with fluctuations at the level of two to three times, which nevertheless means significant changes in the income for fishermen (Figure 24, harvest in AMERBs). The establishment of AMERBs, which was mainly motivated by the closure of the loco fishery at the same time (1989–1992), coincided with a period of increased loco stocks along most parts of the central Chilean coast, probably favoured by the closure and a short period in which fishermen agreed to strictly obey that measure (Stotz 1997). The relationship of the number of fishermen to the size of the management area was at that time generating an acceptable income, but this situation has mostly changed drastically in the following years. At present, many fishermen located at coastal areas with low production are dissatisfied and willing to abandon their AMERBs (compare magnitudes of harvest of AMERBs and number of fishermen in the organisation, shown in Figure 24). Sea urchin fishery The urchin Loxechinus albus fishery in the HCS between 18°S and 35°S is perhaps one of the most variable ones, in this case natural variability probably being increased as a consequence of captures. In central Chile (Regions IV–VIII), urchins are restricted to shallow areas with great surge on the exposed coast, in which the species is partly safe from the predation of the rock shrimp Rhynchocinetes typus (Stotz 2004), a species with abundance and distribution pattern that responds to the variable existence of refuges from its own predators (Caillaux & Stotz 2003). This produces a very patchy distribution of sea urchins. Recruitment occurs mainly inside the adult aggregations, where recruits are protected by the spine canopy (Stotz et al. 1992). Thus, when fishermen, taking advantage of calm weather conditions, reduce the stocks, subsequent recruitment is probably heavily affected. The observation is that once a site is fished for this species, it only recovers about 10 yr later (Stotz 2004). Further north (Regions I–III, mainly between 20°S and 23°S), the fluctuations attenuate slightly and landings increase, as large Macrocystis beds appear (Stotz 2004). There, a more conservative exploitation is carried out inside the AMERBs, conserving patches, and this strategy has allowed an increase in numbers, the sustainable exploitation of which nevertheless still needs to be demonstrated (Figure 22D).

Resource dynamics and management of artisanal benthic fisheries The great natural spatial and temporal variability of resources, and hence fishery production, which characterises the HCS between 18°S and 35°S, poses a great challenge to management. Fishermen, through AMERBs and legal restrictions, are not allowed to move away when a resource goes through a low cycle, as they used to do in the past. Illegal movement, though still occurring, is also increasingly prevented in practice by conflicts with the respective local fishers. This produces 283


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the risk that fishermen continue fishing on a weakened population, perhaps producing its collapse to a degree where recovery is severely compromised. Additionally, the distribution of fishermen along the coast within each region in general is not well adjusted to its spatial productive variability, thus creating great differences of income along the coast, with fishermen’s organisations located at productive sites getting increasingly richer, and others at unproductive sites (as shown with harvest in AMERBs in Figure 24) getting increasingly poorer, which could be a source of conflicts (Gelcich et al. 2005, Stotz & Aburto 2006). The challenge is to advance to an integral management strategy in which fishermen, instead of rotating between fishing zones along the coast, rotate between fisheries of different resources or among other related activities (processing, tourism, etc.) in order to produce income during years or months of poor production. This means a change from a specialist to a generalist strategy, but with very strict control of their numbers, such that they are well adjusted to local production levels. Complementary stock enhancement, using biological and ecological knowledge and aquaculture experience, might help to mitigate natural fluctuations (Stotz et al. 1992, Zamora & Stotz 1994, Pacheco & Stotz 2006). In order to avoid the risk of increased fishing pressure (caused by restricting movements of fishermen) on already weakened populations, the establishment of a network of small reserves along the coast should be considered, which should also aid in reducing natural recruitment variability (see also Stotz & Aburto 2006).

Bioeconomic aspects of industrial crustacean fisheries An important crustacean bottom-trawl fishery exists in northern and central Chile, particularly between Regions II and IV. This fishery, which includes the nylon shrimp (Heterocarpus reedi), yellow squat lobster (Cervimunida johni) and the red squat lobster (Pleuroncodes monodon), is conducted on the continental shelf at depths ranging from ~100 to 600 m. The exploitation of these resources began in the 1950s as fauna accompanying catches of hake Merluccius gayi and was subsequently developed into a fishery specific for these crustaceans (Arana & Nakanishi 1971, SUBPESCA 1999a,b). A growth phase for these fisheries occurred between 1958 and 1968, during which annual landings of nylon shrimp reached 11,000 t. Subsequently, landings declined to less than 3000 t in 1979. Between 1980 and 1986 landings remained below 4000 t, except for 1983, when landings exceeded 6000 t. The last period is interesting in its analysis since it coincided with three fundamental factors: (1) from 1979 to 1982 the exchange rate (Chilean peso/U.S. dollar) was 39 Chilean pesos/U.S. dollar, which produced a depressive economic effect, resulting in the collapse of fisheries companies; (2) in 1983 the occupation of new fishing grounds south of Region IV increased the catch levels above those previously obtained; and (3) from 1984 to 1986 a reorganisation of the fishing fleet occurred due to the opening of the fishery for the red squat lobster (Pleuroncodes monodon) along with an increase in world prices for this resource. Between 1989 and 1991 the fishery was closed in Regions V and VIII, but it remained open in northern Chile. In this northern area after 1986 there was an increase in landings of squat lobster reaching 10,620 t in 1995. After this year there was a clear decrease in these landings, reaching 4000 t in 2000. Since 1995, the three species are subject to catch quotas per boat owner, based on a total allowable catch (TAC), which is determined annually, following direct resource evaluations. A period of difficulty in the crustacean fisheries due to decreasing stocks began in 1999 in northerncentral Chile (SUBPESCA 2005a,b). The TAC per Fisheries Unit of nylon shrimp (Regions II–VIII) dropped over four subsequent years from 10,000 t in 1997 to only 4000 t in 2000 (i.e., a 60% decline). The TAC of the fishery for the yellow squat lobster in Regions III and IV dropped from 6000 t in 1998 to 4000 t in 2000, representing a reduction of 33% during that time period. This substantial reduction in landings caused a major decline in the activities of the fishing fleet and the packing plants. During 2002, some companies closed, with important losses in employment producing socioeconomic impacts in the regions affected (Pérez 2003, 2005). This brief overview 284



underscores that the fishery of these benthic crustaceans is highly dynamic, driven not only by biological factors, but also by administrative decisions and economic considerations. In the following section some relevant information on the biology of these three species is provided before presenting a case study highlighting the relationship between catch-based stock estimates and species biology.

Basic biology of nylon shrimp and squat lobsters The main bathymetric distribution of these species shows a strong overlap with the OMZ between ~50 and ~600 m, where they occur on gravel and mud bottoms (SUBPESCA 1999a,b). Their latitudinal distribution ranges from 25°S to 39°S (Heterocarpus reedi), from 6°S to 40°S (Pleuroncodes monodon), and from 29°S to 38°S (Cervimunida johni) (Acuña et al. 1997, 1998, Quiroz et al. 2005). Little is known about their food resources. All three species themselves are important prey organisms for demersal fish predators (e.g., flounders and hake), and it has been discussed that the OMZ may represent a refuge from predation (Villarroel et al. 2001). Most information about the biology of these crustacean species is based on analysis of specimens obtained from commercial or research catches. Heterocarpus reedi Available data suggest seasonal migrations with the main concentrations of shrimp found at great depths (>400 m) in austral summer, at shallow depths during austral winter (20,000 embryos. Males reach substantially greater sizes and weights than females, and average size at first maturity (of females) varies substantially among years (Acuña et al. 2005). It is generally assumed that each female produces only one clutch per year, and Arancibia et al. (2005) suggested that C. johni is a slow-growing species, with individuals living up to 11 yr. Pleuroncodes monodon Based on the main occurrence of ovigerous females it has been suggested that mating occurs mainly in late summer/early autumn (Palma & Arana 1997). Embryo incubation extends from April to November, with the highest proportion of ovigerous females observed in July–August (Palma & Arana 1997). All embryonic developmental stages were found between June and October (Palma & Arana 1997), suggesting that not all females mate at the same time. The first planktonic larvae 285


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appeared in June, and early developmental stages were found in the plankton until December (Palma 1994), supporting the suggestion that reproduction is not synchronised among females. Gallardo et al. (1994) reported the first benthic stages in March with very high densities of recently settled squat lobsters in April. Roa et al. (1995) identified nursery areas in places with very low oxygen concentrations from where juveniles migrate to nearby adult habitats. It is generally assumed that females produce only one clutch per year, but the extended reproductive period and the apparently short incubation period of 2–3 months (Palma & Arana 1997) could allow some females to produce more than one clutch per year. Fecundity has been estimated at 1800–34,000 eggs in Region VIII (Palma & Arana 1997), and most females reach sexual maturity at carapace lengths of ~25 mm. Red squat lobsters are estimated to live from 5 to 10 yr (Roa & Tapia 1998).

Biology and stock estimates As can be seen from the preceding section, information on the biology of the nylon shrimp and the squat lobsters is mainly based on demographic data (size at first reproduction, fecundity, sex ratio, per cent ovigerous females, embryo developmental stages), but little information is available on their behaviour. A close analysis of monthly landings suggests possible seasonal changes in behaviour during the year and indicates that it is important to incorporate this information in stock estimates (Pérez 2005). Pérez (2005) proposed that the biomass of the nylon shrimp showed well-defined cycles of availability (sensu Menge 1972), which were not related to its true abundance in biomass (sensu Menge 1972). The first cycle (termed the ‘short cycle’) occurs from September to December and is characterised by a decrease in availability (minimum in October), followed by a recovery (maximum availability in December). This leads to a cycle of greater length (termed the ‘long cycle’), characterised by a decline in available biomass, reaching a minimum in April and followed by an increase again reaching a maximum in August. Reasons for these cycles are not clear, but it has been suggested that they are related to the moult and reproduction. In this sense, a sudden decrease in availability could be due to recently moulted individuals hiding from predators (Pérez 2005). The total biomass of Heterocarpus reedi in Region IV in September 1997 was estimated to be about 5000 t, but not all of these shrimp were susceptible to fishing gear, resulting in an available biomass (to the fishery) of 4200 t (for details see Pérez 2005). The estimated total biomass decreased to 2000 t in August 2000 (Week 143), representing a decline of 60% from that estimated at the initiation of the simulation (Figure 25A). However, the percentage variation in the catch per unit effort (CPUE) was not proportional to the decrease in abundance within periods of maximum availability (Figure 25A). At the beginning of the study period, the CPUE was 0.59 t haul−1, and in August 2000 this decreased to 0.45 t haul−1, representing a decrease of only 24%. In recruitment, at each fishing station in 1997–1998 and 1998–1999 the difference between the control curves and availability allowed the identification of two time windows for recruitment, one of lesser intensity in December and one of greater intensity extending through June and July. This seasonal variation in the availability of the resource to the fishing gear has a direct influence on the population estimates and consequently the determination of TACs. Therefore it was recommended to include this biological information in stock evaluation and to conduct the corresponding research cruises during periods of maximum availability of the resource (Pérez 2005).

Crustacean fisheries and economic considerations Fisheries respond not only to fluctuations in stock abundance but also to economic considerations. From the economic perspective, the annual TAC system has produced an increase in the fishing 286



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effort, measured in terms of hauls, resulting in a decrease in the expected economic benefit based on administrative measures. Pérez (2003) explored the bioeconomic effect produced by the decrease in CPUE in the nylon shrimp and yellow squat lobster fisheries in northern Chile during 1997–2000. A biological-technological simulation model was used by Pérez (2005), which took both physical and biological variables into account. An economic submodel was incorporated into this model in order to carry out an integrated analysis of the crustacean fisheries, which included the subsector involving catches and their processing. The economic results obtained, when integrated with the catch results and size of the stocks, allowed the dynamics of this fishery to be explained in Regions III and IV during 1997–2000, when the catch increased by 21%, and final production increased by 24% (measured as frozen tails), although the variable costs increased by only 11% (Figure 25B). In this same period the fisheries costs increased by 117% and the total production cost increased by 93%. The results showed that part of the economic benefit was lost due to the effect of a decrease in the biomasses of both resources and an excessive increase in the average production costs (due to increased costs of fishing and extraction). This example underscores the importance of incorporating economic reference points in addition to biological (biomass) and fishery variables (CPUE and catch). Furthermore, basic biological data (larval ecology, settlement biology, habitat requirements, seasonal migrations, and mating behaviour) still need to be revealed for these three crustacean species. At present, very little is

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known about the influence of oceanographic factors on recruitment success and stock dynamics of crustaceans from the continental shelf off northern-central Chile. Most current information suggests that their life history is driven by seasonal factors, and presently available data do not permit an examination of the effects of interannual variability in oceanographic conditions (e.g., ENSO) on their population dynamics. Future studies should address these questions in order to achieve a sustainable fishery.

Pelagic fisheries and fisheries management, 1980–2005 The Chilean purse seine fleet mainly exploits five pelagic fish species: anchovy (Engraulis ringens), jack mackerel (Trachurus murphyi), chub mackerel (Scomber japonicus), and Pacific sardine or pilchard (Sardinops sagax) and common sardine (Strangomera bentincki). There were doubts about the taxonomic status of jack mackerel (Stepien & Rosenblatt 1996, Oyarzún 1998), but a recent molecular study revealed that the name Trachurus murphyi should be conserved (Poulin et al. 2004). In the official landing statistics it is also recorded as T. murphyi, but the Technical Reports of the Undersecretary of Fisheries refer to it as T. symmetricus.

History of the catches Aguilar et al. (2000) described the development of the Chilean pelagic fisheries based on the landings of anchovy, jack mackerel and pilchard from 1970 to 1995. Herein the period from 1995 to 2005 (also known as the ‘regulated’ period) is included, which adds two additional species: chub mackerel and common sardine, and their landings since 1980. The total annual landings captured by the Chilean purse seine fleet during the study period showed a steadily increasing tendency from 1980 until 1995, from 3.4 million t to almost 6.9 million t, driven mainly by the increase in jack mackerel landings and secondarily by anchovy landings, which replaced the Pacific sardine, the most important species during the early 1980s (Figure 26). After that, the total annual catch decreased to around the same level found at the beginning of the study period, with the exception of 1998, when both species simultaneously showed a sharp decline in their landings. However, as stated, these are ‘regulated’ landings since catches are the result of fixed total annual quotas and therefore not necessarily representative of resource availability. Administratively, four of the five species included in this analysis are latitudinally assigned to fisheries units, each of which comprises two to five administrative zones: Regions I–II (18°25′S–26°06′S), Regions III–IV (26°06′S–32°18′S), Regions V–IX (32°18′S–39°37′S) and Regions X–XII (39°37′S–56°30′S), and there is a strong difference in landings between the respective fishery units (Figure 26). Therefore, annual landings are analysed by species and these latitudinal fisheries units to visualise their importance in each of them. In the following text each Fisheries Unit is defined by the Regions it contains, e.g., Fisheries Unit I-II. Landing statistics of anchovy show the typical characteristics of engraulids, with successive increases and decreases, attaining highest landings of 2.7 million t in 1994, but landings usually do not surpass 1.5 million t. Anchovy landings are most important in northern Chile (Fisheries Unit I–II) (Figure 27). In contrast to all other species, jack mackerel showed a continuous increase until 1995, when annual landings reached 4.4 million t, after which catches continuously decreased again. Throughout this period, catches in central-southern Chile (Fisheries Unit V–IX) had overriding importance (Figure 27). The highest landings of chub mackerel were attained in 2003 with 0.5 million t. This species has shown a steady increase in landings during the study period. At the beginning of the period, landings of chub mackerel were mainly in northern Chile (Fisheries Unit I–II) but since the year 2000 also increased in central-southern Chile (Fisheries Unit V–IX) 288



Figure 26 Total annual landings for the five most important pelagic species caught by the Chilean purse seine fleet during the period 1980–2005; left panel shows SST anomaly, total landings and landings of the five species; right panel shows average landings per latitude and year in each of the four fisheries units, calculated for the time period 1980–2005.

(Figure 27). The highest landings of Pacific sardine were attained in 1985 with almost 2.9 million t. During the 1980s this species was the most important small pelagic fish captured in Chilean waters but catches continuously decreased during the late 1980s, reaching very low levels in the mid1990s. The Pacific sardine was most important in northern Chile (Fisheries Unit I–II) but gained proportionally in importance in central-southern Chile (Fisheries Unit V–IX) during the early 1990s (Figure 27). The highest landings of common sardine were attained in 1999 with 0.75 million t. During the 1980s this species was very scarcely captured in Chilean waters but from then on showed a steady increase (Figure 27). Landings of the common sardine were most important in the centralsouthern Chile (Fisheries Unit V–IX).

Relationships with oceanographic variations The coastal areas off the Chilean coast are known for being typical of an EBC system, where upwelling is a characteristic oceanographic feature. Fonseca & Farías (1987) and other authors described the presence of active upwelling centres in several areas of the Chilean coast, like Iquique (Fuenzalida 1990) and Antofagasta in Fisheries Unit I–II (Blanco et al. 2001), Caldera and Coquimbo (Acuña et al. 1989) in Fisheries Unit III–IV, Valparaíso (Johnson et al. 1980) and Concepción (Cáceres & Arcos 1991) in Fisheries Unit V–IX. 289


Niño 3.4 index

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3 2 1 0 −1 −2 −3

2.500 2.000

ANCHOVY

1.500 1.000 500 0 4.500 JACK MACKEREL

3.500

Landings (1000 t)

2.500 1.500 500 0 350 CHUB MACKEREL

250 150 50 0 3.000

PACIFIC SARDINE 2.000 1.000 0 800 600

COMMON SARDINE

400 200 0 1982 1984 1986 1988 1990 1992 1994 1996 1998 2000 2002 2004

Figure 27 Total annual landings for the five most important pelagic species caught by the Chilean purse seine fleet during the time period 1980–2005 in the respective fisheries units; grey dots represent Fisheries Unit I–II, open dots Fisheries Unit III–IV, grey triangles Fisheries Unit V–IX, and open triangles Fisheries Unit X–XII.

Yáñez et al. (1996) conducted a survey to assess the possibility of introducing the use of SST, obtained from NOAA (National Oceanic & Atmospheric Administration) satellite data, for the small pelagic fisheries resources and found significant relationships between species yields and TGRs. Jack mackerel yields were largely related to strong TGRs next to oceanic waters, while anchovy and common sardine yields were mainly associated with the development of coastal upwelling events. Comparison with the SST anomalies shows that landings of anchovy negatively correlate 290



with SST anomalies (Yáñez et al. 2001); also chub mackerel landings seem to correlate with SST (Figure 27). In contrast, interannual variations in the landings of the other three species seem to be largely independent of variations in SST (Figure 27). The stabilisation of maximum anchovy landings between 1.5 and 2.7 million t during the 1990s and the parallel decline of the landings of the Pacific sardine in the HCS (and the entire southeastern Pacific) reflects another regime shift from the warm ‘sardine regime’ to a cool ‘anchovy regime’ (Chavez et al. 2003, Alheit & Niquen 2004, Halpin et al. 2004). These regime shifts occur on multidecadal scales and are probably related to the PDO, but the mechanisms driving these changes are not yet well understood. Silva et al. (2003) studied the relationship between chl-a concentration, SST, and fishing yields of anchovy, Pacific sardine and jack mackerel in northern Chile during summer and autumn 1999. CPUE superimposed over SST images confirmed that the anchovy has a more coastal distribution than Pacific sardine and jack mackerel, being found in the frontal zone of coastal areas. In the southeastern Pacific, the jack mackerel is a heavily exploited pelagic species, and its presence in the HCS in autumn and winter is assumed to be mainly due to an inshore feeding migration (Bertrand et al. 2004). During warmer years, jack mackerel may immigrate into coastal waters where they are thought to exert high predation pressure on anchovy (Alheit & Niquen 2004). Changes in SST associated with EN events may also affect the migration pattern of jack mackerel, which needs to be taken into account in stock assessment and management plans (Arcos et al. 2001).

Management of pelagic fisheries According to Aguilar et al. (2000) the traditional method used to conserve fish stocks and prevent overfishing is to set a TAC for the fishery. Typically, TACs aim to restrict fishing effort to its MSY (maximum sustainable yield) level. Once these ‘safe biological limits’ are reached, fishing is prohibited. But TACs do not, by themselves, address the overcapitalisation issue. Consequently, many fisheries economists recommend that the designated TAC is distributed to industry participants in the form of individual transferable quotas (ITQs), quasi property rights that restrict additional access to the fishery. Under these rules, failure to acquire an ITQ effectively forces vessels out of the fishery, thereby reducing fishing effort and increasing harvesting efficiency. The Chilean General Law of Fisheries and Aquaculture considers three types of access to the fishery: (1) Full Exploitation Regime, which includes setting an annual quota or TAC by fishing unit, which can be temporally divided during the year and also modified once during that same period (once the full exploitation regime is assigned to a given species no more fishing vessels are allowed to enter the fishery); (2) Recovering Fishery Regime is the fishery under a state of overexploitation, subject to a capture ban of at least 3 yr, to obtain its recovery and where an annual quota (or TAC) can be established; and (3) Early Developing Fishery Regime, which is a demersal or benthonic fishery with open access where an annual quota can be established and where no fishing effort is applied or if it is done it is less than 10% of this quota. In 1993 the first pelagic fisheries in Fisheries Unit I–II were assigned the status of Full Exploitation Regime, and by 2000 this had extended to four of the five fisheries (anchovy, jack mackerel, Pacific sardine and common sardine) and to all fisheries units. The chub mackerel is one of the few commercial species that is still under ‘open access’, with no regulation to date. Alternative management tools have been developed and used in the last years: in 1998, the use of a geographical positioning system for the industrial and artisanal fleets was established; later in 2001 the Límite Máximo de Captura por Armador (LMCA, maximum capture limit per owner) was introduced for the industrial fleet in the fisheries for the small pelagics anchovy, sardines and jack mackerel in all fishery units, which is essentially an ITQ, and at least had the effect of reducing the fishing fleet. The LMCAs are determined using captures (1997–2001) and corrected hold capacity (authorised hold capacity in cubic metres times authorised area length divided by total 291


MARTIN THIEL ET AL.

length of fishery unit, Law 19.173). This action was set to last until 2002, but was later extended until 2012 (Law 19.849). Finally, in 2004 a similar management tool was established for the artisanal fleet, the Regimen Artesanal de Extracción (RAE, Artisanal Capture Regime), which in this case assigns the artisanal fraction of the Annual Global Quota to one or more artisanal fishermen’s organisations in each fishery unit, which in turn divide it between members. As becomes evident from the preceding paragraphs the currently valid management rules (full exploitation, Global Annual Quota, LMCA, etc.) have been implemented when landings started to decrease or stocks had already dramatically declined. Therefore, fisheries of small pelagic fisheries in the HCS off Chile are presently managed at much lower population levels than they used to be in previous decades. In fact, stocks of anchovy and common sardine in central Chile are considered to be overfished (Cubillos et al. 2002). The interaction between administrative effects on landings and fishery-induced impacts on stocks complicate the detection of direct relationships between environmental factors and fish stocks. In addition to fishery-independent stock surveys, studies of the basic biology of individual species, of biological interactions (predators, competitors, food), and of the role of climatic and coastal oceanography are required in order to improve our understanding of the factors driving the population dynamics and distribution in northern-central Chile.

Aquaculture According to the latest reports published by the Food and Agriculture Organization of the United Nations (FAO 2006), the contribution of aquaculture to the world supply of fish and shellfish “continues to grow faster than any other productive sector of animal food origin”. The world production of aquaculture registered in 2002 rose to 51.4 million t (including aquatic plants), with Asian countries producing 91.2% of this. In 2002, Chile contributed only 1.4% of the world’s total production, but it is among the 10 countries showing the fastest growth in aquaculture production. Fisheries and aquaculture are for Chile one of the most important economic activities with a total income of U.S. $2,246 million in 2003, of which aquaculture contributed U.S. $1600 million. The largest share of the Chilean aquaculture production (80%) is from southern Chile (41–46°S), with salmon and mussels and to a lesser extent oysters, seaweeds and more recently red abalones (Haliotis rufescens) being the most important resources (FAO 2006). Aquaculture activities along the coast of northern-central Chile (18–35°S), although not reaching the same levels as in southern Chile, have been continuously growing during the past two decades (FAO 2006). Given that the shorelines of northern-central Chile are relatively exposed to wave action and fully subjected to the effects of ENSO, it is particularly important to take these factors and interannual variability in oceanographic conditions into account. In fact, all aquaculture centres in northern and central Chile are located in relatively sheltered bays. The main natural resources cultured in northern-central Chile are scallops (Argopecten purpuratus) and seaweeds (Gracilaria chilensis), and their natural populations are also exposed to strong seasonal and interannual variations. Small-scale culture of some other species (bivalves Mesodesma donacium and Tagelus dombeii, gastropods Concholepas concholepas and Fissurella spp., sea urchins Loxechinus albus) has also been attempted but has not reached a commercial stage, mainly due to biological (long larval periods) and logistic (food supply) reasons. Several introduced species are also cultured in northern Chile, namely the Pacific oyster (Crassostrea gigas), which has been cultured on a small scale since 1970, and during recent years increasingly abalones (Haliotis rufescens and H. discus hannai). The main resource cultured during the past two decades in northern Chile is the scallop Argopecten purpuratus. Culture centres are located in bays, namely Isla Santa Maria and Bahía Mejillones (22°S), Bahías Caldera, Calderilla and Inglesa (27°S), and Bahías Guanaqueros and Tongoy (30°S). Standing stocks and productivity in these bays are below those of similar bays in 292



Peru (Uribe & Blanco 2001). It has long been recognised that natural scallop stocks vary together with ENSO variations, mainly because settlement of small settlers (spat) is strongly favoured during EN events (Narvarte et al. 2001). Due to this relationship, the local scallop industry is affected in a positive way by EN, which can lead to an increase in spat collection by >300% (Illanes et al. 1985). In a research project on the effect of environmental factors on scallop culture, data of water temperature, gonad indices, larval abundance and recruitment (spat collection) were gathered between 1981 and 1984 in Tongoy Bay (Illanes et al. 1985). The EN event led to a temperature increase of 2°C above the pre- or post-EN levels at the sea surface and of 2.5°C at the bottom (20 m). During the EN period, the Gonad Index of adults registered a maximum (25) and a minimum (6.8) with a massive evacuation of gametes, a situation never observed during normal years (Illanes et al. 1985). Similar observations were made in southern Peru by Wolff (1988), who concluded that A. purpuratus is a continuous spawner with spawning peaks during late austral summer and autumn (February–March). Under the unusually high water temperatures during EN 1982–1983 (2–2.5°C above the normal temperatures), the recuperation time between two spawnings was shortened, indicating that maturation was accelerated and spawning probably intensified under these conditions. This interpretation was confirmed by the highest larval concentration found within the period. Wolff (1988) stated that it is not clear along the Peruvian coast that past EN favoured the scallops stocks. The EN 1972 apparently did favour stocks, as catches were significantly higher than during the 3 yr before and after this event, but the weaker EN 1976 did not have the same effect. In 1979, which was a ‘normal’ year, the catches still exceeded the catches of 1972. Limo (in a personal communication to M. Wolff) reported enormous numbers of A. purpuratus during the strong EN 1925 in Ancon Bay, north of Lima. In order to reduce the dependency on natural variations in supply of small recruits, intense efforts have been undertaken to produce settlers in the laboratory (Uriarte et al. 2001), but this only satisfies part of the requirements for scallop spat. The fact that natural spat is still obtained at much lower costs than spat produced in the laboratory may have led to the strong increase in spat collectors observed during recent years in Tongoy Bay (Figure 28). Possibly, the decreasing spat collection efficiency between 1998 and 2001 (when the total number of spat collectors in the bay exceeded 2.5 million bags) is indication that the carrying capacity is reached and that spat collectors are starting to compete for the available settlers. Culture of Gracilaria chilensis in northern Chile is mainly developed on shallow soft bottoms in sheltered parts of the bays (e.g., Pizarro & Santelices 1993). Since sheltered bays are relatively scarce along the cost of northern and central Chile, aquaculture of this seaweed in this area does not reach the amounts harvested in southern Chile. Edding & Blanco (2001) observed a decrease in productivity and yield of ‘agar’ from G. chilensis cultured in Region IV (29°59′S) during EN 1997–1998. This may be more related to the decrease in nutrient concentrations and increase of visibility instead of the higher water temperatures. Furthermore, these authors cited González (1998), who reported that increased wave action during EN resulted in an important reduction of biomass. Growers of G. chilensis in Region IV also reported heavy damage to culture fields on shallow subtidal soft bottoms caused by storms during EN (R. Rojas personal communication). However, Santelices et al. (1984) stated that recovery of Gracilaria sp. after storms is unexpectedly fast due to regrowth of thalli from the portions buried in the sand. Overall, the production of G. chilensis in northern Chile does not seem to be severely affected by ENSO, but is rather determined by stock density and harvesting frequency (Pizarro & Santelices 1993). The relatively limited interannual variability in extracted and cultured biomass (Figure 28) further supports the suggestion that management rather than environmental conditions drive the production of G. chilensis in northern Chile. During recent years first attempts have been made to culture large kelps from northern and central Chile (e.g., Edding et al. 1990, Edding & Tala 2003) in order to satisfy the growing needs of the abalone culture. This is considered particularly important since EN can have dramatic impacts 293


3 2 1 0 −1 −2 −3 1982

A

1984

1986

1988

1990

1992

1994

1996

1998

2000

2002

2004

Scallop seeds (n bag−1)

600

B

500

30 25

N° of scallop seeds by collector bag N° of collector bags

400

20

300

15

200

10

100

5

0 1982

1984

1986

1988

1990

1992

1994

1996

1998

2000

2002

2004

20 Harvest (1000 t)

2006

16

2006

Collectors (100,000 bags)

Niño 3.4 index

MARTIN THIEL ET AL.

0

C

Natural Gacilaria production Cultivated Gacilaria production

12 8 4 0 1982

1984

1986

1988

1990

1992

1994

1996

1998

2000

2002

2004

2006

Figure 28 (A) SST anomalies for the time period 1982–2006 (from http://iri.columbia.edu/climate/ENSO); (B) average number of scallop Argopecten purpuratus settlers per collector bag (individuals bag−1) and the number of collector bags placed in Bahia Tongoy (30°S); (C) harvest of Gracilaria chilensis between 18°S and 35°S from natural banks and from culture beds. (Source for A and B: SERNAPESCA, Region IV, Chile.)

on the populations of large kelps and other macroalgae in northern Chile (Vásquez 1999). Kelp aquaculture is presently in a developmental phase and has not yet achieved economic importance. Other resources currently being investigated for aquaculture in northern Chile are bivalves from sandy beaches (Mesodesma donacium, Tagelus dombeii). Culture of these bivalves is aiming at stock repopulations after extinction of local stocks (see also Artisanal benthic fisheries, p. 278ff.). Additionally, there are several introduced species that are presently cultured in northern Chile. The Pacific oyster (Crassostrea gigas) was originally introduced in northern Chile in the 1970s, and despite fast growth rates, aquaculture activities then moved to southern Chile because there culture costs are cheaper (bottom culture vs. suspended culture in northern Chile). The production in northern Chile is marginal, and there are only two oyster companies remaining. During the first half of 2004 these produced only 935 t, which represents a 46.8% decrease compared with the production during the first half of 2003. The influence of ENSO on the production of oysters is not well known, but since Pacific oysters have a wide range of temperature tolerance EN effects may be minor (or positive). 294



Other species introduced to the coast of Chile are abalone, originating from California and Japan. These have been mainly cultured in land-based facilities, but due to an increasing production and limited holding capacities on land, sea-based culture (as already established in southern Chile) is also considered for northern-central Chile. The production of abalones increased from 1 t in 1998 to 342 t in 2005, and for the year 2006 it is expected that the Chilean abalone industry will produce >500 t. Only in Region IV, currently five abalone production centres are established and an additional five centres have solicited permits to initiate new aquaculture activities during 2006. The present abalone production in Chile is mainly based on the red abalone (Haliotis rufescens). However, the Japanese abalone (H. discus hannai), which has a better market value, is also raised, but to a lesser degree since culture technology has higher requirements (and costs) than those for red abalone. Although the land-based abalone culture is not directly affected by variations in environmental conditions, ENs may have severe effects on abalone culture because they can produce strong impacts on the population of large kelp, the main food resource presently used in abalone culture. The lack in supply of fresh food algae may produce serious bottlenecks in the culture of abalone. Some of these problems are occurring presently and the National Fisheries Service (SERNAPESCA) is concerned with the overexploitation of kelp, restricting the extraction from natural kelp beds and promoting research for cultivation and management of seaweeds. This scenario presents important challenges for applied research in the near future. In general, aquaculture in northern and central Chile does not reach the levels it has in southern Chile. Some of the main reasons for this are related to the fact that the coast of northern Chile is (1) mostly exposed to wave action and (2) is strongly affected by important interannual variations in oceanographic conditions. Future efforts should probably focus on the development of landbased culture facilities and integrated systems where animals and algae are produced in combination (Chopin et al. 2001).

Conservation of marine biodiversity and Marine Protected Areas The HCS extending from Ecuador to southern Chile is considered as one of the large marine ecosystems for high-priority attention (Boersma et al. 2004). The growing use of coastal areas by human activities is also increasingly threatening marine biodiversity in the HCS. The most important threats are overfishing, aquaculture, pollution by sewage and mining activities, runoffs of chemicals used for agriculture, oil spills and tourism activities (Vásquez et al. 1999, Fernández et al. 2000). All countries, including Chile, that have ratified the Convention on Biological Diversity (CBD) treaty agreed to develop a network of Marine Protected Areas (MPAs, as defined by IUCN 1994) to ultimately protect 10% of the marine environments by 2012, based on an ecosystem approach (Wood 2006). However, the actual establishment rate of MPAs (4.5% annual increase) reveals that the work plan is unrealistic and will not be achieved before the second half of this century (Wood 2006). The Chilean case is a paradox as its economic exclusive zone (EEZ) corresponds to 17.8% of the Latin-American EEZ and is three times larger than its terrestrial territory (~18% of which is already protected; Pauchard & Villarroel 2002). However, only 0.03% of the Chilean EEZ (0.67% of the territorial sea) is protected as MPAs (CONAMA personal communication). Further, only 14% of the surface area of these MPAs are located within the HCS of northern and central Chile (18°S to 41°S). Considering that ~95% of the Chilean population is located between 18°S and 41°S and the growing human impact in coastal areas, this zone represents one of the greatest challenges for marine conservation. The Chilean fisheries management policy, through creation of Management and Exploitation Areas (AMERBs) for benthic resources from coastal habitats, is focusing on economically important 295


MARTIN THIEL ET AL.

target species, largely ignoring habitats, ecological interactions and other important ecosystem components (e.g., species dispersal). Although management areas can provide nursery grounds for target species (e.g., the Chilean abalone, Concholepas concholepas, keyhole limpets Fissurella spp. and the red sea urchin Loxechinus albus), this approach is often criticised for its weakness in providing a long-term ecological and economic viability and uncertainty in its efficiency (NRC 2001). The establishment of MPAs has proven to be a useful tool to achieve conservation and preservation goals (e.g., resources, communities, habitats), either as no-take MPAs or multiple-use MPAs (MUMPAs) (NRC 2001). Among the different tools existing in the Chilean legislation for protecting coastal areas, two types of no-take MPAs (marine reserves and parks) were foreseen in the law since 1989 but have only recently been established (Morales & Ponce 1997, Fernández & Castilla 2005). According to the Chilean laws, marine reserves are not focused on ecosystem protection but rather on exploited resources and their habitats and eventually may allow partial extractions if stocks reach very high levels of abundance. Three marine reserves are located between 18°S and 41°S (Table 6): La Rinconada to preserve a genetic stock of the scallop Argopecten purpuratus and Isla Choros-Damas and Isla Chañaral for allowing the recovery of several overexploited benthic invertebrates (and their habitats). Furthermore, two MUMPAs, Isla Grande de Atacama and Lafken Mapu Lahual, have been created by the National Environmental Agency (CONAMA) with international private funding (GEF) as two of the three Chilean MUMPAs. In contrast to the marine reserves, these MUMPAs aim at the conservation of biodiversity integrating socioeconomic interests by creating not only no-take areas, but also areas where fishery and outdoor activities (e.g., diving, ecotourism) are permitted. While no-take zones have not yet been established in the MUMPA Isla Grande de Atacama, they have recently been identified (C. Gaymer et al. unpublished data) and will be declared in the near future. Finally, a small no-take MPA is located at Las Cruces, mainly for scientific purposes. Although very few MPAs exist in Chile, initiatives for conservation are facilitated by the Chilean law, which established an exclusive zone for artisanal fisheries within 5 nautical miles from the shore (e.g., prohibits trawling). The MPAs Isla Choros-Damas, Isla Chañaral and Isla Grande de Atacama in northern Chile have been subjectively chosen through the use of expert criteria, based on the presence of some important ecological communities representing the HCS. The subtidal zone of these three MPAs is characterised by kelp beds of Lessonia trabeculata, L. nigrescens and Macrocystis integrifolia and several invertebrate populations overexploited during decades, and now vulnerable, such as the Chilean abalone Concholepas concholepas and the red sea urchin Loxechinus albus (i.e., extraction prohibited). Moreover, these MPAs are the habitat of some emblematic and/or endangered species, such as the bottlenose dolphin Tursiops truncatus, the sea otter Lontra felina, the Humboldt penguin Spheniscus humboldti and the Peruvian diving petrel Pelecanoides garnotii, and Isla Grande de Atacama is the southernmost site where the wedge-rumped storm-petrel Oceanodroma tethys kelsalli occurs. Biological corridors (Kaufman et al. 2004) permitting those species along with other seabirds to travel between these three MPAs are not included in the present MPA design. Additional MPAs between the marine reserves and the MUMPA would help to ensure undisturbed migration of marine birds and mammals. Ultimately, given the use of both terrestrial and marine environments by some species (e.g., Humboldt penguins) (Fariña et al. 2003b, Ellis et al. 2006), the selection of priority sites for conservation/preservation should integrate both environments, looking for common hot spots that would increase efficiency and reduce conservation costs. MPAs offer a management tool to preserve hot spots of native species diversity; however, these hot spots can be strongly affected by invasion of exotic species which could compromise the effectiveness of MPAs (Byers 2005, Klinger et al. 2006). Introduction of invasive species in Chile becomes a serious concern with the increase of aquaculture (Castilla et al. 2005a). The Isla Grande de Atacama MPA is south of a bay (Bahia Inglesa) where intense scallop culture takes place and where the highest density of the exotic seaweed Codium fragile for the Chilean coast has been 296

Marine Reserve MUMPA

La Rinconada

297 Marine Reserve No-take MPA MUMPA

Isla Choros-Damas

Las Cruces

Lafken Mapu Lahual

Marine Reserve

Isla Chañaral

Isla Grande de Atacama

Status

Name

2005

2005

2005

2005

2004

1997

Establishment year

33°30′ S 40°43′ S

29°15′ S

29°02′ S

23°28′ S 27°10′ S

Mean latitude

44.6

0.15

25

4.3

35.5

3.4

Size (km2)

791

470

20

213

404

Distance to next MPA (km) to south

Scientific research Biodiversity protection

Overexploited species recovery

Overexploited species recovery

Biodiversity protection

Genetic reserve

Main conservation goals

Conflicts with indigenous people

Illegal extractions

Illegal extractions

Illegal extractions

Invasive species

Illegal extractions

Major threats

Kelp beds, marine mammals and birds

Kelp beds, barrens, marine mammals and birds Kelp beds, barrens, marine mammals and birds Kelp beds, barrens, sea grass, marine mammals and birds Kelp beds, barrens

Scallop bed

Main communities and target groups

Table 6 Main characteristics of marine protected areas (MPAs) in the Humboldt Current System (HCS) between 18° and 41°S

Every year since 1982 1 in 2006

1 in 1999

1 in 1999

Every year since 1997 1 in 2002

Biological survey




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reported (Neill et al. 2006). Recently, patches of C. fragile have been observed within the MPA (R. Villablanca personal observations), highlighting the importance of considering connectivity with surrounding areas and the constraints of aquaculture sites for selecting location of MPAs (Micheli et al. 2004). The establishment of MPAs in Chile has so far been based on anecdotal recommendations (e.g., resource management, tourist attractions) rather than scientific criteria. This approach is considered inadequate to effectively protect biodiversity (Meir et al. 2004, Sutherland et al. 2004). Current strategies for implementing MPA networks require a systematic planning conservation method to identify optimal sites for protection of biodiversity (Sayre et al. 2000, Beck & Odaya 2001). The first step in planning an MPA is the assessment and mapping (Geographic Information System — GIS) of the coastal marine biodiversity, the physical environment and major threats (e.g., human uses) and then the identification of priority sites using Decision Support Systems (DSS) based on algorithms (Sala et al. 2002, Leslie et al. 2003). A DSS based on species richness has been used to identify priority areas for marine vertebrate conservation along the Chilean coast (Tognelli et al. 2005). Habitat classification is generally considered as the conservation goal in the DSS (Roberts et al. 2003). However, benthic surveys along the Chilean coast (C. Gaymer & C. Dumont unpublished data) revealed that communities are probably more appropriate to characterise the benthic environment and consider the ecosystem processes (e.g., trophic cascades; Shears & Babcock 2003), ecological interactions (e.g., predator-prey; Micheli et al. 2004) and population connectivity (e.g., larval dispersal; Palumbi 2003). Moreover, there is an urgent need for more scientific information in Chilean marine biodiversity (e.g., there is a lack of taxonomic expertise), population connectivity (e.g., identifying source and sink populations) and ecological processes (in particular species interactions in subtidal habitats are poorly studied). Although ecological knowledge is a key component in developing MPAs, the management effectiveness is the most important challenge for the success of an MPA (Mascia 2004, Pomeroy et al. 2004). A major difficulty arises from the way in which marine reserves and MUMPAs have been established in Chile. The former were created by an imposition from the central authority (fisheries ministry) without consulting the stakeholders, who are mostly in disagreement with this new status. This establishment strategy has turned enforcement into a complicated task for the fisheries authority, and this may turn into a major threat for the success of present and future MPAs. For example, since its creation in 1997, the marine reserve La Rinconada has been affected by frequent illegal extractions of scallops (M. Avendaño personal observations). Social conflicts due to lack of communication between the authority and the stakeholders are also present within the recently created marine reserves Isla Choros-Damas and Isla Chañaral. In contrast, a participative process took place in the establishment of the MUMPA Isla Grande de Atacama, incorporating most of the relevant actors (i.e., administrative authorities, stakeholders, managers, scientists and fishermen), offering the opportunity to evaluate contrasting interests in order to reduce potential conflicts. Social costs should be evaluated before the establishment of MPAs and a formal educational process should be implemented by the authorities to teach the importance of MPAs in developing sustainable exploitation of resources (Mascia 2004). The government should also negotiate compensations and propose alternative activities (e.g., tourism) to fishermen, who are the ancestral users of the MPA areas, and avoid creating high expectations (Mascia 2004, Sobel & Dahlgren 2004). Ideally, an international MPA network (from Ecuador to Chile) including the connectivity among MPAs should be implemented to effectively preserve biodiversity in the HCS. This should be achieved using the support of international tools and agreements, and international non-governmental organisations (NGOs) in order to co-ordinate and improve the quality of scientific information and reduce the costs (Balmford et al. 2004). Moreover, the Chilean government must contribute to funding for implementation and functioning of MPAs as successful conservation experiences from all over the

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world have demonstrated that self-funding (e.g., through tourism business) is not feasible (Balmford et al. 2004).

Outlook, long-term research vision and future research frontiers An outlook of the pressing scientific questions and tasks to be addressed in the short/mid-term future (during the next decade) within the HCS of northern and central Chile should, as a minimum, include (1) studies of ocean–atmosphere interactions and offshore oceanography, (2) research in inshore and offshore oxygen-minimum ecosystems, (3) research on inner inshore coastal oceanography and benthic-pelagic linkages, (4) development of an ecosystem-based adaptive resource management approach to fisheries that integrates socioeconomic aspects, (5) implementation of a coastal overarching network system for marine conservation-management, (6) novel approaches in coastal mariculture, (7) studies of marine non-indigenous species (NIS), (8) marine molecular biology, particularly on genomics, and (9) training of Chilean marine taxonomists. A long-term HCS vision (during the coming two decades) should also, as a minimum, include (1) intensification of precautionary and integrative ecosystem management; (2) implementation of a high-sea conservation policy; (3) intensification of research on the continental slope, deep-sea and abyssal ecosystems; (4) scientific and technological research on deep-sea gas (methane) hydrates; and (5) evaluations of the effects of future climate change.

Short/mid-term scientific outlook Research on ocean–atmosphere interactions and offshore oceanography The HCS offers unique opportunities for offshore oceanographic studies (e.g., Strub et al. 1998) and considered as a whole, Chile and Peru represent between 15% and 20% of the world’s fishery landings. Large-scale fluctuations in ocean climate (ENSO and the PDO) dominate interannual and interdecadal variability in the ocean, which in turn are linked to upwelling and climate changes (Chavez et al. 2003) and are considered key elements in the HCS functioning. Upwelling systems presently are experiencing ‘anomalous changes’ such as profound changes in the physical and biogeochemical properties in the California Current Systems (Freeland et al. 2003, Grantham et al. 2004), massive nitrogen loss in the Benguela upwelling system (Kuypers et al. 2005), and hydrogen sulphide eruptions in the Atlantic Ocean off southern Africa and linked abrupt degradation of upwelling systems (Bakun & Weeks 2004, Weeks et al. 2004, Arntz et al. 2006). Such interannual and decadal variability and anomalous changes may intensify due to global climate changes, which will also affect the HCS, causing important changes in productivity, biogeochemical cycling and fisheries. Research on these and related oceanographic topics is urgently needed for the HCS. Research on inshore and offshore oxygen-minimum ecosystems Oxygen-minimum zones (OMZs) in the ocean generally form along the EBCs. The decomposition of upwelling-derived biomass in combination with sluggish circulation of mid-water masses strongly enhance the hypoxia conditions, as is the case in the HCS (Levin & Gage 1998, Morales et al. 1999, Levin 2002, Helly & Levin 2004, Ulloa & De Pol 2004). The HCS is characterised by the relative shallowness of the oxygen-minimum layer. The OMZ produces peculiar environments with organisms highly resistant to low oxygen concentrations (Levin et al. 2001, Gallardo et al. 2004). These environments are unique in the HCS and quite different from those off California (Arntz et al. 2006). They offer diverse opportunities to develop frontier research, ranging from

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evolutionary adaptability, primary and secondary production, biodiversity, and species invasions (Castilla & Neill 2007) to impacts on fisheries. Research on nearshore coastal oceanography and benthic-pelagic coupling The nearshore oceanography (coastal border to about 2–3 km offshore) is one of the least-known areas of the world’s oceans. In this regard, during the past 5 yr several lines of research have been developed in central Chile (e.g., Poulin et al. 2002a,b, Kaplan et al. 2003, Wieters et al. 2003, Narváez et al. 2004, Vargas et al. 2004, Piñones et al. 2005) and further research efforts are needed. Unless improvements in the knowledge of nearshore oceanography are achieved, coastal ecology cannot be properly understood. This includes topics such as dispersal of nearshore propagules, benthic resource fisheries and sustainability issues, coastal conservation and pollution impacts. Linking nearshore oceanography and the coastal benthic-pelagic systems remains as a challenge, not only along the HCS (Castilla et al. 2002a, Escribano et al. 2002, Lagos et al. 2002, Marín & Moreno 2002), but also around the world (Shanks 1983, 1995, 2002, Largier 2002). Development of an integrative and adaptive resource management approach to fisheries Pelagic, benthic and demersal marine resource management in Chile is regulated by the Ley de Pesca y Acuicultura 1991 (Fishery and Aquaculture Law, FAL 1991), under the responsibility of the Subsecretary of Fisheries, Ministry of Economy. For the management and administration of fisheries, among other regulations, the law (1) defines the ‘artisanal fishery’ referring to vessels/boats 70% of natural habitats in the habitable portion of the planet (Hannah et al. 1994) and that we are still losing somewhere between 0.5% and 1.5% of wild nature each year (Balmford et al. 2003). Habitat loss is also well recognised as an important threat in the marine environment (Suchanek 1994, Gray 1997, Wolff 2000) but has not been as much a focus of science and conservation as in terrestrial environments. Habitat loss is particularly severe in coastal marine ecosystems, where human activities have been historically concentrated (Suchanek 1994, Lotze 2004, Knottnerus 2005, Lotze et al. 2006, Valiela 2006). Coastal zones occupy 60% of the world’s population (EEA 1999a). Globally, the number of people living within 100 km of the coast increased from roughly 2 billion in 1990 to 2.2 billion in 1995 (Burke et al. 2001), and the population living on the coast is projected to double in the next 30 yr with an expected 75% of the world’s population residing in coastal areas by 2025 (EEA 1999a). As human population has increased in coastal areas, so has the pressure on coastal ecosystems through habitat conversion, increased pollution, and demand for coastal resources. Coastal systems provide many important services to humans such as nutrient cycling, food production, provision of habitat/refugia, disturbance regulation, natural barriers to erosion, control of water quality, and nursery grounds. Indeed the global value of services from seagrasses, estuaries and coastal wetlands is estimated to be 10 times higher than that of any terrestrial ecosystems (Costanza et al. 1997). Recent reviews have examined the extent of habitat loss and fragmentation in tropical environments across large regions for coral reefs (e.g., Sebens 1994, Spalding et al. 2001, Pandolfi et al. 2003, Wilkinson 2004) and mangroves (e.g., Burke et al. 2001, Valiela et al. 2001, Alongi 2002, Wilkie & Fortuna 2003). These studies have done much to advance our understanding of the status and trends of tropical marine ecosystems at multinational and global levels and have been influential in galvanising support for tropical science, conservation and management. Our understanding of the status and trends of temperate marine habitats is surprisingly further behind. Few scientific institutions, organisations or agencies have programmes that focus on temperate marine environments beyond a regional level, and almost no non-governmental organisations (NGOs) or agencies have multinational or global programmes that focus particularly on temperate ecosystems such as seagrasses, salt marshes or oyster reefs or the issue of habitat loss. There have been a few broad reviews of the condition of key temperate habitats (e.g., Kennish 2002, Steneck et al. 2002, Thompson et al. 2002, Lotze et al. 2006) and some recent exemplary efforts to pull together global distribution data on seagrasses (Short & Wyllie-Echeverria 1996, Duarte 2002, Green & Short 2003). Nonetheless, huge gaps still remain in our knowledge of habitat loss on temperate coasts and estuaries. This gap is particularly disturbing because these coasts contain some of the most productive, diverse and, at the same time, degraded ecosystems on Earth (Suchanek 1994, Edgar et al. 2000). In Europe, there has been increasing awareness and concern about the degradation of natural habitats (e.g., Laffoley 2000). Many European coastal habitats have been lost or severely degraded, and it is estimated that only a small percentage of the European coastline ( 2.5 million Outside data coverage

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that European coastal areas will face increasing pressures from population growth (EEA 2005, 2006a). The coasts of the Mediterranean Sea, in particular, have always been among the most densely populated regions on Earth, with an estimated 5700–6600 people km−1 of coastline in 2000 (UNEP/MAP/PAP 2001). Along Mediterranean coasts, the population increased by 46% between 1980 (84.5 million) and 2000 (123.7 million), and it is projected to nearly double between 2000 and 2025 (UNEP/MAP/PAP 2001). Increased land use and development of settlements, agriculture, industries, ports, military installations, mines, power plants and other infrastructures has accompanied population growth in European coastal areas. Their development has posed and still poses severe threats to coastal areas (EEA 2006a). Estuarine and coastal landscapes have been deeply modified and transformed in a process that in some regions, such as the western Netherlands, dates as far back as late prehistoric periods, when the first attempts were made to control the flow of water through the construction of dams and sluices (Rippon 2000). During the Roman times, reclamation of coastal marshes became intensive in some regions (e.g., the Severn Estuary; Rippon 1997), and after the tenth to twelfth centuries, large-scale transformations and reclamations took place systematically around Europe (Wolff 1992, Cencini 1998, Rippon 2000, Reise 2005). In the Wadden Sea region, about 350


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15,000 km2 of wetland, lagoons, coastal lakes and tidal flats have been embanked, drained and converted into arable land and pasture over the centuries (Figure 4; see also Wolff 1992, 1997). In the United Kingdom land reclamation has affected at least 85% of the estuaries since Roman times, with losses of intertidal areas ranging between 25 and up to >80% (Davidson et al. 1991); such widespread claim of estuarine land is continuing at rates of 0.2–0.7% yr−1 and affects also estuaries of recognised international wildlife importance included in the Ramsar/Special Protection Area (SPA) network. Data from the CORINE project indicate that 22,000 km2 of the coastal zone in Europe are covered in concrete or asphalt (EEA 2005), and that artificial surfaces increased by almost 1900 km2 between 1990 and 2000 alone (EEA 2006a). The greatest urban developments occur along the Euro-Mediterranean coasts. At present about two thirds of the Mediterranean coastline is urbanised, with this fraction exceeding 75% in the regions with the most developed industries (UNEP/MAP/PAP 2001). More than 50% of the Mediterranean coasts are dominated by concrete with >1500 km of artificial coasts, of which about 1250 km are developed for harbours and ports (EEA 1999c). Growth of cities (particularly tourist developments) and development of industry in some regions (including the French Riviera, Athens, Barcelona, Marseille, Naples, the north Adriatic shorelines) have taken up to 90% of the coastline (Jeftic et al. 1990, Meinesz et al. 1991, Cencini 1998). In Italy, a survey carried out by World Wildlife Fund (WWF) showed that, in 1996, 42.6% of the entire Italian coast was subject to intensive human occupation (areas completely occupied by built-up centres and infrastructures), 13% had extensive occupation (free zones occupied only by extensive building and infrastructures) and only 29% was free from buildings and infrastructures (EEA 1999c). Coastal zone urbanisation will further increase in the near future, with projected increases of 10–20% for most Mediterranean countries (EEA 2006a). Severe decreases of water quality have generally followed population growth with organic pollution as a major driving factor (Jansson & Dahlberg 1999, Diaz 2001, van Beusekom 2005). Excessive nutrient enrichment has been a problem in European waters historically (Islam & Tanaka 2004). Hoffman (2005) reports that archaeological signs of eutrophication from dense, mainly urban populations were detected on the Bodensee shore at Konstanz (Germany) in late-mediaeval times, and that in 1415 a royal ordinance tried to mitigate the low water quality of the Seine below Paris. Nutrient loads started to rise probably around 1700–1800, increased significantly in the early 1900s and steeply accelerated after the 1950s (Lotze et al. 2006). It is estimated that in the Baltic and North Sea regions nitrogen (N) and phosphorus (P) loads from land and atmosphere have increased about 2–4 and 4–8 times, respectively, since the 1940s (Nehring 1992, EEA 2001, Karlson et al. 2002). Historical reconstructions of the preindustrial trophic status in the Wadden Sea suggest about 5-fold greater organic matter turnovers nowadays compared with preindustrial conditions (van Beusekom 2005). The historical development in nutrient loads to the Mediterranean and Black Seas is unknown, but is probably of the same magnitude (UNEP/FAO/WHO 1996, EEA 1999c). For example, in the north Adriatic Sea nutrient load has been increasing since at least 1900 and it markedly intensified after 1930 (Barmawidjaja et al. 1995, Sangiorgi & Donders 2004), with a doubling of nutrient loads in the Po river between 1968 and 1980 (Marchetti et al. 1989). In the Black Sea, concentrations of nitrate have increased 5 times and phosphate 20 times from the 1960s to 1980s (Gomoiu 1992). The increased eutrophication has, as a secondary effect, led to increased oxygen consumption on the sea bed and expansion of areas with hypoxia and anoxia (Diaz 2001, Karlson et al. 2002). In the Black Sea up to 90% of the waters are anoxic. The Kattegat has been affected by seasonal hypoxia since the beginning of the 1980s, which has followed a more than 3-fold increase in N input in the 1960s and 1970s (Rosenberg et al. 1990). Similarly, in the north Adriatic Sea the first signs of hypoxia started around 1960 and developed into severe anoxic events over the past 20 yr (Barmawidjaja et al. 1995, Diaz 2001). Since the middle of the 1980s the phosphorus load has 351


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generally levelled off or declined locally. In some areas such as the North Sea there have been declines in P up to 50% due to improved sewage treatment, reduced industrial discharges and a change to phosphorus-free detergents (Frid et al. 2003). However, there do not yet seem to be discernible European-scale reduction of nitrogen inputs, marine eutrophication or extent of anoxic areas (Karlson et al. 2002). Increased loads of sediments have followed changes in land use both inland and along the coasts of Europe, but long-term data on water turbidity and sediment load are limited even at local scales (Lumb 1990). The greatest impacts were felt when forests were extensively cleared for timber, agriculture or urban developments, which together with interferences in the natural course of rivers caused dramatic acceleration of natural soil erosion (Airoldi 2003). In Europe clearing of forested catchments for agriculture commenced several thousand years ago (e.g., see the historical reconstruction by Cencini 1998). Episodes of accelerated erosion following phases of expansion of arable lands were common during mediaeval times (Hoffmann 2005) and became particularly severe during the nineteenth century (Pukaric & Jorissen 1990, Barmawidjaja et al. 1995). Chemical pollution has also affected European estuaries and coastal waters since ancient times (Islam & Tanaka 2004), particularly in the Mediterranean Sea, where overall 101 pollution ‘hot spots’ have been identified, generally located in semi-enclosed gulfs and bays near important harbours, big cities and industrial areas (UNEP/MAP/PAP 2001, EEA 2006b). Pollution from shipping, oil spill traffic, drilling activities and related accidents is particularly severe in Europe (EEA 1998, 1999c, 2006a, Thompson et al. 2002). Some of the busiest shipping lanes in the world are found in the Baltic Sea, North Sea and Mediterranean Sea (Frid et al. 2003), and about 22% of the total world petroleum traffic passes through the Mediterranean Sea (Jeftic et al. 1990). Marine pollution has become a major concern in Europe, and many E.U., trans-national and national initiatives (see next section, p. 356) have helped to control the disposal of urban and industrial pollutants in coastal areas. Even so, there are still large pollution loads, and long-term contamination of sediments is a major problem. Marine food resources have been used by Europeans since prehistory. At some heavily populated localities, particularly along the Mediterranean shores, the most valued species had severely declined in abundance and size by the end of the Roman era (Hoffmann 2005) with detectable effects on coastal systems (Sala 2004). Exploitation increased during late-mediaeval times, when fisheries became subject to market exploitation, and in subsequent centuries growing food demand and technological progress led to almost unrestricted overexploitation of coastal resources (Hoffmann 2005, Wolff 2005, Lotze et al. 2006). The total fish landings in European sea regions peaked at 12 million t in 1997, but have decreased since in both quantity and quality, down to 7.6 million t in 2002 (EEA 2006a). Disruptive fishing techniques are considered among the major causes of physical destruction of marine coastal habitats at global scales (Watling & Norse 1998, Turner et al. 1999, Thrush & Dayton 2002). In Europe, bottom trawls, bivalve dredging, pneumatic hammering of date mussels, explosives and other disruptive fishing techniques have a long history of use, mainly in estuaries, bays and continental shelf waters (Fanelli et al. 1994, Bavestrello et al. 1997, Lindeboom & de Groot 1998, Cicogna et al. 1999, EEA 1999c, Johnson 2002, Hall-Spencer et al. 2003, Tudela 2004). In Britain, concern about the adverse effects of fishing on marine habitats and wildlife populations dates back to the fourteenth century when it was noted in a petition presented to Parliament in the year 1376–1377 (quoted in Hore & Jex 1880) that “the hard and long iron of the said ‘wondyrchoun’, [an oyster dredge] … destroys the spawn and brood of the fish beneath the said water, and also destroys the spat of oysters, muscles [sic], and other fish by which large fish are accustomed to live and be supported”. The use of trawls and other mobile fishing gears accelerated sharply with the introduction of diesel engines in the 1920s (Watling & Norse 1998). The sea bed in Europe

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has been trawled to a depth of over 1000 m since the 1970s, affecting extensive areas of benthic habitats. Aquaculture, which can have some benefits, has had increasingly adverse effects on coastal habitats. World aquaculture production has increased by >300% since 1984, with growth of about 10% a year in the 1990s, making it the fastest-growing food production activity (Mock et al. 1998). In Europe the culture of fish, shrimp, shellfish and seaweeds has been used as an alternative source of marine food at least since Roman times (Hoffmann 2005), and in regions such as the Po delta area salt marshes were transformed centuries ago into artificial fishing lagoons (Cencini 1998). Unprecedented growth in production has occurred in the last decades (Váradi 2001, EEA 2006a), with significant impacts on bottom habitats and assemblages (e.g., Holmer et al. 2001, EEA 2006b). In 1998, total marine aquaculture production in Europe was >1.3 million t, with most production concentrated in Norway, France, Spain and Italy (Váradi 2001). In the Mediterranean region, marine aquaculture production has increased from 19,997 t in 1970 to 339,185 t in 2002 (EEA 2006b), and the total production of salmon in fish farms (mainly in Norway and Scotland) has increased from 70,000 t in 1990 to 148,000 t in 1996 (EEA 2002) up to 540,000 t in Norway alone in 2003 (EEA 2006a). More recent pressure and threats to European coastlines are from tourism and development of recreational infrastructures, particularly in the Mediterranean region. Before the 1930s, tourism was a relatively minor phenomenon, although it did lead to the beginning of urbanisation in seaside areas (EEA 1999c). From the 1930s onward and especially after World War II, mass tourism started to grow, and the phenomenon was amplified by the development of transport facilities (e.g., Cencini 1998). Nowadays, the Mediterranean coast is the world’s leading holiday destination, accounting for 30% of the world’s tourism, and in some countries coastal tourism represents up to 90% of all tourism. In 1990 alone, 135 million vacationers flocked to the Mediterranean coast, and by 2025 the annual crowd is projected to increase to 235–350 million tourists (EEA 1999c). Effects on coastal habitats have been devastating. In Spain, tourist developments occupy 42% of the entire coast (Jeftic et al. 1990), with peaks in areas such as the Catalonia coast, where tourist developments make up 337 km of the total 580 km. Similarly, buildings, roads, bathing establishments and other recreational facilities located directly over the beaches and sand dunes almost entirely occupy the Italian coast of the north Adriatic Sea (Cencini 1998). The demand for marinas and yacht harbours has been growing all over the Mediterranean coasts, with an estimated growth for France of 1.5–2.6% yr−1 (EEA 2006a). Increased coastal erosion and flooding (often indirectly related to human activities) also pose serious threats to European coastlines (EC 2004). A recent inventory of coastal evolution in Europe undertaken within the CORINE programme showed 55% of the coastline to be stable, 19% to be suffering from erosion problems and 8% to be depositional (Stanners & Bourdeau 1995). Some coastal regions are also gradually subsiding (Bondesan et al. 1995, EEA 2006a), with subsidence sometimes enhanced by groundwater or petrochemical extraction (Bird 1993), while land lift up to 9 mm yr−1 is occurring in areas of the Baltic Sea (HELCOM 1998). Erosion mitigation schemes have been put in place to respond to the problem of coastal erosion, which affected about 7600 km of coasts in 2001 (EC 2004). Defence measures include a variety of hard defence structures (e.g., breakwaters, groynes, seawalls, dykes or other rock-armoured structures), which have proliferated in the second half of the twentieth century, leading to severe hardening of coastal areas and changes in sediment structure (Airoldi et al. 2005). In the north Adriatic Sea, >190 km of artificial structures, mainly groynes and breakwaters, seawalls and jetties (Figure 5), have been built along 300 km of naturally low sedimentary shores (Bondesan et al. 1995, Cencini 1998). This hardening has caused severe losses and alterations of shallow sedimentary habitats (e.g., Martin et al. 2005) and has introduced new artificial habitats, with dramatic effects on native habitats and assemblages (Bacchiocchi & Airoldi 2003, Bulleri & Airoldi 2005). Similar 354



Figure 5 Coastal defence structures along the highly urbanised Italian shores of the north Adriatic Sea. (Photo by Giorgio Benelli. With permission.)

examples occur in many other European coasts (e.g., Davidson et al. 1991, Anthony 1994, Reise 2005), presumably affecting an overlooked enormous amount of benthic habitats. Overall >15,000 km of coasts in Europe are now actively retreating, some of them in spite of coastal protection works (2900 km), and another 4700 km are artificially stabilised (EC 2004). Globally the problem of erosion and flooding is becoming much more serious because of rising sea levels and an increased storm frequency as a result of global climate change (Bray & Hooke 1997, Valiela 2006). During the past century, the mean global sea level has risen between 10 and 25 cm (Burke et al. 2001). The Intergovernmental Panel on Climate Change (IPCC) Working Group 355



has projected a global sea-level rise of 15–95 cm by the year 2100. The recession of coastlines is expected to continue even in the absence of new human activities (Bondesan et al. 1995). Approximately 450–600 non-indigenous marine species have been added to European coastal fauna and flora, often facilitated by human-mediated processes such as shipping, aquaculture and aquaria (Reise et al. 2006 and references therein). Some introductions occurred hundreds of years ago, as is the case of the sand-gaper, Mya arenaria, which was probably transported as food from North America to Europe by the Vikings (Petersen et al. 1992). Rates of introduction rose dramatically in the past century, probably in relation to increased shipping and aquaculture. In the Mediterranean Sea, the number of introduced species has nearly doubled every 20 yr since the beginning of the twentieth century (Boudouresque & Verlaque 2002). The number of introduced species is often greatest in estuaries, lagoons, embayments, closed seas, canals and harbours, probably because of low species richness combined with strong anthropogenic change (OcchipintiAmbrogi 2001, Reise et al. 2006). They have profoundly altered European coastal ecosystems and are displacing some native species. One notorious example is the invasion of Caulerpa taxifolia along the coastlines of Spain, France, Monaco, Italy, Croatia and Tunisia (Meinesz et al. 2001) although there is debate about the area of benthos affected (Jaubert et al. 2003). Nevertheless, these invasions do not seem to have caused large-scale extinctions in recipient biota and losses of native coastal habitats (Wolff 2000).

E.U. coastal policies and trans-national agreements The European Union has been involved in efforts to protect the European natural heritage for the past 30 yr (Table 1). At the international level, the European Union has signed a number of important conventions aimed at nature protection, including the Ramsar Convention on the Conservation of Wetlands, the Bonn Convention on Migratory Species, and the Rio Convention on Biological Diversity (CBD), and shares the international commitment of the World Summit for Sustainable Development to establish a globally representative system of marine and coastal protected areas by 2012 (Kelleher et al. 1995, Green & Paine 1997). At the European level, the Bern Convention has led the development of policy and action in nature conservation in Europe. It lists protected species, including a number of marine plants and invertebrates, and requires its parties to prevent the disappearance of endangered natural habitats. The Sixth Environmental Action Plan, setting the European Union’s environmental policy agenda until 2012, highlighted nature and biodiversity as a top priority, and the E.U. leaders in Gothenburg in 2001 launched the E.U. Sustainable Development Strategy to halt the loss of biodiversity in the European Union by 2010. The European Union has adopted a Biodiversity Strategy and Action Plans (currently under review), and Member States have developed — or are developing — their own national strategies and action plans (e.g., U.K. Biodiversity Group 1999). Other policies, in particular the Birds Directive and the Habitats Directive, have been promoted to rectify or reduce damage to European natural habitats and associated species. Following the criteria set out in the directives, each Member State must draw up a list of sites hosting wild fauna and natural habitats and put in place a special management plan to protect them, combining longterm preservation with economic and social activities, as part of a sustainable development strategy. The final aims are the creation of a European Ecological Network of Special Protection Areas (SPAs) and Special Areas of Conservation (SACs), called NATURA 2000, and the integration of nature protection into other E.U. policy areas, such as agriculture, fishery, industry, regional development and transport. Indirect protection to a variety of habitats also comes from a number of E.U. Directives that regulate water quality, including the Dangerous Substances, Shellfish Waters, Integrated Pollution

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Table 1 Summary of main protection initiatives adopted by the European Union and State Members that directly or indirectly address issues related to the protection of European marine coastal habitats and associated assemblages Initiative

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Web site

Ramsar Convention

Ramsar Convention on Wetlands, Ramsar (1971). Provides the framework for the conservation and wise use of wetlands of international importance especially as waterfowl habitat. Includes salt marshes and some lagoon systems and marine waters to a depth of 6 m. Convention on the Conservation of Migratory Species of Wild Animals, Bonn (1979). Intergovernmental treaty, aiming to conserve terrestrial, marine and avian migratory species throughout their range. Convention on Biological Diversity, Rio de Janeiro (1992). Provides legal framework for biodiversity conservation and sustainable development. The Jakarta Mandate (1995) leads activity in marine biodiversity management and conservation. Convention on the Conservation of European Wildlife and Natural Habitats, Bern (1979). Aims at preserving wild flora and fauna (including some marine species) and their natural habitats through national programmes using the co-operation between European States. Convention for the International Council for the Exploration of the Sea, Copenhagen (1964). Coordinates and promotes marine research in the North Atlantic, including the Baltic Sea and North Sea, and the Common Fisheries Policy on the protection of the marine environment and the regulation of fisheries. Convention for the Protection of the Marine Environment of the northeast Atlantic, Paris (1992). Merged the 1972 Oslo Convention on dumping waste at sea and the 1974 Paris Convention on land-based sources of marine pollution. It guides the protection of the marine environment of the northeast Atlantic and the identification of priority habitats and species. Six declarations produced at as many International Conferences on the Protection of the North Sea (first in Bremen, 1984). Political commitments to the protection of the North Sea environment, addressing, e.g., species and habitats issues, pollution and fisheries. Convention on the Protection of the Marine Environment of the Baltic Sea Area, Helsinki (1992). Guides the protection of the marine environment of the Baltic Sea from pollution and the identification of priority habitats and species for protection. Joint Declaration of The Netherlands, Denmark and Germany, Copenhagen (1982). Aimed at co-ordinating the protection of the Wadden Sea National Park. In 1997, a Trilateral Wadden Sea Plan was adopted.

www.ramsar.org

Bonn Convention

Rio Convention

Bern Convention

ICES Convention

OSPAR Convention

North Sea Conference Declarations

Helsinki Convention

Trilateral Wadden Sea Cooperation

www.cms.int

www.biodiv.org/convention/default.shtml

www.coe.int/T/E/Cultural_Co-operation/ Environment/Nature_and_biological_ diversity/Nature_protection/

www.ices.dk/aboutus/convention.asp

www.ospar.org

www.sweden.gov.se/sb/d/6363/a/57475; jsessionid=abPgelqjfxJ8

www.helcom.fi

www.waddensea-secretariat.org

(continued on next page)

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Table 1 (continued) Summary of main protection initiatives adopted by the European Union and State Members that directly or indirectly address issues related to the protection of European marine coastal habitats and associated assemblages Initiative

Description

Web site

Barcelona Convention/ Mediterranean Action Plan (MAP)

Amended in 1995 as the Convention for the Protection of the Marine Environment and the Coastal Region of the Mediterranean, Barcelona (1976). Provides legal framework to MAP (1975), under UNEP Regional Seas Programme. Aims to control human impacts (e.g., marine pollution, tourism) and protect the marine and coastal Mediterranean environments. Convention on the Protection of the Black Sea against Pollution, Bucharest (1992). Aims to control and prevent pollution and preserve biodiversity in the Black Sea. Council Directive on the Conservation of Wild Birds. Identifies 194 endangered species and subspecies of birds for which the E.U. Member States are required to designate Special Protection Areas (SPAs). Over 4000 SPAs have been designated to date, covering 8% of E.U. territory. Council Directive on the Conservation of Natural Habitats and of Wild Fauna and Flora. Aims to protect wildlife species and habitats which have conservation that requires the designation of Special Areas of Conservation (SACs). These sites, together with the SPAs of the Birds Directive, make up the NATURA 2000 network, currently covering about 15% of E.U. coasts. Marine habitats broadly defined, and few marine species listed. Council Directive on the Quality Required of Shellfish Waters. Aims to ensure a suitable environment for shellfish harvest. Member States are required to designate coastal and brackish waters that need improvement to support shellfish fisheries. Integrates and updates existing E.U. water legislations (e.g., Discharges of Dangerous Substances, Urban Waste Water Treatment, Nitrates Directive) and provides for water management. Complemented by the recently revised Bathing Water Directive (2006/7/EC). The proposed directive aims to define common objectives and principles at E.U. level to achieve good environmental status of the European marine environments by 2021. It will establish European Marine Regions as management units for implementation.

www.unepmap.org

Bucharest Convention

Birds Directive (79/409/EEC)

Habitats Directive (92/43/EEC)

Shellfish Waters Directive (79/923/EEC)

Water Framework Directive (2000/60/EC) Marine Strategy Directive

Note: All web sites last accessed 8 August 2006.

358

www.blacksea-commission.org

www.ec.europa.eu/environment/nature/ nature_conservation/eu_nature_legislation/ birds_directive/index_en.htm

www.ec.europa.eu/environment/nature/ nature_conservation/eu_nature_legislation/ habitats_directive/index_en.htm

www.europa.eu.int/eur-lex/en/consleg/pdf/ 1979/en_1979L0923_do_001.pdf

www.europa.eu.int/comm/environment/ water

www.ec.europa.eu/environment/water/ marine.htm



Control, Urban Waste Waters, Bathing Waters and the Water Framework (Table 1). These commitments provide for the regulation of discharges to the sea and have set targets and quality standards covering many metals, pesticides and other toxic substances. In addition to E.U. initiatives, a number of trans-national agreements have been developed to address some conservation issues within the main European seas, including the OSPAR Convention, the North Sea Conference Declarations, the Helsinki Convention, the Trilateral Cooperation on the Protection of the Wadden Sea, the Mediterranean Action Plan and the Black Sea Environmental Programme (Table 1). These programmes generally address water quality and fishery concerns and are not specifically focused on habitat loss and protection, although initiatives have also included commitments toward establishing an integrated network of Marine Protected Areas (MPAs). A more focused initiative for the Atlantic Ocean and Baltic Sea is the commitment of the Joint Ministerial Meeting of the Helsinki and OSPAR Commissions to complete by 2010 a joint network of MPAs that, together with the NATURA 2000 network, would be ecologically coherent. In recent years there has been increasing awareness that past efforts to protect European marine coastal habitats and associated species have been marginal relative to terrestrial environments and that there is limited co-ordination of national and transnational initiatives at a European level. The global MPA database indicates that there are 1129 MPAs in Europe (Figure 6) covering 236,000 km2 (MPA Global 2006). Most of these MPAs are small, and they often lack adequate political and

N

0

250 500

1,000 Km

Figure 6 Marine Protected Areas in Europe (UNEP/WCMC 2006; data extracted August 2006). These MPAs include, for example, nature reserves, national parks, habitat/species management areas, RAMSAR Wetlands of International Importance and World Heritage Sites. The map should be considered indicative of general distribution not areal extent of MPAs.

359



financial support and effective enforcement (an extensive review of MPAs in Europe is offered in Kelleher et al. 1995, and specifically for the Mediterranean Sea in Badalamenti et al. 2000). The application of the Birds and Habitats Directives to the marine environment is also presenting major challenges, with significant delays in the selection and designation of the marine sites of the NATURA 2000 network. In response the European Union has set the protection of the marine environment as a major priority and has launched the Marine Strategy Directive, specifically aimed at protecting and conserving marine ecosystems and promoting the sustainable use of marine resources through the development of an integrated, coherent policy for the marine environment. A 2004 conference in Malahide (Ireland) has also set the necessary steps to complete the selection of the marine NATURA 2000 sites by the end of 2008.

Coastal habitats in Europe Despite its relatively small geographic size, Europe has a very long coastline, approximating 325,892 km, including islands (Pruett & Cimino 2000). It comprises the main marine regions of the northeast Atlantic, part of the Arctic, the Baltic Sea, the North Sea, the Mediterranean Sea and the Black Sea (Frid et al. 2003, EEA 2006a). It includes primarily temperate environments as well as some Arctic and subtropical climate environments and covers a variety of geomorphological features (EEA 2002). There are a large variety of extremely productive, diverse and valuable natural coastal habitats, including sea-bed communities of macroalgae and seagrasses, tidal mudflats, salt marshes and biogenic reefs (Stanners & Bourdeau 1995). However, there does not seem to be any comprehensive summary of the current distribution and status of native habitats along European coastlines as a whole. Databases are available or are being prepared for some habitats (e.g., wetlands (Nivet & Frazier 2004) and seagrasses (Green & Short 2003)), countries (e.g., the United Kingdom; Hiscock & Tyler-Walters 2006), and regions (e.g., northwest Europe, the OSPAR regions and the Baltic Sea; ICES 2006a). However, for most European coasts information, if any, is scattered, fragmented and not easily accessible. Even less information is available about long-term trends of habitat loss or degradation. With some notable exceptions (e.g., the Wadden Sea; Lotze & Reise 2005), there is little comprehensive historical information on coastal habitats prior to about 1900 (e.g., Hiscock & Kimmance 2003). In this section published information, from both the scientific and grey (reports, web sites) literature is critically documented for major coastal habitats. The aim was not to reconstruct trends from local historical sources, maps or databases but to cover the most pertinent literature that reported data and information at regional, national, trans-national or European levels. The data summarised in this section are of variable quality and it should not be inferred that they provide a complete picture of the status of European coastal habitats.

Coastal wetlands and salt marshes Current distribution and status Much of the European coastline consists of a chain of extensive estuaries, lagoons and intertidal bays interspersed through stretches of rocky shore and sandy beaches. These coastal wetlands represent some of the most productive and biologically diverse components of near-shore ecosystems (Dugan 1993, Keddy 2000). They provide nursery grounds for commercially important fishes, habitat for shellfish, birds and a variety of biota and play a fundamental role in flood control, nutrient cycling and sediment dynamics. These coastal wetlands are patchworks of sand, mud flats

360



and salt marshes. Salt marshes, with their vegetated complex surfaces, form one of the most important components of these wetlands (Adam 1990). Coastal salt marshes are distinctive habitats that can be delineated relatively easily on remotely sensed images (e.g., Ekebom & Erkkilä 2003). In Europe, they have been the target of studies for a long time (see, among others, Dijkema 1984, Allen & Pye 1992, Jones & Hughes 1993, Rippon 2000, Adam 2002). However, quantitative information on their distribution and status is limited even at local scales. Information is often available for the more general category of ‘coastal wetlands’ and here the focus is mainly on this broader habitat type. Accounts exist at regional and local scales but most European countries lack comprehensive inventories of the extent and status of coastal wetlands, with numerous countries lacking almost any organised information on these conspicuous habitats (Table 2). One of the greatest difficulties in completing such inventories is in identifying intact wetlands, that is those wetlands that are not so severely transformed or deteriorated as to be functionally extinct (Allen 2000, Nivet & Frazier 2004). In the Severn Estuary, England, for example, there are a total of about 14 km2 of intact marshes and about 840 km2 of enclosed marshes, and this ratio of intact to enclosed coastal marshes may be common across Ireland, France, the Netherlands, northwest Germany and Denmark (Allen 2000). Other difficulties in estimating coverage include incompatible information (e.g., inconsistent classifications and methodologies) and the lack of co-ordination between different studies or for different wetland types (Nivet & Frazier 2004). A European-scale review of the current distribution and coverage of coastal wetlands has been recently completed by Nivet & Frazier (2004) that integrates and updates previous inventories by Jones & Hughes (1993) and Finlayson & Spiers (1999). According to this inventory, the total cover of marine/coastal wetlands in Europe is around 51,910 km2, and detailed information for individual countries is summarised in Table 2. An inventory of the distribution of European wetlands, including coastal wetlands, is also available in the CORINE database (EEA 1999b). A broad map of the distribution of salt marshes in Europe is given in Figure 7. There is little comprehensive information on the status of coastal wetlands and salt marshes in Europe but there are clear indications that the historical concentration of human activities in European coastal wetlands has deeply modified their structure and function (Dijkema 1984, U.K. Biodiversity Group 1999, Allen & Pye 1992). Adam (2002) points out that nowadays minerogenic sedimentation prevails over autogenic (organic matter) sedimentation in the majority of European marshes. Most wetlands have deeply altered flow regimes (e.g., Cencini 1998, Reise 2005) with associated important effects on sediment dynamics as well as nutrient and salinity regimes (e.g., Allen 2003) and often are heavily polluted (e.g., Trombini et al. 2003). Their vegetation composition is the product of centuries of use and management (e.g., Wolff 2000) and their fauna and flora have been deeply transformed by introduced species (Reise et al. 2006). Historical losses and causes Coastal wetlands have suffered some of the most serious habitat loss rates (Dugan 1993, Suchanek 1994, Rippon 2000, Valiela 2006) and some estimates suggest that over time temperate estuaries and coastal areas may have lost approximately 67% of the wetlands that existed (Lotze et al. 2006). Even when wetlands have not been completely lost, significant degradation of their environment has often occurred, impairing their functions (Dugan 1993, Wolff 2000). Exploitation of coastal wetlands and salt marshes in Europe dates back to at least the Neolithic, when salt marshes were used for salt production (e.g., Rippon 2000). Since then, these habitats have been increasingly exploited, providing location for settlement, agriculture and harbours; source of food, water and raw materials; and a focus for transport, trade and exploration (Rippon 1997,

361

362

n.a. n.a. n.a. 60% ^ since 1870 n.a 50% ^ 86% in twentieth century

8.3 n.a.

n.a. n.a. n.a 8851* n.a. 501 3 813

n.a. 6809*

1011

n.a. n.a.

n.a. 550–600 km since mid-1900s n.a. n.a.

150

2

Albania

Loss

Cover (km2)

Country

D, S, U, A

D

WQ, D

D, U

2500

>250 salt marshes are located/mapped; 65 km2 of estuarine area reclaimed in the Shannon estuary and 20 km2 in the Wexford Harbour mainly in nineteenth century 60% of the estimated losses occurred in 1938–1984; on the Po river delta, >70% of salt marshes reclaimed in 1870s–1960s; in Friuli Venezia Giulia 631 km2 lost in 1877–1990

In the Wadden Sea, 200 km2 of salt marshes lost during 1950–1984, and only 400 km2 remain today; in the whole Wadden Sea (not only Germany) 14,650 km2 of coastal wetlands lost since the eleventh century (33% of salt marshes in 1930–1987)

Losses of 22.8% ^ between 1950 and 1985 In Bretagne, 40% of coastal wetlands lost since 1960, and 66% of the remaining seriously affected by drainage; ongoing losses in the Languedoc Roussillon wetlands

D D, U

Limited settlement during Iron Age

Regional data/additional information

Country’s coastline shortened by 1168 km (14%)

2000

2000

Exploitation history (yr)

D

D

Cause of loss

Stanners & Bourdeau 1995, Cencini 1998

Curtis & Skeffington 1998, Healy & Hickey 2002

EEA 2006a

Dugan 1993, Rippon 2000, Reise 2005

EEA 2006a

Rippon 2000

Additional references

Table 2 Estimates of actual cover and historical losses of coastal wetlands (and when possible salt marshes) for European countries (and eventual additional regional information), main attributed causes of loss and known history of exploitation



363

n.a. n.a. n.a. 4043 (85 of salt marshes)

n.a. n.a. 795 5297 5786* n.a. 1041 6000

n.a. n.a. 5966 (450 of salt marshes)

Malta Monaco Montenegro Netherlands

Norway Poland Portugal Romania Russiaa Slovenia Spain Sweden

Turkey Ukraine United Kingdom

n.a. 70% ^ during the past 30 yr n.a. n.a. n.a. 4000 km2 of salt marshes between twelfth century and second half twentieth century n.a. 89.8% ^ n.a. 4000–6600 km2 ^ n.a. n.a. 60% 23% ^ since early 1800s n.a. n.a. >50% of salt marshes since Roman times; 38 lagoons in the latter half of 1980s; >913 km2 of estuary area D, U, E

Coastal/marine wetlands comprise (in km2) 14 sand dune slack, 2658 estuarine waters, 2793 intertidal flats, 450 salt marsh and 50 saline lagoons; salt marsh loss most significant in the 1800s but still ongoing (100 ha yr−1).

Most of the losses occurred in 1965–1990 >50% degraded since early 1800s; in some areas losses up to 80–90%

D, U D

Humans affected salt marshes since at least the Iron Age; earliest dam dates 175 B.C.

70% of salt marshes in the western Algarve lost before 1988

2500

>2500

D D, S, E

D, WQ, U

D

Rippon 1997, 2000, Davidson et al. 1991, Boorman 2003

EEA 2006a

Allen 2003

Wolff 1992, 1997, 2000, Rippon 2000

a

Including Asian Russia.

Note: If not otherwise specified in the reference column, information is derived from Nivet & Frazier (2004) and references therein, while information about history of exploitation is generally derived from Rippon (2000). A = aquaculture, D = drainage/embankment/land claim/conversion (e.g., to agriculture), E = erosion/sea-level rise/coastal squeeze, n.a. = not available, S = altered sediment loads (e.g., from inland deforestation), U = urban/harbour developments, WQ = water quality degradation. Note that the definition of coastal wetland is vague for most countries, and in many cases (indicated as *) only important or large marine wetlands were included. Concerning habitat loss, in many cases (indicated as ^) estimates refer to total wetlands because no distinction was made between coastal and inland wetlands, and often no time span is indicated. Overall, estimates should be considered as broad indications.

1426 413

Latvia Lithuania





Major complexes Isolated sites >500 ha Small 100 ha). South of Arcachon there is the largest Zostera nolti bed in Europe (70 km2 in 1984) and a large bed of Z. marina (4 km2 in 1984). The Glenan Archipelago lost 6 km2 of seagrasses during 1932–1992. In the Mediterranean many local documented losses of P. oceanica (e.g., close to Marseille, 5–6% per decade between 1900 and 1970, nowadays 90% is deteriorated; along the French Riviera, 30.57 km2 lost since 1800; seagrasses virtually extinct from the region of Toulon). 3184 ha affected by the invasion of C. taxifolia.

Seagrasses presumably not affected by the wasting disease.

Considered as widespread; severe declines of Posidonia oceanica in Istria between 1938–1998. Local invasions by Caulerpa taxifolia. Considered as widespread. Total cover is the sum of 30 km2 in the Wadden Sea and 1350–1680 km2 on the eastern coasts; vertical depth distribution reduced by 50% between 1900 and 1996–1997 (from 5–6 to 2–3 m in estuaries and from 7–8 to 4–5 m in open waters); 261 km2 lost in Limfjiorden between 1900 and 1994 and 559 km2 lost in Oresund between 1900 and 1996–2000; local losses continue nowadays (e.g., in 1994 eelgrass temporarily lost at Funen Island due to summer anoxia).


(continued on next page)

Meinesz et al. 1991, 2001, Glemarèc et al. 1997, Duarte 2002, EEA 2002

Reise 1994, Rask et al. 1999

Zavodnik & Jaklin 1990, Meinesz et al. 2001


Table 3 Estimates of actual cover and historical losses of seagrasses for European countries (and eventual additional regional information) and main attributed causes of loss



370

n.a.

n.a.

n.a.

Norway

Poland

Portugal

n.a. n.a. n.a. n.a.

Greece Iceland Ireland Italy

n.a. Extinct? n.a. n.a. n.a. 200

n.a. 170

Georgia Germany

Latvia Lithuania Malta Monaco Montenegro Netherlands

Cover (km2)

Country

n.a.

No losses since 1930 n.a.

n.a. 100%? n.a. n.a. n.a. n.a. Virtually extinct in Wadden Sea

n.a. n.a. n.a. 30–40% of P. oceanica in recent decades

n.a. n.a.

Loss

Most seagrasses are in southwest estuaries; in the Grevelingen estuary, the construction of two dams facilitated the growth of Z marina (from 12 km2 in 1964 to 34 km2 in 1985) and the disappearance of Z. noltii, which before was the most common seagrass (large-scale, unexplained die-off of Z. marina since 1986–1987). In the Wadden Sea 145 km2 lost in 1919–1971 and 3 km2 since) and only 1–2 km2 are left nowadays.

WD, Ec, S, EU, Fd

Virtually extinct in Puck Lagoon in 1957–1987, with subsequent recolonisation in some areas.

Severe invasion by C. taxifolia.

I

MGc, EU

Considered abundant over thousands of hectares in the past.

Most beds are mapped, and coverage of many is known (estimated 2350 km2 in Liguria, Lazio, Sardinia, Veneto and Friuli). Estimate of loss is derived from estimates from France. P. oceanica virtually extinct in the north Adriatic Sea, in some areas since earlier than 1850s. Z. marina disappeared from areas of the Venice lagoon. 9414 ha affected by the invasion of C. taxifolia.

S, EU, P, I, T

MGc

Most seagrasses occur in Wadden Sea; only present on eastern coasts. Baltic Sea: in Kiel Bight decreased from 6 to 2 m in depth between 1960 and end of 1980. No losses in Greisvald Lagoon between 1930 and 1988. North Sea: Z. marina extinct at Helgoland island as early as 1928. Considered as widespread.


WD, EU, S, MGb, LC

Cause of loss

Reise 1994, Short & Wyllie-Echeverria 1996, Wolff 2000, 2005

Meinesz et al. 2001

Barmawidjaja et al. 1995, Piazzi et al. 2000a,b, Guidetti 2001, Meinesz et al. 2001, Milazzo et al. 2004

Reise 1994, Bartsch & Tittley 2004


Table 3 (continued) Estimates of actual cover and historical losses of seagrasses for European countries (and eventual additional regional information) and main attributed causes of loss



371

n.a.

n.a. n.a. n.a.

Sweden

Turkey Ukraine United Kingdom n.a. n.a. n.a.

n.a.

n.a. n.a. n.a. 30–40% of P. oceanica in recent decades

WD

EU, MG

S, F, I, A, T Estimate of loss is derived from estimates from France. In the Mediterranean, 57% of P. oceanica beds were under regression in 1996 along 1000 km of coasts, and some are now extinct, with losses concentrated over 400 km of coasts. 58% of P. oceanica beds and 52% of seagrasses are degraded along the Catalan coasts and near Alicante, respectively. Local invasions by C. taxifolia. At least 60–130 km2 along the southern Baltic coasts. Losses of 58% (10.61 km2) of seagrasses in the Skagerrak in 1980s–2000. Considered as widespread. 950 km2 in the northwestern bays. Most beds mapped, and coverage of many known. About 140 sites of Z. marina and about 70 sites of Z. noltii, covering from 12 km2 (Cromarty Firth) to 20–40 hae. The Maplin Sands hosts one of the largest surviving population of Z. noltii in Europe (325 ha). Before 1900s, seagrasses were common; they severely declined in 1920s–1930s and have not recovered yet, particularly in southern and eastern England.

Davison & Hughes 1998, U.K. Biodiversity Group 1999

Baden et al. 2003

Marbà et al 1996, Meinesz et al. 2001, Duarte 2002, EEA 2002

b

Decrease of maximum Secchi depth from 12 m in 1900 to 6 m in the 1990s. Seagrasses replaced by filamentous algae. This may not necessarily be the cause of seagrass loss. c Including the closure of the Zuidersee in 1932. d Eelgrass harvesting until 1930, shell fisheries after 1970. e Other important sites are the Exe Estuary, the Solents marshes and the Isles of Scilly, Morfa Nefyn, Milford Haven, the Moray Firth, Carlingford Lough, Dundrum Bay, Strangford Lough and Lough Foyle.

a

Note: If not otherwise specified in the reference column, information is derived from the World Atlas of Seagrasses (Green & Short 2003) and references therein. A = aquaculture, E = engineering works and embankments, EU = eutrophication, F = fisheries, I = invasive species, MG = growth of ephemeral macroalgae (often a consequence of eutrophication), LC = land claim/waterfront development, n.a. = no comprehensive estimate available, P = urban and/or industrial pollution, S = increased water turbidity/load of sediment, beach replenishments, T = tourism, WD = wasting disease. Note that the estimates of covers should be regarded as minimal representation of the actual coverage in most cases. Also note that most often time span is not indicated.

n.a. n.a. n.a. n.a. >1000 in Mediterranean

Romania Russia Slovenia Spain





N

0

250 500

1,000 Km

Figure 9 Map of the distribution of seagrasses in Europe. (Data courtesy of UNEP World Conservation Monitoring Centre.) The map should be considered indicative of general distribution not areal extent.

(Sandulli et al. 1994). The extent of seagrass degradation is, however, difficult to quantify even locally. Historical losses and causes There is increasing awareness about the severe degradation of seagrass meadows (e.g., see, among others, the reviews by Short & Wyllie-Echeverria 1996, Hemminga & Duarte 2000, Duarte 2002). Reports consistently identify a long-term trend of worldwide seagrass decline, about 70% of which can be probably attributed to direct human-induced disturbance (Short & Wyllie-Echeverria 1996). Less information is available concerning the degradation caused by indirect impacts (Duarte 2002). It has been estimated that a global loss of 12,000 km2 occurred during the 1990s alone (Short & Wyllie-Echeverria 1996), which represents about 7% of the known distribution of seagrasses (Green & Short 2003). Data covering longer time spans are rare. Based on data from 12 temperate estuaries around the world, it has been estimated that over time these systems may have lost approximately 65% of their seagrasses (Lotze et al. 2006). No comprehensive organised historical information seems to be available for Europe and information is limited to restricted areas (Table 3). There are different trends for seagrass losses in northern and Mediterranean European countries (Green & Short 2003). In northern Europe, before the early 1900s, several seagrass species, including Zostera marina, were common. They were harvested for a variety of uses, including use for fuel, packing, upholstery, insulation, roof material, filling of mattresses and cushions, feeding and bedding for domestic livestock, fertiliser and as resource to obtain salt. Their abundance was, however, severely reduced in the 1930s by a ‘wasting disease’, caused by the pathogenic slime mould Labyrinthula zosterae (e.g., Den Hartog 1987). The disease led to the catastrophic die-back 372



of eelgrass (Zostera marina) meadows along the north Atlantic coasts, with loss of almost 90% of the Zostera populations in north Atlantic western Europe. Some beds progressively recovered, but substantial areas remain lost from most beds. There is uncertainty about the causes of this disease outbreak and there has been much debate about whether concomitant human impacts that already weakened the plants contributed to the outbreak (Den Hartog 1987). The decline particularly affected sublitoral beds, while intertidal populations were less affected. Probably the best records of the disruptive effects of the wasting disease are from Denmark, where records of eelgrass distribution date back to 1900 (Boström et al. 2003). In 1900 eelgrass covered about 6726 km2 (Figure 10); by 1940, 93% of the distribution of vegetated areas was lost. Since 1960, there has been a slow recolonisation. Currently, there is approximately 20–25% of the distribution recorded in 1900 (i.e., a total loss of 75–80%). The greatest loss has been in deep Zostera populations and in Denmark the vertical depth distribution of Z. marina was reduced by about 50% during the twentieth century, from a historical depth limit of 5.6–11 m to a recent limit of 2.5–8 m in sheltered and exposed areas, respectively (Hemminga & Duarte 2000, Baden et al. 2003). Similar dramatic losses are described for the United Kingdom (Davison & Hughes 1998 and references therein) and the Wadden Sea (Reise 1994, Wolff 2000, Lotze 2005). The distribution of seagrasses in the United Kingdom was only systematically described in the 1930s after the outbreak of wasting disease, when Z. marina was already scarce and restricted to few sheltered lagoons but there are indications that seagrasses were widespread until 1917 (Davison & Hughes 1998). There is some uncertainty about when recovery started. According to some, recovery began in 1933 and was quite rapid, with some beds fully recovering within a few years of the 1930s epidemic, while according to others the disease continued to affect Zostera populations until the mid-1940s and recovery did not really begin until the 1950s. Nowadays, most Zostera beds have not fully recovered, particularly in southern and eastern England where the species was once abundant, and only 20 of Britain’s 155 estuaries have eelgrass meadows >1 ha in extent (Davison & Hughes 1998). Before the 1930s the Wadden Sea also contained large subtidal, seagrass beds. These have been exploited since historical times for construction and insulating material and to fill mattresses and cushions. In the Dutch Wadden Sea, from the thirteenth century to 1825, eelgrass was used to build dykes. The construction of 1 m of seawall required about 8–20 m3 of compacted eelgrass, equivalent to about 40–100 m3 of fresh eelgrass, which in some years corresponded to about 1–10% of the annual production (Wolff 2005). Decline of seagrasses appears to have occurred over two phases (Reise 1994): one acute in the 1930s, caused by the wasting disease, after which most subtidal eelgrass beds did not recover, and one more gradual that began in the 1960s, mostly driven by eutrophication. These declines first affected subtidal eelgrass beds, and subsequently also intertidal ones, leading to the almost extinction of seagrasses in some regions of the Wadden Sea (e.g., the Dutch Wadden Sea, where cover dropped from 150 to 1–2 km2; see Table 3), and to the disappearance of numerous species associated with seagrasses (Wolff 2000). Along Mediterranean coasts, reliable estimates made by direct observation of the area of seagrass lost or degraded are limited (Green & Short 2003). It is estimated that in the past Posidonia oceanica meadows may have covered 50,000 km2 in the whole basin (Duarte 2002), which considering present estimated covers of seagrasses in the Mediterranean and Euro-Asian Seas (Green & Short 2003) would make an overall loss >85% (but probably many existing seagrass meadows are not presently documented). Rapid local regression (up to complete disappearance) of P. oceanica meadows is known to have occurred at numerous localities in France, Italy and Spain (Table 3). It is estimated that shoot density of P. oceanica in the western Mediterranean has decreased by up to 50% over a few decades, with major losses between 10 and 20 m depth (EUCC 1998). For the French mainland coast, overall habitat loss is estimated as about 10–15%, which would increase up to 30–40% if the decline in shoot density is also taken into account. Overall, these figures are 373



North Sea

1901

Sweden

Kattegat Denmark ∅resund

Baltic Sea Germany

North Sea

1941

Sweden


Baltic Sea

Germany North Sea

1994

Sweden


Baltic Sea

Germany

Figure 10 Area distribution of eelgrass Zostera marina (in dark grey) along Danish coasts in 1901, 1941 and 1994. (Modified after Boström et al. 2003. With permission.) In the 1930s, eelgrass populations were severely affected by a wasting disease, and in 1941 they covered only 7% of the areas occupied in 1901. Recolonisation took place after the 1960s, but in 1994 cover was still only 20–25% of that in 1901.

374



considered a good estimate for most western Mediterranean coastlines, with notable exceptions around the islands and in the eastern Mediterranean (EUCC 1998). There have been clear losses of seagrasses on the Italian coast of the north Adriatic Sea. Geological data have shown that seagrass beds were probably common before the 1800s and experienced dramatic regressions to virtual extinction in the last two centuries (Barmawidjaja et al. 1995, Caressa et al. 1995, Rismondo et al. 1997). For example, faunal changes and the sudden disappearance of epiphytic foraminiferans in a sediment core in front of the Po river delta suggest that seagrass beds were present up to 1840 in this area, and that increased load of fine sediment and nutrients between 1840 and 1870, due to substantial changes to the main outflow canals of the Po river, was probably the main cause of their sudden disappearance (Barmawidjaja et al. 1995). Similarly, dead beds of P. oceanica have been found at several sites about 8 miles offshore from the Venice lagoon, indicating the likely past presence of P. oceanica (Rismondo et al. 1997). In the Gulf of Trieste the regression has been more recent and there has been some direct documentation. P. oceanica was reported as common in the Gulf of Trieste at the beginning of the 1900s (Caressa et al. 1995). In 1938 the species had declined but was still present all over the coastlines of the Istrian peninsula, while just about 30 yr later only the Koper meadow was detected. Nowadays, P. oceanica is present only in a fragmented meadow along the coastline of Koper (Slovenia, at the southern side of the Gulf of Trieste) and in a very small area along the coasts of Grado (on the Italian side of the Gulf of Trieste). This drastic reduction has been attributed to a steep increase in water pollution during the last 50 yr as a consequence of industrial and harbour development in the Gulf of Trieste (Caressa et al. 1995). Another case of rapid regression of P. oceanica meadows is in the French Riviera, from Menton to the Rhône delta (Meinesz et al. 1991). Intensive waterfront development started around 1800 covering natural coastal habitats with recreational harbours, artificial beaches, landfills (for the tourist industry) and large commercial and military complexes, ports and airports. A total of 185 reclamation projects ‘occupied’ 106 km (16.2%) of the coastline, directly removing 30.57 km2 of bottom substrata and greatly affecting surrounding areas (by modifying water movements and sedimentation patterns). Overall a total of 9.7% of the shallow water zone between 0 and −20 m and 14.5% of the zone between 0 and −10 m were irreversibly destroyed. The vast majority of this area was occupied by P. oceanica meadows, which were estimated to originally cover a total of 200 km2. Trends and threats Many anthropogenic factors are considered responsible for the ongoing degradation and decline of seagrasses in Europe as well as globally (e.g., see, among others, the reviews by Short & WyllieEcheverria 1996, Davison & Hughes 1998, Hemminga & Duarte 2000, Duarte 2002, Green & Short 2003). The most important of these threats is likely to be poor water quality from pollution, eutrophication and excess sedimentation. These impacts are associated with, and enhanced by, urban and tourist waterfront developments, port constructions, beach replenishments and other interventions for shoreline stabilisation. Severe seagrass loss is still in progress in Europe, as evident along 200 km of the Skagerrak Swedish coast (Baden et al. 2003). Here, 50 of 69 mapped meadows of Zostera marina have shown average declines of 58% between the 1980s and 2000, corresponding to a lost surface of about 1061 ha. Most declines are related to a reduction of the upper and the lower depth distribution of seagrasses, resulting in narrower meadows, but in some areas seagrass meadows have disappeared completely, with dramatic effects on the fish assemblages (Pihl et al. 2006). The reasons for this continuing loss of seagrasses are not known but could be related to an excess growth of phytoplankton and filamentous or other ephemeral macroalgae as a consequence of eutrophication. These micro- and macroalgae outcompete seagrasses (Hauxwell et al. 2001) and decompose on the bottom, 375



favouring anoxia events like those that caused the recent temporary disappearance of Z. marina in 1994 around the island of Funen in Denmark (Rask et al. 1999). There is also current seagrass loss in the Mediterranean Sea as a consequence of the invasion of Caulerpa taxifolia (Meinesz et al. 2001, but see Jaubert et al. 2003). This species competes for space and resources with the seagrass Cymodocea nodosa (Ceccherelli & Cinelli 1997) and is thought to be able to damage Posidonia oceanica beds, particularly when these are already under stress (e.g., de Villèle & Verlaque 1995). Protection measures Green & Short (2003) report that worldwide there are only some 247 MPAs that are known to include seagrasses, spread in 72 countries and territories. This estimate may be conservative but these authors note that this is far smaller than the number of MPAs with coral reefs (more than 660) or mangrove forests (over 1800). Even for these 247 MPAs, there are questions about their effectiveness in protecting seagrass ecosystems particularly from threats such as poor water quality (e.g., Marbà et al. 2002, Milazzo et al. 2004). In Europe, numerous initiatives arising from the Rio CBD, the Habitats Directive and the Birds Directive (Table 1) have led to seagrass meadows being specifically targeted for conservation and restoration. Seagrasses are a named component of several habitats in the Habitats Directive, including ‘Coastal lagoons’ (a priority habitat), ‘Sandbanks slightly covered by seawater all of the time’, ‘Large shallow inlets and bays’, ‘Estuaries’ and ‘Mudflats and sandflats not covered by seawater at low tide’ (EC 2003). Furthermore, Posidonia oceanica beds are a priority habitat in Annex I of the Habitats Directive. International concern about the conservation of seagrass beds has also led to the banning of trawling on seagrasses in EC waters (Tudela 2004). As a response to these initiatives, European States have also developed national strategies and initiatives. The U.K. Biodiversity Action Plan, for example, includes a Habitat Action Plan for seagrass beds (U.K. Biodiversity Group 1999). Accordingly, areas of seagrass are included in some coastal ASSIs/SSSIs, Ramsar sites, SPAs and voluntary MPAs. Two of the three U.K. Marine Nature Reserves have seagrass beds and the habitat occurs in a number of areas proposed as SACs.

Macroalgal beds Current distribution and status For the purpose of this review ‘macroalgal beds’ refer to kelps, fucoids and other complex, erect brown and red macroalgae that produce relatively large biogenic habitats. Macroalgal beds form diverse, productive and valuable temperate coastal ecosystems (Steneck et al. 2002). They are widespread on shallow hard substrata around Europe (Birkett et al. 1998b, Steneck et al. 2002, Thibaut et al. 2005 and references therein), including rock, boulders, cobbles and human-made structures from the intertidal down to more than 30 m in depth. Laminaria and Fucus are the main genera along the coasts of northwest Europe, while Cystoseira and Sargassum are the main genera in the Mediterranean Sea. The macroalgal flora of the European coasts are among the best known and studied. Furthermore, large macroalgae have been the target of ecological studies for over a century and there is extensive literature on physical and biological factors operating in these habitats (see, among many, Lüning 1990, Ballesteros 1992). Surprisingly, however, quantitative information on the distribution of macroalgal beds is limited and their extent is unknown. At present, there do not appear to be comprehensive inventories for the macroalgae of any European country and it is difficult to estimate the areal coverage even at regional or local scales. Distribution maps of rocky coast biotopes, 376



including kelps, are available for the United Kingdom (Connor et al. 2004), and ongoing inventories by the Joint Nature Conservation Committee (JNCC) are likely to result in estimates of the cover of macroalgal beds for this country, at least for intertidal habitats, while little information is available for the subtidal. Intensive research on kelp distribution, ecology and effects of human harvesting has also been carried out along Norwegian coasts, where it is estimated that the dominant kelp Laminaria hyperborea might cover between 5,000 and 10,000 km2 (Jensen 1998). Limited quantitative information on the distribution of macroalgal beds is available at a few regional or local scales (e.g., the Albères coast, Thibaut et al. 2005; Kiel Bay, Vogt & Schramm 1991; the Öregrund archipelago, Eriksson et al. 1998; the Gullmar Fjord, Johansson et al. 1998). Historical losses and causes A perceived worldwide decline of macroalgal beds has been reported during the last decades (Steneck et al. 2002, Airoldi 2003). Such decrease most often appears to be parallelled by a trend of increasing abundance of turf-forming, filamentous or other ephemeral algae (e.g., Munda 1993, Eriksson et al. 1998, Benedetti-Cecchi et al. 2001, Lotze 2005) that, once established, often inhibit recolonisation of canopy-forming algae and other organisms (Airoldi 1998). Canopy-forming algae and turfs have been suggested to represent alternative states in shallow temperate rocky reefs under different disturbance and stress regimes (Worm et al. 1999, Airoldi 2003, Connell 2005). Macroalgal beds have also been replaced by coralline dominated ‘urchin barrens’, where outbreaks of urchins may have been the primary cause of macraolgal extirpation (Hagen 1995, Steneck et al. 2002, Guidetti et al. 2003), or by mussel beds (Thibaut et al. 2005). In Europe, almost no information on trends in abundance of macroalgae is available before the 1900s. In the twentieth century, conspicuous losses, sometimes to virtual local disappearance, of complex macroalgae have been documented for coastal areas in several countries, including Iceland, Norway, Britain and Ireland (Hagen 1995, Steneck et al. 2002 and references therein); Sweden (Lundälv et al. 1986, Eriksson et al. 1998, 2002, Johansson et al. 1998, Nilsson et al. 2004); Denmark (Middelboe & Sand-Jensen 2000); Finland and Germany (Kangas et al. 1982, Messner & von Oertzen 1991, Vogt & Schramm 1991, Schories et al. 1997); Lithuania (Olenin & Klovaité 1998); Italy (Sfriso 1987, Cormaci & Furnari 1999, Benedetti-Cecchi et al. 2001, Guidetti et al. 2003, L. Airoldi unpublished data); France and Spain (Rodríguez-Prieto & Polo 1996, Thibaut et al. 2005 and references therein); Croatia (Munda 1993, 2000) and Romania (Zaitsev 2006). The causes of these losses are various but mainly include outbreaks of sea urchins and decreased water quality as a consequence of pollution and/or enhanced sediment loads. These trends have been generally traced through comparisons with historic floristic records, which are difficult to translate into an estimate of the extent of habitat loss. Furthermore, complex native species were often replaced by simpler macroalgae or non-native species, possibly affecting the status and functioning of the systems but not their extent. Frequently, the declines of macroalgae have been greater at depth so that their depth range has become shallower (Lumb 1990, Messner & von Oertzen 1991, Bokn et al. 1992, Munda 1993, Eriksson et al. 1998, 2002, Pedersén & Snoeijs 2001, Thibaut et al. 2005). In Kiel Bay (Germany, western Baltic), for example, there has been a >90% decline in the biomass of Fucus spp. between 1950 and 1988, from about 40,000–45,000 t wet wt down to only 2400 t wet wt (Vogt & Schramm 1991). In the 1950s, Fucus spp. were the dominant macrophytes down to 6 m in depth and were still frequent in the 1970s, whereas at the end of the 1980s, the species were not found at depths >2 m. In the Öregrund archipelago (Sweden) the average depth penetration of F. vesiculosus decreased significantly by 2 m between 1943 and 1996 and this species had completely disappeared at depths >8 m (Figure 11). The amount of macroalgal habitat lost as a consequence of this depth reduction has not been estimated (B.K. Eriksson personal communication) but it must have been 377



0

Max. depth

Mean depth

∗ 2

Depth (m)

4 6 8 1943–44 10

1984 1996

12

Figure 11 Changes in the depth (maximum and mean) distribution of Fucus vesiculosus in 1943–1944, 1984 and 1996 in the Öregrund archipelago. (From Eriksson 2002, based on Kautsky et al. 1986, Eriksson et al. 1998 and Eriksson & Bergstrom 2005; courtesy B.K. Eriksson.) *No mean depth data were available for 1984. Data are averages ± 1 standard error. Maximum and mean depth penetration of F. vesiculosus decreased significantly both between 1943–1944 and 1984 and between 1943–1944 and 1996 (t-test, n = 5 sites, p < .05).

considerable because quantitative measures of the distribution of F. vesiculosus indicated that this species covered on average 20–50% of the substrata. Similarly, at Stora Bornö Island, on the Swedish Skaggerak coast, the depth distribution of macroalgae declined on average by 2.8 m between 1941 and 1998 (Eriksson et al. 2002). This loss was particularly severe for large, complex macroalgae (>50 cm), which showed up to 8-m reductions in their depth distribution (Figure 12). These complex macroalgae were replaced by simpler, thin filamentous and sheet-like forms. Losses to virtual extinction of macroalgal species have been reported from other regions in the Wadden Sea, southern France, the Venice lagoon and the Black Sea, for example. In the Wadden Sea, at least 10 species of macroalgae have become extinct during the past 2000 yr, probably because of the transformation of brackish waters into fresh waters and the destruction of native eelgrass beds and oyster reefs that provided solid surfaces for attachment (Wolff 2000). Along the 1941

1998

2

Depth (m)

4 6 8 10 12 14 16 300

200

100 0 100 200 Estimate of cover (%)

300

Figure 12 Depth distributions of macroalgae with different thallus shape (thin filamentous and sheet-like = white bars; coarse filamentous and thick, leathery algae = dotted bars) in 1941 and 1998 at Stora Bornö Island. (Based on data from Eriksson et al. 2002. Courtesy B.K. Eriksson.) Bars are mean percentage cover in each depth interval pooled for two vertical profiles (error lines are not shown for clarity).

378



Albéres coasts (southern France) dramatic reductions in abundance to extinction of populations of Fucales (Cystoseira spp. and Sargassum spp.) occurred between the early 1900s and 2003 (Thibaut et al. 2005). Only 9 of 14 species of Fucales documented in 1912 were present in 1978, with the genus Sargassum entirely lost, and only 5 species were found in 2003. Seven of the extinct species were considered frequent to abundant in 1937. Of the 5 species remaining in 2003, only 1 did not show signs of regression. In the Venice lagoon, some 141 algal species were documented in 1938, 116 in 1962, 107 in 1987 and 96 in 1991 (Sfriso 1987, Sfriso & La Rocca 2005). Some of the most damaged species in the lagoon were previously dominant large brown algae, including species of Cystoseira, Fucus virsoides, and Sargassum hornschuchii. These algae were adversely affected when channel excavations limited water exchange, leading to an increase of nutrient levels and eutrophication within the lagoon, and in the 1980s their abundance dramatically decreased, sometimes to complete and permanent disappearance. These complex macroalgae were replaced by ephemeral species, mainly green algae, and invasive species. In the last 10 yr, water quality has improved in the Venice lagoon, and recent sampling detected an increase in the number of macroalgae (Sfriso & La Rocca 2005), in part also linked to the introduction of several invasive species, including the now-abundant large brown algae Undaria pinnatifida and Sargassum muticum. Along the coasts of the Black Sea, several complex macroalgae (including brown algae of the genera Cystoseira and Phyllophora) have virtually disappeared during the past 30 yr (Zaitsev 2006). Estimates at some Romanian localities indicate that in 1971 these species attained considerable biomasses (about >10,000 t fresh wt). Trends and threats Several factors are thought to be responsible for the continuing decline of kelps and canopy-forming macroalgae and to pose serious threats to the future of rocky reefs in general (Steneck et al. 2002, Thompson et al. 2002). Urbanisation is thought to have the most disrupting effects on kelps and other canopy-forming algae, particularly by affecting water clarity and quality as well as other habitat-related changes (e.g., Vogt & Schramm 1991, Munda 1993, Eriksson et al. 1998, 2002, Benedetti-Cecchi et al. 2001). In some northern European regions, including the west coast of Norway, the French channel coast and parts of the U.K. coast, harvesting is also an issue (e.g., Christie et al. 1998). Kelps washed up on the shore have been traditionally collected for centuries in some regions, for use as an agricultural fertiliser and to improve the soil structure. Nowadays kelps and fucoids are harvested from living beds to be used as basic resource in the alginate industry to produce emulsifying and gelling agents. The most commonly harvested species include Laminaria hyperborea, L. digitata, Ascophyllum nodosum and Fucus spp. with 70,000–80,000 t of seaweeds collected each year around the coasts of both Brittany (Birkett et al. 1998b) and Norway (EEA 2002). Modern methods of kelp harvesting (e.g., by trawling) seem to have a significant direct influence on kelp biotopes (Birkett et al. 1998b, Christie et al. 1998). Reef habitats and associated macroalgal beds are also severely damaged by disruptive fishing techniques. For example, the collection of the date mussel Lithophaga lithophaga by use of hammers and chisels, pneumatic hammers and explosives is still a widespread practice in most Mediterranean countries, despite its legal ban. This practice directly and irreversibly destroys the rocky environment, causing the loss of canopy-forming seaweeds and the formation of barrens (Fanelli et al. 1994, Guidetti et al. 2003). The introduction of harbour piers, jetties, dykes, seawalls, coastal defences and other armoured artificial structures has in some regions led to an expansion in the distribution of native and nonnative macroalgae and other rocky-bottom species (Moschella et al. 2005). In the Wadden Sea, for example, where hard substrata were naturally scarce, about 730 km of artificial structures have 379



introduced about 2–4 km2 of hard substrata, providing new habitats for a variety of rocky-bottom species, including Laminaria saccharina (Reise 2005). Along the north Adriatic shores, which are naturally devoid of rocky substrata, >190 km of rock-armoured structures (Figure 5), built mainly in the past 40 yr (Bondesan et al. 1995), have introduced about 1 km2 of artificial hard substrata within natural sandy depositional environments, which are now extensively colonised by the nonindigenous, invasive canopy-forming macroalga Codium fragile ssp. tomentosoides (Bulleri & Airoldi 2005). The extent of hard coastal structures is expected to increase in the future, with profound but overlooked ecological consequences on native coastal environments (Airoldi et al. 2005). Protection measures Although kelp beds and other macroalgal habitats are not specifically targeted in the Habitats Directive, species of the genus Fucus, Laminaria and Cystoseira and other macroalgae are named components of ‘Reefs’ habitat (EC 2003). Other European initiatives also include the protection of some species of complex macroalgae. For example, six Mediterranean species of the genus Cystoseira and two species of Laminaria are listed in Annex I of the Bern Convention. The Action Plan for the Conservation of Marine Vegetation in the Mediterranean Sea, adopted within the framework of the Barcelona Convention, identifies the conservation of Cystoseira belts as a priority. Several complex brown macroalgae are listed in the Red Books of Mediterranean and Black Seas as endangered (e.g., Boudouresque et al. 1990, Zaitsev 2006). Furthermore, Lithophaga lithophaga is included in Annex IV of the Habitats Directive and its collection is banned in most Mediterranean countries to protect rocky reefs and associated macroalgal beds from the destructive consequences of the fishery for this rather abundant date mussel (Russo & Cicogna 1991). There are also national initiatives. For example, the commercial harvesting of kelps is strictly regulated in Norway and in Brittany (Birkett et al. 1998b), including a system of rotation of harvested areas introduced by the Norwegian government to ensure that each area of kelp forest is harvested only once every 4 yr.

Biogenic reefs: oyster reefs Current distribution and status The native European flat oyster (Ostrea edulis) is a sessile, filter-feeding, bivalve mollusc that used to be very abundant throughout its range (Korringa 1952). It is associated with highly productive estuarine and shallow coastal water habitats with sediments ranging from mud to gravel. The natural distribution of O. edulis is along the European Atlantic coasts from Norway to Morocco and across the coasts of the Mediterranean and Black Seas. Their abundance declined significantly during the nineteenth and twentieth centuries and wild native beds were considered scarce in Europe as early as the 1950s (Korringa 1952, Yonge 1966, Mackenzie et al. 1997). Remains of wild native oyster beds still occur in various regions, including the rivers and flats bordering the Thames Estuary, the Solent, River Fal, the west coasts of Scotland and Ireland (Kennedy & Roberts 1999, U.K. Biodiversity Group 1999, Tyler-Walters 2001, Jackson 2003), the western part of the Swedish Kattegat region of the Baltic (Lozan 1996), the Limfjord region of Denmark (Korringa 1952), the Adriatic Sea, where O. edulis is still captured in the wild (Barnabe & Doumenge 2001), the Mar Menor (Spain), where a large flat oyster population, estimated at over 100 million individuals, still produces large amounts of spat (Cano & Rocamora 1996), and areas of the Black Sea, where the species was still valuable commercially until the 1970s (Zaitsev 2006). Limited information, however, is available about the current status of these oyster reefs and there 380



is debate about whether the fragmented patches of wild oyster habitats are self-sustaining or owe their survival to the inputs of larvae from cultivated oysters (Korringa 1952). Nowadays, aquaculture provides the main supply of native oysters in most European countries (Mackenzie et al. 1997, Ocean Studies Board 2004). This industry has also been seriously affected by epidemic diseases in recent decades, with documented losses of commercial stocks above 80% in France (Kennedy & Roberts 1999, Ocean Studies Board 2004), and most Mediterranean native oyster beds are in such poor conditions that they are unable to support intensive culture (Barnabe & Doumenge 2001). Although marketplace demand for native oysters remains strong, the introduced Pacific oyster Crassostrea gigas, which is easier to cultivate than the native oyster, now provides the major share of oyster production in Europe (Cano & Rocamora 1996, Kennedy & Roberts 1999, Lotze 2005; see Figure 13C). Historical losses and causes There is some documentation on the declines and loss of native oyster reefs, mostly from fishery landing records; direct quantitative data are uncommon. In Europe oysters have been an extremely popular food for centuries (Jackson 2003). Both ancient Greeks and Romans highly valued oysters. Romans fished and imported them from all over European and Mediterranean coastlines and extensively cultivated them (Günther 1897), and in some British estuaries there are archaeological signs of overexploitation of native oyster beds since the first century (Rippon 2000). For centuries, Ostrea edulis reefs supported a productive commercial fishery (Mackenzie et al. 1997). In the eighteenth and nineteenth centuries, large offshore oyster grounds in the southern North Sea and the English Channel produced up to 100 times more than today’s 100–200 t (U.K. Biodiversity Group 1999, Berghahn & Ruth 2005). The richest natural oyster beds in Europe until the nineteenth century were probably around Britain, from Stornoway to the Solway in the west and from the Orkney Islands to the Firth of Forth in the east (Berghahn & Ruth 2005). In the mid-nineteenth century these were heavily exploited; dredging of the oyster beds was one of the largest fisheries, employing about 120,000 men around the coast in the 1880s (Tyler-Walters 2001), with an annual yield of >50 million oysters (Berghahn & Ruth 2005). Oyster reefs at Strangford Lough, in Ireland, once supported up to 20 boats employed in oyster dredging (Kennedy & Roberts 1999). In the Wadden Sea, the commercial fishery for oysters started in the eleventh century and flourished in the eighteenth century (Figure 13): in 1765 large oyster beds between Texel and Wieringen supported profitable fishery by 145 vessels, with catches over 100,000 oysters yr−1 vessel−1 (Wolff 2005). By the late nineteenth century, beds of O. edulis were already severely depleted or physically destroyed around most European coasts (Ocean Studies Board 2004). Regulations and fishery closures were imposed in some regions. In the Wadden Sea, for example, management strategies including minimum landing size, fishery closures, a licence system and maximum yield per year have been applied since the seventeenth century (Berghahn & Ruth 2005). Similar initiatives were taken in France. The decline could not, however, be halted and in the twentieth century catches collapsed (e.g., Figure 13A,B). Overfishing and wasteful exploitation, combined with outbreaks of diseases, habitat loss and change or destruction, reduction in water quality and other large-scale environmental alterations, adverse weather conditions, and the introduction of non-native oysters (and associated parasites and diseases, such as the protozoan Bonamia ostreae) for aquaculture and other non-native species (e.g., the invasive gastropod Crepidula fornicata) were blamed for the decline (Korringa 1952, Mackenzie et al. 1997, Wolff 2000, Jackson 2003, Berghahn & Ruth 2005). Virtual extinction of native oyster beds has been documented in the Wadden Sea, where wild oysters largely disappeared by 1950 (Wolff 2000, 2005, Lotze 2005); in Helgoland (Germany), where beds largely disappeared by the mid-1900s (Korringa 1952, Franke & Gutow 2004, OSPAR Commission 2005); in the Dutch Easter Scheldt (van den Berg et al. 2005); in Belgium (OSPAR Commission 381



A

0.3 0.2 0.1 0.0

1770 1780 1790 1800 1810 1820 1830 Oysters harvested (× 106)

6 B

5 4 3 2 1 0

1870 1880 1890 1900 1910 1920 1930 25 C 20 15 10 5 0 1970 1975 1980 1985 1990 1995 2000 Figure 13 Annual landings of native European oysters, Ostrea edulis, (A) in the east Frisian Wadden Sea during 1770–1830, (B) in the north Frisian Wadden Sea during 1868–1930 and (C) landings of cultured Pacific oysters, Crassostrea gigas, in the Netherlands during 1970–2000. (From Lotze 2005. With permission.)

2005); in all deeper waters of the southern North Sea, such as in the Oyster Grounds (OSPAR Commission 2005); in most areas of Galicia (Cano & Rocamora 1996) and in some bays in the Black Sea (Zaitsev 2006). In the Firth of Forth (Scotland), which in past centuries had hosted one of the most famous oyster banks, no oysters were found in 1957 (Dodd 2005). Dramatic stock decreases have been reported as well on the Atlantic coasts in French Brittany, the Netherlands, Denmark, Norway, Ireland and England and in the Mediterranean Sea (Korringa 1952, Mackenzie et al. 1997, U.K. Biodiversity Group 1999, Barnabe & Doumenge 2001). In the United Kingdom, where 700 million oysters were consumed in London alone in 1864, the catch fell from 40 million in 1920 to 3 million in the 1960s and has not recovered (Tyler-Walters 2001). In Archachon Basin (France), 382



wild oysters, which had been exploited for ages, were nearly commercially exhausted in the mid1800s, leading to the introduction first of the Portuguese hollow oyster (Crassostrea angulata) and then of the Pacific oyster (Crassostrea gigas). Trends and threats There is limited information about current trends and threats to remaining native oyster reefs in Europe. Ostrea edulis is a relatively long-lived species and reproduces sporadically (Korringa 1952). Thus, presumably, times of recovery from overexploitation or other causes of damage are very long and are estimated to take up to 20 yr (Jackson 2003). O. edulis is considered to be highly sensitive to substratum loss, smothering, contamination by synthetic compounds (particularly tributyltin (TBT) antifouling paints used on ships and leisure craft, which, in the early 1980s, caused stunted growth of oysters and probably affected reproductive capacity), oxygen depletion, reduced freshwater inputs, introduction of microbial pathogens/parasites, introduction of non-native species and direct extraction (U.K. Biodiversity Group 1999, Jackson 2003, Hiscock et al. 2005). All these factors impair recovery as well as restoration efforts. The main factors that probably threaten native oyster reefs nowadays include illegal fishing as well as by-catch in trawling targeting other species, poor water quality and pollution, changes to the environment (e.g., habitat loss due to coastal development) and the introduction of non-native competitors, predators and diseases (Jackson 2003, OSPAR Commission 2005). Protection measures Nowadays, the sparse remains of wild native oyster beds are probably one of the most endangered marine habitats in Europe. Ostrea edulis, however, does not seem to be the target of any specific protection measure, conservation legislation or convention at a European level. ‘Reefs’, including biogenic reefs, are listed as a conservation feature in Annex I of the Habitats Directive; however, native oyster reefs are not mentioned as a component (EC 2003). Since 2003, O. edulis beds are included in the OSPAR list of threatened and/or declining species and habitats for all OSPAR areas (OSPAR Commission 2005). Ostrea edulis is also included in the ‘Red’ lists of some regions (e.g., Wadden Sea, Black Sea). Indirect protection to native oyster reefs may also come from a number of EC Directives related to shellfish, such as the 95/70/EC, which sets community-wide rules to prevent the introduction and spread of the most serious diseases affecting bivalve molluscs, and the Shellfish Waters Directive (Table 1). Fisheries for native oysters are regulated (sometimes prohibited) at a national level (e.g., Hiscock et al. 2005, Zaitsev 2006) but other national or regional conservation initiatives seem to be rare. There is little evidence that this management is leading to recovery of stocks. In the United Kingdom, O. edulis is included in a Species Action Plan under the U.K. Biodiversity Action Plan (U.K. Biodiversity Group 1999) and naturally occurring native oyster beds are considered a nationally scarce habitat (Jackson 2003), although complex regulations still allow some harvesting.

Biogenic habitats: maerls Current distribution and status Maerls (also known as rhodolith beds) comprise several species of crust-forming, free-living (i.e., unattached), calcareous red algae (Donnan & Moore 2003). Over time, they can become abundant enough to form substantial banks of live and dead material, with some European beds dated as older than 5500 yr (Grall & Hall-Spencer 2003). The major maerl-forming species in European waters are Phymatolithon calcareum, Lithothamnion corrallioides and L. glaciale. They can occur 383



in exposed and sheltered environments, from the surface down to 100 m in depth (e.g., near Corsica and Malta), but most typically occur at 20–30 m (OSPAR Commission 2005). Maerl beds are structurally and functionally complex habitats that support rich and diverse assemblages and host many species unique to those habitats. There is also growing evidence that maerl beds have considerable value as nursery grounds for species of commercial interest (Barbera et al. 2003). Records of the presence or absence of maerl biotopes on European coasts are patchy (Birkett et al. 1998a). Detailed studies of maerl habitats have been undertaken only in the past 40 yr and at only a few locations, mainly in France, Norway and Ireland. Large, historically accessible maerl banks are relatively well recorded as a result of commercial interests. The locations of other maerl sites are known from the results of grab-and-dredge sampling during scientific research cruises. In more recent times, scuba divers have reported maerl banks. However, the extent of a maerl bed at any given location, its species composition and the species associated with it remain largely unknown (Birkett et al. 1998a). The most recent and comprehensive overview of maerl beds in Europe was compiled under the EC MAST-funded BIOMAERL project (Donnan & Moore 2003). Unpublished databases are expanding for the United Kingdom and France but an overall inventory has not been attempted. In Europe maerl beds are patchily distributed and relatively restricted in size (Donnan & Moore 2003). They are found throughout the Mediterranean Sea, with important beds in Algeria, at Marseilles, in Corsica and Sardinia and in the Aegean (Birkett et al. 1998a). Maerl beds are also common on the Atlantic coasts, from Norway to Portugal. Spanish maerl deposits are confined mainly to the Ria de Vigo and Ria de Arosa (Galicia, northwest Spain). Maerl beds are relatively rare in the eastern English Channel, Irish Sea, North Sea and Baltic Sea (Barbera et al. 2003), whereas they are particularly abundant in Brittany, with more that 70 beds >1 km2 and some of the largest and thickest beds in Europe and the world (Grall & Hall-Spencer 2003). In Ireland, maerl is widely distributed in the south and southwest (e.g., Galway Bay, Bantry Bay, Roaringwater Bay; De Grave & Whitaker 1999), and Scotland is home to some of the most extensive maerl beds in Europe (Birkett et al. 1998a). Information on the status of present maerl beds in Europe is limited. Most Breton maerl beds are affected by human activities and the only pristine grounds remaining are small compared with the extensive maerl beds that covered several square kilometres in the 1960s (Grall & Hall-Spencer 2003). In 1999, surveys at one of the largest maerl beds in Brittany (Glenan), which was covered with living maerl until extraction started some 35 yr ago, showed that live maerl was rare over most of this bank while species-poor assemblages on muddy bottoms prevailed (Grall & HallSpencer 2003). Even maerl beds included in Breton NATURA 2000 sites are far from pristine and many are severely degraded. Historical losses and causes Information on historical losses of maerl beds in Europe as a consequence of human activities is virtually absent. Maerl has been harvested on a small scale in Europe for thousands of years for use in animal food additives, water filtration systems, acid lake and pond treatment, biological denitrification, toxin elimination, surfacing garden paths and in the pharmaceutical, cosmetics, medical and nuclear industries, but mostly as a cost-effective source of calcium/magnesium soil additive in agriculture and horticulture (Barbera et al. 2003, Grall & Hall-Spencer 2003). Initially, the quantities extracted were small, being dug by hand from intertidal banks, but in the 1970s about 600,000 t of maerl were extracted per year in France alone (Birkett et al. 1998a). Maerl extraction still forms a major part of the French seaweed industry, both in terms of tonnage and value of harvest, although amounts have declined to about 500,000 t yr−1 (Grall & Hall-Spencer 2003). In the United Kingdom and Ireland maerl has been harvested since the seventeenth century (De Grave 384



& Whitaker 1999) with up to 30,000 t yr−1of maerl harvested commercially in the River Fal from 1975 to 1991 (Birkett et al. 1998a). Currently only limited extraction of maerl takes place in the United Kingdom and Ireland (De Grave & Whitaker 1999). The extent to which this historical extraction has affected maerl beds is not known. Comparisons between museum collections made in 1885–1891 and again in 1995–1997 at maerl beds in the Firth of Clyde (Scotland) showed extensive changes, with substantial reduction in size and number of living thalli of Phymatolithon calcareum (Hall-Spencer & Moore 2000). Such changes have been attributed to mechanical impacts of scallop dredging, which started in that area in the 1930s and became particularly intensive in the 1960s through the advent of more powerful boats, more efficient dredges and better processing facilities. The wholesale removal of maerl habitats and significant reductions in diversity and abundance to adjacent areas at five sites around the coasts of Brittany have also been reported (Barbera et al. 2003) and attributed to commercial extraction (Grall & Hall-Spencer 2003). Trends and threats Maerl habitats are considered highly sensitive to overexploitation and other human activities that result in physical disturbance or deterioration in water quality (Barbera et al. 2003), particularly smothering by fine sediments (Wilson et al. 2004). This sensitivity is compounded by long recovery times due to the slow growth (approximately 1 mm yr−1) and accumulation characteristics of maerl beds. The coralline algae that form the maerl are among the slowest-growing species, and substantial deposits take centuries to millennia to accumulate (Hall-Spencer et al. 2003) so that any effects of habitat removal are irreversible over timescales relevant to humans. The major threats to maerl habitats have been recently reviewed (Barbera et al. 2003). The most obvious threats are from the ongoing commercial extraction. The three main areas of commercial exploitation in Europe have been Brittany, Cornwall and the west of Ireland. Nowadays, only limited extraction takes place in Ireland and the United Kingdom (De Grave & Whitaker 1999) but maerl extraction still forms a major part of the French seaweed industry (Grall & Hall-Spencer 2003). It has been predicted that if extraction rates persist at current levels the large Glenan deposit (Britttany) could be exhausted within 50–100 yr (Grall & Hall-Spencer 2003). In addition to the direct effects of harvesting, other direct and indirect impacts on maerl beds have been noted. Damage to the surface of the beds is caused by towed demersal fishing gear, such as scallop dredges, which significantly reduce bed complexity, biodiversity and long-term viability (Hall-Spencer & Moore 2000, Barbera et al. 2003, Hall-Spencer et al. 2003). Hall-Spencer & Moore (2000) found that scallop (Pecten maximus) dredging in the Clyde Sea led to a 70% reduction of live maerl on a previously unexploited bed with no signs of recovery over the subsequent 4 yr. Permanent moorings for pleasure boats can have similar, more localised, effects (Birkett et al. 1998a). The negative effects of increased eutrophication and turbidity in coastal waters both from silt loads and nutrient runoff from agricultural land and aquaculture have been documented in Galicia and in the Bay of Brest (Birkett et al. 1998a, Barbera et al. 2003). Smothering of maerl beds as a consequence of the invasion of the gastropod Crepidula fornicata has been observed in Breton bays (Grall & Hall-Spencer 2003). Maerl beds are also threatened by land reclamation and proliferation of coastal structures that alter circulation patterns (Birkett et al. 1998a, Barbera et al. 2003). Protection measures Although maerl is confined to a relatively small proportion of European shallow sublittoral waters, their conservation importance is being increasingly recognised (Birkett et al. 1998a, Donnan & 385



Moore 2003). Two of the most common maerl-forming species, Phymatolithon calcareum and Lithothamnion corallioides, are now the only algal species specified as requiring appropriate management measures under the Habitats Directive (Annex V). Free-living Corallinaceae are also a named component of the habitat ‘Sandbanks which are slightly covered by sea water all of the time’ (EC 2003). Maerl beds are also included in the OSPAR list and Mediterranean Red Book of threatened habitats (Boudouresque et al. 1990, OSPAR Commission 2005). In the United Kingdom, maerl is the subject of a Habitat Action Plan (U.K. Biodiversity Group 1999) and both L. corallioides and Phymatolithon calcareum are on the long list of species in the U.K. Biodiversity Steering Group Report (U.K. Steering Group 1995). Furthermore, in the JNCC interpretation of the EC Habitats Directive, maerl is identified as a key habitat within the Annex I category ‘Sand banks which are slightly covered by seawater at all times’. This means that a number of SACs being designated under the directive will provide protection to the maerl that they contain. Maerl beds occur in three of 12 demonstration SACs within the United Kingdom, while the Fal and Helford (Cornwall) candidate SAC includes the largest maerl bed in England. Recently, the Board of Falmouth Harbour Commissioners in Cornwall has decided to cease licensing maerl extraction (Hall-Spencer 2005). France has also recognised biogenic reefs, such as maerls, as vulnerable habitats, and some maerl grounds in Brittany lie within SACs (Grall & Hall-Spencer 2003). However, many of these are already severely degraded and are affected by dredge fishing, eutrophication and the spread of Crepidula fornicata. Maerl extraction in Brittany is under the control of the French mining management scheme, with quota schemes (80,000 t in 2001 on the Glenan bank) and regular environmental surveys. However, such quotas are considered not compatible with regeneration of the resource (Grall & Hall-Spencer 2003).

Biogenic formations There are other examples of biogenic habitats that are severely affected and presumably have been subject to massive losses over the centuries. However, most of these losses have probably passed unnoticed and the information is scattered and mainly anecdotal. For example, off-shore rocky formations in the north Adriatic Sea, both organogenic and fossil in nature, have been flattened and reduced in size by trawling or other destructive forms of fisheries (Bombace 2001), sometimes to virtual extinction. Mediterranean ‘coralligenous’ reefs, which are considered one of the most valuable and diverse habitats in the Mediterranean Sea, are degraded and highly threatened by a variety of human activities (Ballesteros 2006) but how much coralligenous habitat might have been lost in the past as a consequence of these activities is not known. Significant declines in the extent of wild intertidal mussel beds have been reported from large coastal areas of Germany, the Netherlands and Denmark, and Mytilus edulis beds are now rare in the Wadden Sea (OSPAR Commission 2005, Wolff 2005) and are considered under threat in the United Kingdom (Hiscock et al. 2005). Large subtidal Sabellaria spinulosa reefs in the German Wadden Sea, which provided an important habitat for a wide range of associated species, have been completely lost since the 1920s (U.K. Biodiversity Group 1999, Wolff 2000) and similar losses have been reported also from areas of the northeast Atlantic and the United Kingdom (OSPAR Commission 2005). A significant contraction in the range of S. alveolata reefs on the south coast of England has occurred over a period of at least 20 yr until 1984. Declines have also been reported in the western part of the north Cornish coast, the upper parts of the Bristol Channel and in North Wales and the Dee Estuary but the causes of this regression are not known (U.K. Biodiversity Group 1999). The scarcity of information probably underlies the lack of adequate policies and protection measures for these biogenic habitats. Biogenic reefs and concretions are all broadly covered as ‘Reef’ by the Habitats Directive but most are not specifically mentioned (EC 2003) and even 386



national and regional initiatives are limited. Coralligenous assemblages are listed in the Red Book of Mediterranean assemblages (Boudouresque et al. 1990) and S. spinulosa reefs are included in the Red List of Macrofaunal Benthic Invertebrates of the Wadden Sea and the OSPAR list of threatened habitats (OSPAR Commission 2005). In the United Kingdom, both S. spinulosa and S. alveolata are the subject of Habitat Action Plans (U.K. Biodiversity Group 1999).

Sedimentary habitats (mudflats, sandflats and subtidal soft bottoms) Current distribution and status Coastal areas are dominated by soft-sediment habitats. The marine biotope classification for Britain and Ireland identifies a number of littoral and sublittoral sediment biotopes (U.K. Biodiversity Group 1999). These are grouped into several major categories (gravels and sands, muddy sands, muds and mixed sediments) and subdivided further according to depth (littoral, infralittoral or circalittoral) and sediment size. Muddy habitats usually occur in sheltered areas, such as sea lochs, enclosed bays and estuaries, whereas sandflats and coarser sediments tend to develop in more exposed situations on the open coast. Distinctions are also sometimes made between estuarine and marine habitats (e.g., OSPAR Commission 2005). Despite the importance of these highly productive soft-sediment habitats, which support large numbers of predatory birds and fishes, providing nursery, feeding and resting areas, and the long history of studies on many aspects of their ecology, no comprehensive inventory of their extent and status is available at a European level, and even regional or local initiatives are rare. The only habitat for which there are some rough estimates of the amount and distribution are intertidal mudflats. In the OSPAR region, the largest continuous area of intertidal mudflats borders the north coasts of Denmark, Germany and the Netherlands in the Wadden Sea, covering around 4990 km2 (OSPAR Commission 2005). In the United Kingdom, intertidal mudflats are widespread, with significant examples in the Wash, the Solway Firth, Mersey Estuary, Bridgwater Bay and Strangford Lough; overall they are estimated to cover about 2700 km2 (U.K. Biodiversity Group 1999). Historical losses and causes The extent of historical losses of soft-bottom habitats is virtually unknown for any country and even regional or local information is scarce. Available information suggests that loss or deep alteration of these types of habitats may have been extremely high, particularly in estuaries and enclosed bays but there is also the possibility that unvegetated soft bottoms have increased their area at the expense of losses in other more structurally complex habitats (e.g., seagrass beds). Past losses are likely to have been related mainly to land claim for agriculture, ports and industrial and urban developments. In the United Kingdom, for example, it is estimated that at least 88% of estuaries have lost intertidal habitats and about 25% of overall estuarine intertidal flats have been removed with peaks of up to 80% in some estuaries such as the Tees (U.K. Biodiversity Group 1999, OSPAR Commission 2005). The amount of intertidal mudflats and sandflats may have also changed significantly over time in relation to coastal erosion, changes in sea levels and human interventions to control these factors (e.g., Lee 2001). Most often, however, these habitats have probably been deeply altered in their fundamental characteristics, including sediment structure and composition, accretion or erosion rates, and inhabiting fauna. As an example, in Europe the massive use of hard defence structures and beach replenishment schemes has deeply changed the structure of shallow surf-zone sediments along whole coastlines in past decades (Airoldi et al. 2005, Martin et al. 2005), as is the case of the north Adriatic Sea (e.g., Figure 5), presumably affecting an enormous and overlooked amount of shallow 387



soft-bottom habitats. Similarly, most sedimentary benthic systems on the continental shelf of Europe have been modified by fishing activities, particularly bottom trawls and dredging, in the last 100 yr (Ball et al. 2000, Frid et al. 2000). In the southern North Sea fishing is thought to have long been the main ecological structuring force on the benthos, to an extent that makes it difficult even to design robust field experiments due to the virtual lack of control areas (Hall-Spencer & Moore 2000). How much of this transformation should be considered as habitat loss or degradation is difficult to quantify. Trends and threats There is limited organised information on the current trends and threats to sedimentary environments (Brown & Mclachlan 2002). Today the major threats to sedimentary habitats are more likely to be linked to further land claim, construction of marinas and slip ways, the widening and dredging of channels for navigation, pipe and cable laying, oil and gas extraction, tourist developments and infrastructures and the construction of sea defences. Some of these threats have slowed considerably in recent years, at least in some European countries, but they have not stopped. In the United Kingdom, many coastal areas, including estuaries, are now either licensed or available for exploration and development (U.K. Biodiversity Group 1999). Pollution from sewage discharge, aquaculture activities, industries and shipping are also important threats to sedimentary environments and associated fauna, leading to anoxic conditions particularly in estuaries and enclosed basins, as observed in Scandinavian and Baltic waters (Karlson et al. 2002), and to long-term accumulation of contaminants (Islam & Tanaka 2004). Physical disturbance by fishing and aggregate dredging activities also represents a major threat to Europe’s sedimentary habitats and associated biota (Lindeboom & de Groot 1998, Tudela 2004), although nowadays highly disturbed seabeds may appear to be relatively unaffected by fishing activities or other physical disturbances (e.g., Hall-Spencer & Moore 2000). On the Dutch continental shelf, the fisheries are now so intensive that every square metre is trawled, on an average, once to twice a year (Lindeboom 1995), and this broadly applies to the entire sea bed of the North Sea (Gray 1997). Direct extraction of sands and gravels for coastal developments, use in the construction industry and beach nourishments are also major, increasing threats for sedimentary habitats (Newell et al. 1998, van Dalfsen et al. 2000). Extraction of sands has steadily increased in most north European countries during the past few decades (ICES 2006b). In the Netherlands, for example, extraction of sands has increased from 35 million m3 yr−1 in 2001, and in the coming decades an average request of 19–43 million m3 yr−1 is expected. Much of the extracted sand is used for beach recharges and coastal defence. Beach nourishments are being increasingly used along European coasts as a ‘soft’ measure to counteract erosion (Hamm et al. 2002), but the consequences of both the extraction and the disposal of sands on sedimentary habitats and biota have received limited attention (Desprez 2000, van Dalfsen et al. 2000, Simonini et al. 2005). Some projections of loss are available for intertidal areas in relation to possible future changes in sea level, recession of coastlines and coastal ‘squeeze’. For example, sea-level rise is projected to cause a loss of 80–100 km2 of intertidal flats in England between 1993 and 2013 (U.K. Biodiversity Group 1999), particularly in southern and southeast regions; the major firths in Scotland will probably also be affected. Protection measures Protection for intertidal and shallow mudflats and sandflats is provided by various international and E.U. agreements, including the Ramsar Convention, the Bonn Convention, the Bern Convention, 388



and the Birds and Habitats Directives (Table 1). In particular, ‘Mudflats and sandflats not covered by seawater at low tide’ and ‘Sandbanks which are slightly covered by sea water all the time’ are listed in Annex I of the Habitats Directive. Mudflats are also included within several other designated Annex I Habitats: ‘Estuaries’, ‘Lagoons’ and ‘Large shallow inlets and bays’. Mudflats are also in the list of OSPAR threatened habitats (OSPAR Commission 2005). Some countries also have national protection measures. For example, in the United Kingdom mudflats are the subject of a Habitat Action Plan (U.K. Biodiversity Group 1999); furthermore, over 300 SSSIs including mudflats have been designated in estuaries and 10 coastal ASSIs in Northern Ireland contain significant areas of mudflats. Soft bottoms deeper than 30 m do not seem to be the target of any specific protection measure at a European level (Hiscock et al. 2005), although a number of EC Directives that regulate water quality provide indirect protection from some types of impacts (Table 1). There are, however, national initiatives. For example, in the United Kingdom ‘Sublitoral sands and gravels’ and ‘Mud habitats in deep waters’ are the subjects of Habitat Action Plans (U.K. Biodiversity Group 1999). Commercial fishing activities are excluded from a number of estuaries and bays around the coast of the United Kingdom, which are important nursery areas for juvenile commercial species (e.g., River Exe, River Conwy and Filey Bay). Fishing activities are prohibited within 500 m of gas and oil platforms, from firing ranges and in close proximity to certain military installations (U.K. Biodiversity Group 1999).

Discussion Population density along European coasts has been growing since ancient times. There are increasingly greater demands and impacts on the habitats and resources in coastal environments. The present review has shown that such intensive exploitation has caused dramatic losses and severe deterioration of native coastal habitats (e.g., Tables 2–4). There are many policies and directives Table 4 Summary of main characteristics of European coastlines and habitats based on reviewed sources Characteristic

Value

Main references

Coastline length Population within 50 kmb Degraded coastlines Years of impactc Artificial coastlines Defended/eroding coastlines Increase in N/P loads 1940s–1980s

325,892 km 200 × 106 85% 2500 yr 22,000 km2 7,600/20,000 km 2- to 4-/4- to 8-fold

Number of invasive species MPAs (Number/Total surface) Present coastal wetlands/loss since 1900s Present seagrasses/historical lossesd Present wild native oyster reefs/historical lossesd Present macroalgal beds/historical lossesd

450–600 1,129/ 236,000 km2 51,910 km2/>65% 7290 km2/>65% Scarce/>90% Unknown/2–4 m in depth

Pruett & Cimino 2000 Stanners & Bourdeau 1995 EEA 1999a Rippon 2006, Lotze et al. 2006 EEA 2005 EC 2004 Nehring 1992, EEA 2001, Karlson et al. 2002 Reise et al. 2006 UNEP/WCMC 2006, MPA Global 2006 Nivet & Frazier 2004, EEA 2006a Duarte 2002, Green & Short 2003 Mackenzie et al. 1997 Vogt & Schramm 1991, Eriksson 2002

a

a b c d

Including islands. In the 1990s. Since beginning of modification and transformation of coastal landscapes. Estimate based on reviewed local to regional sources.

389



Table 5 Past (earlier than 1900 = P), recent (twentieth century = R) and present (N) main drivers of habitat loss along European coasts based on reviewed sources Wetlands Impact Claim/conversion Coastal development Coastal defence Exploitation Water quality Diseases/pests/predators Destructive fishing Aquaculture

Seagrasses

Macroalgae

P

R

N

P

R

N

P

*** ***

*** *** **

** *** ***

** ***

** *** *

* *** **

?

**

** *

** *

*

**

*** *** *** *

*** ** * **

** ** *

? ? ?

R

N

* ** g * *** ** **

** *ga * *** ** *

Biogenic habitats P

R

*

*

*** ** ** ** *

*** ** *** *** **

N

*** ** *** *** ***

Sediments P

R

N

*** ***

*** *** ** *

** *** *** **

?b

?b

Note: Habitats are coastal wetlands (including salt marshes); seagrass meadows; macroalgal beds (kelps, fucoids and other complex macroalgae); biogenic habitats (including oyster reefs and maerls); and sedimentary habitats (mudflats, sandflats and subtidal soft bottoms). Impacts are drainage, embankment, land claim and habitat conversion (e.g., into agriculture land or into freshwater lake); coastal development including urban and industrial developments, ports and infrastructures, marinas and tourist and recreational developments; coastal defence which includes hard structures, beach nourishments and other measures associated with impacts from erosion, storms and sea-level rise; direct exploitation or harvesting; water quality including organic and chemical pollution and altered sedimentation regimes; outbreaks of diseases, introduced species, competitors or predators; destructive fishing techniques (e.g., trawling); aquaculture. Blank cells = nil or modest, * = low, ** = moderate/locally high, *** = high and widespread, g = habitat gain, ? = not known. a b

Negative effects of beach nourishments, enhancement of substratum availability from hard structures. Fishing affects soft bottoms extensively, but it is difficult to evaluate how many soft sediment habitats are lost.

aimed at reducing and reversing these losses (e.g., Table 1) but their overall positive benefits have been low. Coastal habitats are affected by many threats with a few dominant threats that vary somewhat by habitat and over time (Table 5). The greatest impacts to wetlands have been land claim and coastal development with the latter rising in importance over time. The greatest impacts to seagrasses and macroalgae are presently associated with degraded water quality whereas in the past there have been more effects from destructive fishing and diseases. Coastal development remains an important threat to seagrasses in particular. For biogenic habitats, some of the greatest impacts have been from destructive fishing and overexploitation with additional impacts of disease, particularly to native oysters. Coastal development and defence have had the greatest known impacts on soft-sediment habitats with a high likelihood that trawling has impacted vast areas while not causing loss of soft-sediment habitats per se. Shellfish, oyster reefs in particular, have been among the most severely affected of all coastal habitats by overexploitation and other human-driven changes to the environment. By the late nineteenth century, overfishing combined with outbreaks of diseases, habitat transformation, and the introduction of non-native competitors and parasites, had already wiped out wild Ostrea edulis reefs around much of the European coastline. Where documentation is available, it is evident that the loss of shellfish habitat is a major cause of species extirpation and declines in biodiversity (Wolff 2000, Lotze et al. 2005). Currently most native oyster reefs are functionally or entirely extinct in most coastal areas of Europe and they are probably one of the most endangered marine habitats. However, loss and threats to these habitats are largely overlooked and shellfish beds do not seem to be covered by adequate protection measures, conservation legislation or convention at a European level. Oyster reefs are not mentioned specifically in the Habitats Directive, which sets the present framework for habitat protection in Europe. Although less documented and more 390



localised, similar losses seem to have occurred also to other types of shellfish reefs and biogenic formations, such as maerl beds, intertidal mussels beds, Sabellaria spp. reefs and coralligenous formations. The current losses in European coastal habitats are alarming; even worse is the fact that these losses are only measured against recent distributions with little recognition of the compounding impact of centuries and millennia of habitat loss. This historical short-sightedness has been referred to as ‘shifting baselines’, where we only recognise declines in the natural environment relative to the baseline of recent memory in each generation (sensu Dayton et al. 1998, Jackson et al. 2001). At present, there seems to be limited public, political and even scientific awareness of the extent, importance and consequences of such a long history of coastal habitat loss (Lotze 2004). The evidence reviewed in the present work clearly indicates that in some European regions most estuarine and near-shore coastal habitats were probably already severely degraded or driven to virtual extinction well before 1900 and sometimes much earlier than that. Ecological descriptions of coastal marine habitats are rather recent (mid-1900s), and long-term documentation of habitat declines and losses are virtually missing for most systems. Even when documentation is available, evidence tends to be overlooked. For example, Wolff (2000) points out how most Dutch people are inclined to ignore the profound habitat changes and losses that occurred during the past 2000 yr in the Wadden Sea, and they tend to consider present-day systems to be in a ‘natural’ state. Nowadays only a small percentage of the European coastline is considered in ‘good’ condition (EEA 1999c). The degradation has continued at high rates over the past few decades. In France, for example, it is estimated that 15% of natural areas on the coast have disappeared since 1976 and are continuing to do so at the rate of 1% a year (Stanners & Bourdeau 1995). In Italy, around 7000 km2 of coastal marshes were present at the beginning of the 1900s, no more than 1920 km2 in 1972 and fewer than 1000 km2 today (Stanners & Bourdeau 1995). Losses of coastal wetlands and seagrasses exceeding 50% of original area have been documented for most countries where long-term data were available, with peaks above 80% for many regions. Beds of complex macroalgae have been under severe recession since at least the early 1900s. An impressive number of local to regional extinctions of habitats have been documented. Overall, it is estimated that every day between 1960 and 1995, a kilometre of ‘unspoilt’ European coastline has been developed (EUCC 1998). Those fragments of native habitats that remain are under continued threat.

Recommendations for conservation and management Fortunately there is recognition at the European level that the current management and conservation of marine diversity and habitats is insufficient. In particular, there have been recent and significant E.U. policies to identify the coastal and marine habitats of Europe and to develop networks of MPAs around them. These efforts have been inspired in part by the international commitments developed as part of the Rio CBD and the World Summit for Sustainable Development. The development of E.U. MPAs is included in the Habitats Directive and the Birds Directive and embodied in particular in the development of protected areas in the NATURA 2000 network. There are also other regional (e.g., OSPAR) and national initiatives with similar aims for protection of endangered habitats and species (Table 1). These policies are necessary but not sufficient. Indeed MPAs alone have limited use for addressing threats such as poor water quality, disease or even coastal development. Some key needs and opportunities for enhancing the overall conservation and management of coastal and marine habitats in Europe are suggested below. Currently, there is no comprehensive summary of the distribution of habitats along European coastlines and their management is not well informed by adequate knowledge of their distribution and status. Despite millennia of reliance on the resources from coastal and marine ecosystems, we cannot accurately describe the distribution of even the shallowest habitats. Detailed habitat mapping 391



should therefore be given high priority to promote conservation and sustainable management practices. Some databases are available or are being prepared for some countries or habitats but most information is scattered, fragmented or limited to few case studies. At least there has been good progress in developing a consistent habitat classification as part of the European Union Nature Information System (EUNIS) (Davies et al. 2004), which is a necessary precursor to developing a consistent database of habitat distribution. These habitat distributions then need to be compared with the existing protected areas to identify gaps in protection. To fill these gaps, there should be systematic planning for placement of new protected areas and other management measures. Increasingly scientists, agencies and organisations are using systematic planning approaches to target the placement of their protection and management efforts particularly at regional levels (e.g., Beck & Odaya 2001, Possingham et al. 2002, Airamè et al. 2003, Leslie et al. 2003, Beck et al. 2004). These approaches enable decision makers to develop a range of solutions for protection and management and to examine how changes in decisions can affect solutions. The values of these coastal habitats also need to be better assessed to provide real estimates of the ecosystem services that they provide such as pollution regulation, storm hazard reduction, productivity of nurseries for fisheries and recreation (Agardy & Alder 2005). Better valuations of these services will illustrate for communities and governments the real costs of this habitat loss and should provide impetus and economic incentives for their protection and restoration. There are still reasonable opportunities for conservation of coastal habitats in key areas throughout Europe. This protection should be put in place quickly because conservation is cheaper than restoration. The current EC policies are mostly aimed at these protections but they do need to be implemented. Given the extent of damage to coastal habitats, restoration will be required in many places to meet any reasonable goals for conservation and management. Information on historical distributions and loss is extremely important because management goals for these habitats should be based on historical estimates of the distributions of these habitats, not the vastly reduced current distributions (Beck 2003). Even extremely modest goals of 10% protection of the historical distributions of European coastal habitats will require some restoration for many habitats. To meet these goals for conservation and restoration there should be greater involvement by nongovernmental organisations and community groups and there are tools that they can use to contribute directly to conservation and management. These groups have often been involved in efforts to develop MPAs and restore coastal habitats and these efforts should be encouraged and expanded. There are also new tools, such as the private leasing and ownership of marine lands and resources that can be employed more often by private groups to help protect and restore these coastal habitats (Beck et al. 2004). For example, the National Trust in the United Kingdom leases intertidal lands and sea beds from the government along some 180 km of the coast for conservation and restoration. Recently there has been new policy adopted by the French government that allows private groups also to lease subtidal lands as is commonly done by many business interests (e.g., aquaculture industry). Most coastal habitats lie within the exclusive economic zones of individual countries and thus the individual coastal zone management of these countries must be strengthened to protect and manage these habitats. The European Union and member nations have been dedicated to developing better Integrated Coastal Zone Management (ICZM) for some time. The development of strong and effective ICZM programmes has been slow for most nations. These programmes need to advance further to slow and reverse coastal habitat loss. There is much academic and agency interest in developing a more ecosystem-based management (E-BM) approach for managing the many marine resources and the overlapping stakeholder needs for access to these resources (e.g., Browman & Stergiou 2005). E-BM has been incorporated 392



as a central goal of the European Union’s emerging ‘Marine Strategy’ (Table 1). While this approach is needed and sensible, it will take many years to develop and its development should not be allowed to slow efforts to protect and restore habitats now.

Conclusions Europe has seen decades, centuries, even millennia of coastal habitat loss and it continues today. Estuaries, enclosed bays and near-shore shelf habitats around the European coasts are some of the most degraded environments on Earth as they have been used intensively for thousands of years. These functionally valuable coastal ecosystems are still a focal point for human colonisation and use. Before all the services from these ecosystems are lost, efforts should be redoubled to protect the remaining habitats, slow and stop future losses and restore some of these habitats. Recent efforts toward a more sustainable use of these coastal resources have not reversed the trend. There is no single best strategy for addressing these losses and there are many approaches that can help. Many of these are commonsense recommendations that are common in policies within and between nations, but the progress toward them has been slow with a few exceptions. The United Kingdom clearly has made more progress than most nations in the European Union and internationally. If the long history of physical destruction, fragmentation and transformation is further neglected, and if significant, irreversible thresholds are passed, the future sustainability of those few fragments of native or semi-native habitats that remain may ultimately and finally be compromised.

Acknowledgements The work was supported by a grant from The Nature Conservancy. L.A. was further supported by an Assegno di Ricerca of the University of Bologna. We are grateful to Heike Lotze and Keith Hiscock for valuable input and for comments on an earlier draft of the work and to Robin Gibson for his unfailing support. Many other people stimulated our work and gave useful input/references, including Claudio Battelli, Erik Bonsdorff, Paolo Guidetti, Fiorenza Micheli, Karsten Reise and Enric Sala. We are grateful to all the authors who provided figures from their work and the publishers who agreed to their publication. We thank particularly Klemens Erikkson, who kindly prepared original figures for this work, and Sandy Beck for lending her artistry to Figure 2. We also wish to thank Giovanna Branca, Kendra Karr, Caitlyn Toropova, Dan Dorfman, Chris Shepard and Marco Abbiati for their support at various stages of the work. L.A. is particularly grateful to Elena Fuschini for her invaluable assistance with library searches. This publication is contribution number MPS07016 of the E.U. network of Excellence MarBEF.

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CLIMATE CHANGE AND AUSTRALIAN MARINE LIFE E.S. POLOCZANSKA1, R.C. BABCOCK2, A. BUTLER1, A.J. HOBDAY3,6, O. HOEGH-GULDBERG4, T.J. KUNZ3, R. MATEAR3, D.A. MILTON1, T.A. OKEY1 & A.J. RICHARDSON1,5 1Wealth from Oceans Flagship — CSIRO Marine & Atmospheric Research, PO Box 120, Cleveland, Queensland 4163, Australia E-mail: [email protected] 2Wealth from Oceans Flagship — CSIRO Marine & Atmospheric Research, Private Bag 5, Floreat, Western Australia 6913, Australia 3Wealth from Oceans Flagship — CSIRO Marine & Atmospheric Research, GPO Box 1538, Hobart, Tasmania 7001, Australia 4University of Queensland, Centre for Marine Studies, St Lucia, Queensland 4072, Australia 5University of Queensland, Department of Mathematics, St Lucia, Queensland 4072, Australia 6University of Tasmania, School of Zoology, Private Bag 5, Hobart, Tasmania 7001, Australia Abstract Australia’s marine life is highly diverse and endemic. Here we describe projections of climate change in Australian waters and examine from the literature likely impacts of these changes on Australian marine biodiversity. For the Australian region, climate model simulations project oceanic warming, an increase in ocean stratification and decrease in mixing depth, a strengthening of the East Australian Current, increased ocean acidification, a rise in sea level, alterations in cloud cover and ozone levels altering the levels of solar radiation reaching the ocean surface, and altered storm and rainfall regimes. Evidence of climate change impacts on biological systems are generally scarce in Australia compared to the Northern Hemisphere. The poor observational records in Australia are attributed to a lack of studies of climate impacts on natural systems and species at regional or national scales. However, there are notable exceptions such as widespread bleaching of corals on the Great Barrier Reef and poleward shifts in temperate fish populations. Biological changes are likely to be considerable and to have economic and broad ecological consequences, especially in climate-change ‘hot spots’ such as the Tasman Sea and the Great Barrier Reef.

Introduction The global climate is changing and is projected to continue changing at a rapid rate for the next 100 yr (IPCC 2001, 2007). Average global temperatures have risen by 0.6 ± 0.2°C over the twentieth century and this warming is likely to have been greater than for any other century in the last millennium. The 1990s were the warmest decade globally of the past century; and the present decade may be warmest yet (Hansen et al. 2006). Most of the warming observed during the last 50 yr is attributable to anthropogenic forcing by greenhouse gas emissions (Karoly & Stott 2006). The increase in global temperature is likely to be accompanied by alterations in patterns and strength of winds and ocean currents, atmospheric and ocean stratification, a rise in sea levels, acidification of the oceans and changes in rainfall, storm patterns and intensity. Evidence is mounting that the

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changing climate is already impacting terrestrial, marine and freshwater ecosystems (HoeghGuldberg 1999, Walther et al. 2002, Parmesan & Yohe 2003, Root et al. 2003, Walther et al. 2005). Species’ distributions are shifting poleward (Parmesan et al. 1999, Thomas & Lennon 1999, Beaugrand et al. 2002, Hickling et al. 2006), plants are flowering earlier and growing seasons are lengthening (Edwards & Richardson 2004, Wolfe et al. 2005, Linderholm 2006, Schwartz et al. 2006) and timing of peak breeding and migrations of animals are altering (Both et al. 2004, Lehikoinen et al. 2004, Weishampel et al. 2004, Jonzén et al. 2006, Menzel et al. 2006). Most of this evidence, however, is from the Northern Hemisphere, with few examples from the Southern Hemisphere and only a handful from Australia (Chambers 2006). The lack of observations in Australia is attributed to a lack of studies of climate impacts on natural systems and species at regional or national scales. Further, the extent of historical biological datasets in Australia is largely unknown, many are held by small organisations or by individuals and the value of these datasets may not be recognised (Chambers 2006). Because of the unique geological, oceanographic and biological characteristics of Australia, conclusions from climate impact studies in the Northern Hemisphere are not easily transferable to Australian systems. Including fringing islands, Australia has a coastline of almost 60,000 km (Figure 1) that spans from southern temperate waters of Tasmania and Victoria (~45°S) to northern tropical waters of Cape York, Queensland (~10°S). Australia is truly a maritime country with over 90% of the population living within 120 km of the coast. Most of Australia’s population of 20 million live in the southeast with the west and north coasts being sparsely populated. Around 40% of Australia’s population live in the cities of Sydney and Melbourne alone (Australian Bureau of Statistics 2006).

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Figure 1 (See also Colour Figure 1 in the insert following page 344.) Map of Australia indicating the locations discussed in the text. The 200 nm EEZ for Australia is marked by the dashed line, and the 200 m depth contour by the solid line.

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Australia has sovereign rights over ~8.1 million km2 of ocean and this area generates considerable economic wealth estimated as $A52 billion per year or about 8% of gross domestic product (CSIRO 2006). Fisheries and aquaculture are important industries in Australia, both economically (gross value over $A2.5 billion) and socially. Marine life and ecosystems also provide invaluable services including coastal defence, nutrient recycling and greenhouse gas regulation valued globally at $US 22 trillion ($A27 trillion) per annum (Costanza et al. 1997). The annual economic values of Australian marine biomes have been estimated: open ocean $A464.7 billion, seagrass/algal beds $A175.1 billion, coral reefs $A53.5 billion, shelf system $A597.9 billion and tidal marsh/mangroves $A39.1 billion (Blackwell 2005). This assessment assumes Australian marine ecosystems are unstressed so actual values may be lower for degraded systems. Compared to other countries, relatively little is known about the biology and ecology of Australia’s maritime realm, mainly due to the inaccessibility and remoteness of much of the coast as highlighted by the discovery of living stromatolites (representing the one of the oldest known forms of life on Earth) in Western Australia in the 1950s (Logan 1961). Australia is unique among continents in that both the west and east coasts are bounded by major poleward-flowing warm currents (Figure 2), which have considerable influence on marine flora and fauna. The East Australian Current (EAC) originates in the Coral Sea and flows southward before separating from the continental margin to flow northeast and eastward into the Tasman Sea (Ridgway & Godfrey 1997, Ridgway & Dunn 2003). Eddies spawned by the EAC continue southward into the Tasman Sea bringing episodic incursions of warm water to temperate eastern Australia and Tasmanian waters (Ridgway & Godfrey 1997). The Leeuwin Current flows southward along the Western Australian coast and continues eastward into and across the Great Australian Bight reaching the west of Tasmania in austral winter (Ridgway & Condie 2004). The influence of these currents is evident from the occurrence of tropical fauna and flora in southern Australian waters at normally temperate latitudes (Maxwell & Cresswell 1981, Wells 1985, Dunlop & Wooller 1990, O’Hara & Poore 2000, Griffiths 2003). The importance of these major currents in structuring marine communities can be seen in the biogeographic distributions of many species, functional

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Figure 2 Major currents and circulation patterns around Australia. The continent is bounded by the Pacific Ocean to the east, the Indian Ocean to the west and the Southern Ocean to the south. Figure courtesy of S. Condie/CSIRO.

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Figure 3 (See also Colour Figure 3 in the insert.) Phytoplankton provinces around Australia. In northern shelf waters westwards from Torres Strait tropical diatom species dominate, with slight regional differences in relative abundances and absolute biomass (1a-c). The shallow waters of the Great Barrier Reef region (3) are dominated by fast-growing nano-sized diatoms. The deeper waters of the Indian Ocean and the Coral Sea are characterised by a tropical oceanic flora (2a and 2c, respectively) that is dominated by dinoflagellates and follows the Leeuwin Current (2b) and the East Australia Current and its eddies (2d). South-eastern coastal waters harbour a temperate phytoplankton flora (4) with seasonal succession of different diatom and dinoflagellate communities. Waters south of the tropical and temperate phytoplankton provinces are characterised by an oceanic transition flora (5a,b) that communicates to the subantarctic phytoplankton province (6) and is highly variable in extent. The phytoplankton provinces are associated with surface water masses and the zooplankton fauna likely shows a similar pattern (Figure prepared by G.M. Hallegraeff for CSIRO and National Oceans Office).

groups and communities. For example, there is broad agreement between phytoplankton community distributions and water masses (Figure 3). Australian waters are generally nutrient poor (oligotrophic), particularly with respect to nitrate and phosphate because the boundary currents are largely of tropical and subtropical origins and there is little input from terrestrial sources. In general, Australia has a low average annual rainfall and this rainfall is highly variable. Much of the interior is desert and in the west the aridity extends to the coast. Monsoonal rains fall in the tropical north during the wet season (December to March) with cyclones common at this time, but there is little or no rainfall during the rest of the year. Australian soil is generally low in nutrients and this, together with the high variability in rainfall, results in little terrestrial nutrient input into the surrounding sea. The generally oligotrophic status of Australian marine waters contrasts with many mid-latitude productive coastal areas around the world. This distinction is particularly strong on the western coast of Australia where the Leeuwin Current replaces the upwelling systems produced by the highly productive eastern boundary currents characteristic of all other major ocean basins. The impact of changing productivity on marine oligotrophic systems is largely unknown; they may not be as resilient to stress and disturbance, including climate change, as more productive 410



systems that commonly experience considerable interannual variability. Changes in the terrestrial climate also impact Australia’s marine ecosystems to a greater degree than other parts of the world, so it may not be possible to generalise easily from knowledge elsewhere. Aeolian dust input may be an important regulator of coastal primary production. In regions south of Tasmania, where macronutrient concentrations are always high, iron availability influences growth, biomass and composition of phytoplankton (Sedwick et al. 1999, Boyd et al. 2000). In the macronutrient-limited regions more typical of the waters around continental Australia, the atmospheric supply of iron may stimulate nitrogen-fixing phytoplankton, which have a higher iron requirement than other phytoplankton and therefore influence phytoplankton community composition (Jickells et al. 2005). Climate-induced changes in wind or rainfall may thus have disproportionately large consequences for waters around Australia. Climate change will influence physiology, abundance, distribution and phenology of species both directly and indirectly, although impacts will usually become most apparent at an ecosystem level. Given the intrinsic complexity of ecosystems and the uncertainties in future climate projections, predicting consequences for biodiversity is difficult and highly speculative. Response rates will depend on the magnitude of changes and on longevity of the species involved in a particular system. Plankton systems will therefore respond quickly (Hays et al. 2005), whereas a lag might generally be expected in responses of long-lived species. The ability for adaptation to change will also vary among species but the rapid rate of present climate change coupled with high exploitation and destruction or alteration of habitats will compromise the resilience of many populations and ecosystems (Travis 2002). Strategies for adaptation and mitigation of climate change impacts must begin with the identification of ecosystems or populations that are most vulnerable to change and those most vulnerable to other anthropogenic stressors. In this review, we address the potential impacts of climate variability and climate change on Australian marine life from the intertidal zone through pelagic waters and into the deep sea. We provide a synopsis of climate change projections for Australia of key climate variables known to regulate marine ecosystems from the only IPCC (Intergovernmental Panel of Climate Change) climate system model constructed in the Southern Hemisphere, the Commonwealth Scientific and Industrial Research Organisation (CSIRO) Mk3.5 model. Our focus is on the critical variables that regulate processes in marine ecosystems, namely, temperature, winds, currents, solar radiation, mixed-layer depth and stratification, pH and calcium carbonate saturation state, storms and precipitation, and sea level. We review the expected impacts on species and communities of changes in each of these variables based on laboratory, modelling and field work and concentrate on biological groups found in three broad ecosystems: coastal, pelagic and offshore benthic.

Australian marine biodiversity Australia has highly diverse and unique marine flora and fauna, ranging from spectacular coral reefs in the tropics to giant kelp forests in Tasmanian waters. The biodiversity of tropical Australia is high because it is a continuation of the Indo-Pacific biodiversity hot spot, but much of this fauna is threatened by overharvesting and unregulated development in this region including countries to the north of Australia. The species diversity of seagrasses and mangroves is among the world’s highest, particularly in tropical Australia (Walker & Prince 1987, Kirkman 1997, Walker et al. 1999). Temperate Australian waters contain high numbers of endemic organisms due to their long history of geographic isolation from other temperate regions (Poore 2001). Australian waters also harbour species and ecosystems that are of international importance. The best-known example is the Great Barrier Reef, which is the world’s largest World Heritage Area and extends some 2100 km along the coast of northeast Australia.

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Although Australian temperate waters have lower species diversity than the northern tropical waters, they harbour much higher numbers of endemic species (Poore 2001). Approximately 85% of fish species, 90% of echinoderm species and 95% of mollusc species in these southern waters are endemic (Poore 2001). This high endemism is also documented in Australia’s temperate macroalgae (Bolton 1996, Phillips 2001). High endemism along the southern coastline is partly the result of low dispersal abilities of species and the presence of ecological barriers to dispersal along the southern coastal waters such as a sharp temperature gradient near the cessation of the Leeuwin Current and the absence of near-shore rocky reefs in the centre of the Great Australian Bight and at other locations along the southern Australian coastline. Australia’s fish fauna is extremely diverse and endemic by world standards due to a high diversity of tropical and temperate habitats and due to the geographic isolation of the temperate regions. Pelagic fish found around Australia include iconic species such as tuna, billfish (swordfish and marlin) and sharks. The continental shelf waters off southern Queensland have been identified as a biodiversity hot-spot for large pelagic fishes (Worm et al. 2003). In contrast to the pattern elsewhere, this Australian pelagic fish hot spot is located in an area of high catch rates and fishing effort (Campbell & Hobday 2003). Valuable fisheries exist, despite the generally low productivity of Australian marine waters; these include the Northern Prawn Fishery, the Southern Bluefin Tuna Fishery, the Eastern Tuna and Billfish Fishery and the Western Rock Lobster Fishery. Small pelagic species, such as sardines, jack mackerel, redbait and squid are captured in lower-value but highvolume coastal fisheries operating from a number of Australian ports. For many of these, there are well-known correlations between environmental factors and the productivity of the fishery. For example, the size of the Western Rock Lobster Panulirus cygnus Fishery, which is Australia’s most important single-species fishery and the world’s largest rock lobster fishery, varies in a predictable manner with the strength of the Leeuwin Current (Caputi et al. 2001). Similarly, size of banana prawn Penaeus merguiensis catches in some areas of northern Australia is correlated with wet season rainfall (Staples et al. 1982, Vance et al. 1985). These variables are likely to change as climate changes. Further offshore, cold-water corals are found on seamounts and the continental rise, particularly within the Tasmanian Seamounts Marine Reserve. Cold-water corals are hot spots for biodiversity, comparable to shallow tropical coral reefs, although little is known of their ecology, population dynamics or distribution in Australian waters. Over 850 macro- and megafaunal species were recently found on seamounts in the Tasman and southeast Coral Seas, of which 29–34% were potential endemics or new to science (Richer de Forges et al. 2000, Williams et al. 2006). Globally significant populations of many other groups occur in Australia including populations of marine turtles, marine mammals and seabirds. Six of the seven living species of marine turtle forage and breed in Australian tropical waters. Marine turtles home to their natal area to breed and large rookeries used by tens to hundreds of thousands of turtles occur along the northern Australian coastline and the southern Great Barrier Reef area (Marsh et al. 2001). The flatback turtle Natator depressus nest only on Australian beaches so can be considered endemic to Australia. The dugong Dugong dugon forages on seagrasses in tropical Australasian waters. This species is highly threatened in much of its range and a large proportion of global dugong stock is believed to be in Moreton Bay in eastern Australia and Shark Bay in Western Australia (Marsh et al. 2001). Australian fur seals Arctocephalus pusillus doriferus, the world’s fourth rarest seal species, and the endemic Australian sea lion Neophoca cinerea, one of the most endangered pinnipeds in the world, breed at sites along the southern coast of Australia. These non-migratory pinniped species remain in southern Australian waters for their entire lives. Around 45 species of whales, dolphins and porpoises are found in Australian waters including large baleen whales such as the southern right whale Eubalaena australis and the humpback whale Megaptera novaeangliae, which migrate from their Southern Ocean feeding grounds to temperate waters around the southern parts of Africa, South America and Australia and to the tropical waters of the Pacific to breed. 412



A diverse seabird fauna breeds on mainland and island coastlines around Australia; for example the Houtman Abrolhos Islands on the west coast are an important nesting area for Australian seabirds in terms of biomass and species diversity (Ross et al. 2001). One of the largest documented colonies of crested terns Sterna bergii globally (13,000–15,000 nesting pairs) occurs in the Gulf of Carpentaria in Australia’s tropical north (Walker 1992). Planktivorous seabirds occur in high numbers in Australia’s southern temperate waters. For example an estimated 23 million short-tailed shearwaters Puffinus tenuirostris nest in southeast Australia (Ross et al. 2001).

Climate change projections for Australia A number of climate models have been used to investigate the response of the ocean-atmosphere system to increased levels of greenhouse gases and aerosols (Cubasch et al. 2001). This review examines aspects of climate simulations that are relevant to determining how marine ecosystems will respond to global climate change. In general, climate model simulations using future greenhouse gas emission scenarios project oceanic warming, an increase in oceanic stratification and alteration of mixing depth, changes in circulation, increased pH and rise in sea level, alterations in cloud cover and ozone levels and thus solar radiation reaching the ocean surface and altered storm and rainfall regimes (Figure 4). It is very likely that such changes will cause considerable alterations in marine biological communities (Bopp et al. 2001, Boyd & Doney 2002, Sarmiento et al. 2004). We use future climate projections over the next century from the CSIRO Mk3.5 climate model (hereafter called the CSIRO climate model; Appendix 1) using the IS92a future emissions scenario, often referred to as the ‘business-as-usual’ scenario. Although there are subtle differences between the CSIRO climate model and other international models, many of the general trends in these fields are similar and we use the CSIRO climate model to suggest the magnitude of the projected changes in the set of variables that follow. HUMAN ACTIVITIES Increased greenhouse gas concentration

Change in UV radiation levels

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ff Rise in sea-level Altered nutrient supply and stratification (mixed layer depth)

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Figure 4 Important physical and chemical changes in the atmosphere and oceans as a result of climate change.

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Ocean temperature Waters around Australia are projected to warm by 1–2°C by the 2030s and 2–3°C by the 2070s (Figure 5). The CSIRO climate model projects the greatest warming off southeast Australia and this is the area of greatest warming this century in the entire Southern Hemisphere. This Tasman Sea warming is associated with systematic changes in the surface currents on the east coast of Australia; including a strengthening of the EAC and increased southward flow as far south as Tasmania (Figure 5). This feature is present in all IPCC climate model simulations, with only the magnitude of the change differing among models. Changes in currents leading to the Tasman Sea warming observed to date is driven by a southward migration of the high-latitude westerly wind belt south of Australia, and this is expected to continue in the future (Cai et al. 2005, Cai 2006).

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Winds Under global warming scenarios, the southeasterly trade winds strengthen east of northern Australia, but weaken to the west of the continent (Figure 5). Westerly winds in southern Australian waters will weaken. In the Australian coastal region, downwelling will prevail due to the dominating winds and density structure of the upper ocean. Increasing wind intensity may suppress localised upwelling in the northeast. However, decreasing wind intensity in southern waters may facilitate localised upwelling there.

Ocean currents Surface currents on the east coast will show a systematic change (Figure 5) including EAC strengthening and increased southward flow as far south as Tasmania. On the west coast there will be no obvious strengthening of the Leeuwin Current. In the south, the Great Australian Bight region will experience more westward transport as global temperatures rise. Along the northwest and northeast coasts there will be an increase in the northward flow.

Mixed-layer depth and stratification The Australian coastal region is generally a downwelling region due to prevailing winds and density structure of the ocean. In oligotrophic marine regions of Australia, the dominant mechanism of nutrient supply to the upper ocean is winter convective mixing due to cooling of surface waters. Under these conditions the seasonal evolution of the mixed-layer depth and density differences between this layer and the water below play an important role in the supply of nutrients to the upper ocean. Surface ocean warming will stabilise the upper ocean and reduce the supply of nutrients to the surface. The CSIRO climate model simulations project a decline in the annual mean mixedlayer depth by the 2070s (Figure 5).

CO2, pH and calcium carbonate saturation state Over the last 200 years, oceans have absorbed 40–50% of the anthropogenic CO2 released into the atmosphere (Raven et al. 2005). Rising atmospheric CO2 concentrations via fossil fuel emissions will lead to enhanced oceanic CO2 as the ocean re-equilibrates with the perturbed atmosphere (McNeil et al. 2003). Elevated CO2 in the upper ocean will alter the chemical speciation of the oceanic carbon system. As CO2 enters the ocean it undergoes the following equilibrium reactions: CO 2 + H 2O ⇔ H 2CO3 ⇔ HCO3− + H + ⇔ CO32− + 2H + Two important parameters of the oceanic carbon system are the pH and the calcium carbonate (CaCO3) saturation state of sea water (Ω). Ω expresses the stability of the two different forms of CaCO3 (calcite and aragonite) in sea water. Increasing CO2 concentration in the surface ocean via uptake of anthropogenic CO2 will have two effects. First, it decreases the surface ocean carbonate ion concentration (CO32−) and decreases Ω. Using an ocean-only model forced with atmospheric CO2 projections (IS92a), Kleypas et al. (1999) predicted a 40% reduction in aragonite saturation (Ωarag) by 2100. Laboratory experiments

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have shown that some species of corals and calcifying plankton (Gattuso et al. 1998, Langdon et al. 2000, Orr et al. 2005) are highly sensitive to changes in Ω, which has led to the hypothesis of large decreases in future calcification rates under elevated atmospheric CO2 (Kleypas et al. 1999). Second, when CO2 dissolves in water it forms a weak acid (H2CO3) that dissociates to bicarbonate, generating hydrogen ions (H+), which makes the ocean more acidic (pH decreases). Using an ocean-only model forced with atmospheric CO2 projections (IS92a), Caldeira & Wickett (2003) predicted a pH drop of 0.4 units by the year 2100 and a further decline of 0.7 by the year 2300. They argued that the oceanic absorption of anthropogenic CO2 over the next several centuries may result in a pH decrease greater than inferred from the geological record over the past 300 million years, with the possible exception of those resulting from rare, extreme events such as meteor impacts. Changes in surface pH and in Ωarag reflect changes in the speciation of carbon within the ocean and are a function of temperature, salinity, alkalinity and dissolved inorganic carbon concentrations. McNeil & Matear (2006) showed that climate change does not alter the projected change in surface pH. The projected pH decrease is controlled by the future levels of atmospheric CO2. However, the decline in Ωarag due to rising CO2 levels in the ocean is slightly reduced (~15%) because of the increase in Ωarag due to the increase in surface temperature. For the Australian region, the pH and Ωarag for the 1990s are shown along with the corresponding change in these values relative to 1990s (Figure 6). We see significant declines in these parameters but with the greatest declines occurring off northeast Australia. A major unknown in this region is whether any dissolution of the tropical coral reefs would buffer the pH decreases. Because of the enhanced levels of CO2 in the atmosphere and rates of fossil fuel burning, the process of ocean acidification is essentially irreversible over the next century. It will take thousands of years for ocean chemistry to return to a condition similar to that of preindustrial times.

Solar radiation Highly energetic ultraviolet radiation (UVR) penetrates the ocean surface and is known to have detrimental effects on marine organisms. UVR penetration to the earth’s surface increased during the last quarter of the twentieth century as stratospheric ozone was depleted by chlorofluorocarbons (CFCs), halons, hydrochlorofluorocarbons and other compounds. Stratospheric ozone levels appear to have stabilised, however, due to the 1989 implementation of the Montreal Protocol designed to phase out the production of CFCs and other compounds that deplete the ozone layer (de Jager et al. 2005). Most climate models predict that the ozone layer will recover and thicken throughout the twenty-first century (de Jager et al. 2005), so UVR penetration should decline (McKenzie et al. 2003). However, these predictions are somewhat uncertain, especially in the timing of the rethickening, due to uncertainties in projections of greenhouse gas emissions and degradation and due to the complex ways that chemical, radiative and dynamic processes will affect stratospheric ozone. For example, chemical reactions of some greenhouse gases (such as methane) can reduce total ozone in the stratosphere but the level of methane emissions is difficult to predict. Climate change will also affect UVR penetration indirectly by influencing other factors such as aerosols, clouds and snow cover. Aerosols can scatter more than 50% of the UV-B — the biologically important component of UVR — and aerosols increased in the atmosphere during most of the twentieth century, although they have shown declines since 1990 (Schiermeier 2005). Clouds can attenuate 15–30% of the UV-B, and cloud reflectance measured by satellite has shown a long-term increase in some regions of the world (McKenzie et al. 2003). All these factors introduce considerable uncertainty in future levels of UVR at the ocean surface, and it has been suggested that climate warming will slow the recovery of the ozone layer by up to 20 yr (Kelfkens et al. 2002). 416



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Precipitation and storms Changes in the amount or timing of rainfall and the associated river runoff affect the salinity regimes of estuaries and adjacent coastal waters, while in comparison salinity is relatively constant throughout the year in most oceanic waters. Despite the high uncertainty of rainfall projections in Australia, there is a tendency for decreased rainfall over most of Australia and over the oceans in climate model simulations (Figure 7). This general reduction in rainfall may be offset by an increase in the frequency of intense storms (Emanuel 2005, Webster et al. 2005), which will increase rainfall intensity and the associated runoff of freshwater and suspended sediments. In northern Australia, tropical cyclones are important extreme rainfall events. A recent study under 3 times the baseline levels of CO2 conditions based on levels prior to the industrial revolution in the mid-1800s, projected a 56% increase in the number of simulated tropical cyclones over northeastern Australia with peak winds greater than 30 ms−1 (Walsh et al. 2004). However, the behaviour of tropical cyclones under 417



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global warming is uncertain because they are not currently well resolved by global or regional climate models (Pittock et al. 1996, Walsh & Pittock 1998).

Sea level Rising sea level around Australia will flood existing coastal environments and alter their marine habitats. With global warming, the CSIRO climate model projects a doubling in the rate of sealevel rise from the observed 1.44 mm yr−1 for the twentieth century (Church et al. 2001). By the 2080s, sea level is projected to rise by 0.06–0.74 m above the 1990 value (Gregory et al. 2001). These projections take into account both the mean global projections from the IPCC scenarios and the non-uniform spatial distributions of sea-level change related to thermal expansion produced by the climate simulations. However, they do not include vertical land movement, which can be locally important. Sea-level rise projected by the CSIRO model for just the thermal expansion shows an increase in the entire Australian region but with large spatial variability (Figure 7). The variability in sea-level rise reflects how the excess heating of the planet due to global warming is stored in 418



the oceans, and this large variability is supported by reconstructed sea-level estimates from the past decade (Willis et al. 2003). Therefore, over this century the local impact of sea-level rise may substantially deviate from the global averaged value. For the Australian region, much greater sealevel rise is projected on the east coast than the west coast due to the increased southward penetration of the warm EAC, which causes water here to expand more than in other regions.

Climate impacts on Australian marine life In this section we describe the impacts of climate variables on marine life in coastal, pelagic and offshore benthic systems. We consider the climate variables that have greatest impact on structuring marine communities within these systems and for which projections over the next 100 yr are available from global climate models. Where applicable, we review impacts on physiology, distributions and abundance, and phenology of marine organisms. Studies of climate impacts from both field and experimental research from Australia are discussed and supplemented with studies and observations from international research. Results of this section are summarised in Table 1.

Ocean temperature Elevated water temperatures stress plants and animals already near the upper limits of their optimal temperature range, slowing growth and impairing reproductive capacity (Philippart et al. 2003, Roessig et al. 2004, Helmuth et al. 2005, Keser et al. 2005). This is because most biological processes have an optimal temperature range and outside this range physiological efficiency declines. Coastal systems Physiology Extreme temperatures, both warm and cool, if severe or prolonged can lead to irreparable damage and death of coastal organisms as well as photosynthetic inhibition in marine plants (Bruhn & Gerard 1996, Ralph 1998, Davenport & Davenport 2005, Campbell et al. 2006). Large diebacks of marine fauna and flora in the intertidal and shallow subtidal occur on very hot days particularly when these coincide with low tides during the middle of the day (Tsuchiya 1983, Perez et al. 2000). Such a situation may have been responsible for the major dieback of seagrass beds in southern Australia during early 1993 when over 12,000 hectares were lost (Seddon et al. 2000). Probably the most widely publicised mass mortalities induced by warmer-than-average temperatures are those resulting from tropical coral reef bleaching events (Hoegh-Guldberg 1999). During bleaching events, the symbiosis between the coral and the unicellular algae (dineflagellates from the genus Symbiodium) that live within the coral tissues disintegrates. Bleached corals may recover their symbiotic populations of Symbiodium in the weeks and months after a bleaching event if the conditions triggering the event are mild and short-lived, but mortality has reached 100% in bleached corals when stressful conditions have persisted for days to weeks. Recent warming throughout tropical oceans has led to repeated coral bleaching events, not seen anywhere in the world before 1979, affecting hundreds to thousands of square kilometres of coral reefs in almost every region of the world where coral reefs occur. In the most severe global episode of mass coral bleaching (1998), 16% of corals that were surveyed before that event had died by the end of the year (Hoegh-Guldberg 1999, Knowlton 2001). Mass bleaching events over large sections of the Great Barrier Reef have occurred six times during the past 30 years: in 1983, 1987, 1991, 1998, 2002 and 2006. Mortality rates in this region were relatively low however, primarily because warming on the Great Barrier Reef was less severe than in other parts of Australia and the world. For example, in 1998 a very warm pool of water sat 419



Table 1 Expected and observed impacts of climate change on Australian marine life and field or experimental evidence from outside Australia

Expected change in climate

Species group/ natural system

Expected climate impact in Australia

Observations in Australia

Increasing temperature

Seagrasses and mangroves

Poleward shift in species ranges and a shift in abundance toward species tolerant of warmer waters Earlier flowering and fruiting

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Increased frequency and intensity of large-scale diebacks with increase in frequency and intensity of extreme temperatures Poleward shift in species ranges and a shift in abundance toward species tolerant of warmer waters

Seagrass distributional limits linked to temperature1 Flowering of seagrasses in temperate Australia linked to water temperature2 Southern Australia early 1993 (>12,000 hectares)3

Rocky shore, fauna and macroalgae

Kelp communities

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Zooplankton

Increased frequency and intensity of large-scale diebacks with increase in frequency and intensity of extreme temperatures Contraction of kelp ranges, declines in abundance, local extinctions, particularly in Tasmania Poleward shift in species ranges and a shift in abundance toward warm-water species

A decline where warming enhances stratification Earlier appearance of plankton in summer in temperate waters Increase in frequency and intensity of harmful and nuisance blooms Poleward shift in species ranges and a shift in abundance toward warm-water species A decline where warming enhances stratification

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Diebacks in Tasmania and South Australian hot days5

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Rocky shores in Europe, United States and South America over past 50 yr4 European and Japanese coasts6

Decline of kelp in Tasmanian waters over past 50 yr7

Loss of kelp in east Pacific following El Niño8

Southward extension of a coccolithophore and a dinoflagellate in southeast Australia9

Poleward shift in North Atlantic10

North Atlantic11 North Sea12 Norwegian coast13

Large poleward range shifts (>1000 km) in North Atlantic14 North Atlantic15



Table 1 (continued) Expected and observed impacts of climate change on Australian marine life and field or experimental evidence from outside Australia



Coral reefs

Demersal and pelagic fish

Seabirds and wetland birds

Marine turtles and mammals

Expected climate impact in Australia Earlier appearance of zooplankton in summer in temperate waters Increase in frequency and severity of coral bleaching and mortality Increase in local extinctions of coral-associated fauna with bleaching events Poleward shift in species ranges and a shift in abundance toward species tolerant of warmer waters

Earlier dates of mean migration and spawning in temperate and subtropical species Poleward shifts in species ranges and a shift in abundance toward species tolerant of warmer waters Earlier arrival in migratory species in temperate and subtropical regions Earlier nesting and laying and protracted breeding seasons in temperate and subtropical species Poleward shift in species foraging ranges

Earlier breeding

Skewing of turtle sex ratios toward females

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North Sea16

Six severe bleaching events in past 30 yr (Great Barrier Reef, Ningaloo Reef)17

Coral reefs globally18

Coral reefs globally19 Tasmanian fish distributions shifting south with increase in fish that prefer warmer waters20

North Atlantic fish shifting northward21

Earlier migrations in northeast Atlantic fish22 Southward shift of seabird distributions in Western Australia and increase in abundance23 Southern Australian wetland birds24

Terrestrial, wetland and seabirds globally25

Western and southern Australian seabirds26 Northward shift of cetaceans and turtles in northeast Atlantic27 Earlier nesting in marine turtles in United States28 Experimental and modelling evidence that warmer temperatures produce more females29 (continued on next page)








Alteration of winds

Phyto- and zooplankton

Increased productivity where wind mixing is enhanced and a reduction where wind strength declines

Coastal fish

Recruitment strength linked to wind strength Reduction of breeding success with prolonged periods of strong winds Local extinctions of cold-water species in southeastern Australia with increased flow of EAC, appearance of tropical species further south on east coast

Production pulses correlated with peaks in wind oscillation in Tasmanian shelf waters30 Rocky reef fish32

Seabirds

Alteration of currents including strengthening of EAC

Seagrasses & mangroves

Rocky shore, fauna and macroalgae

Kelp communities


Local extinctions of cold-water species in southeastern Australia with increased flow of EAC, appearance of tropical species further south on east coast Local extinctions of cold-water species in southeastern Australia with increased flow of EAC, appearance of tropical species further south on east coast Poleward extension of warm currents will transport tropical plankton more southward

Decline in mixed-layer depth/increasing stratification


Decrease in abundance

Increased CO2 and decrease in pH and aragonite saturation state

Mangroves

Increase in productivity with rising atmospheric CO2

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Observations elsewhere or experimental evidence Decreased production in central North Pacific during low-wind regimes31

Breeding colonies on Great Barrier Reef33 Seagrass distributional limits further south on west coast than east coast due to influence of warmwater Leeuwin Current34 Tropical species already found at temperate latitudes on east coast35

Expansion of longspined urchin to Tasmania facilitated by larval transport by EAC36 High abundance of a tropical coccolithophore off southeast Australia37 Phytoplankton productivity in central North Pacific declines as mixed-layer depth decreases38 Experimental evidence39







Seagrasses

Increase in productivity with increase dissolved CO2 and deepening of depth limits Impaired growth in calcifying fauna and macroalgae and increase in mortality of early life stages Changes in growth and community composition; longterm decline in abundance and distribution of calcifying species Impaired growth in calcifying species, particularly pteropods; midterm decline in abundance and distribution Impaired growth rates and possible dissolution

Rocky shore, fauna and macroalgae Phytoplankton

Zooplankton

Coral reefs

Possible increase in UV

Cold-water corals Seagrasses

Mangroves

Rocky shore fauna and macroalgae Kelp and subtidal macroalgae Phytoplankton

Zooplankton

Coral reefs

High threat of impaired growth rates and possible dissolution Reduction of growth rates and biomass in UV-sensitive species Reduction of growth rates and biomass in UV-sensitive species Increase mortality of early life stages and reduction of growth rates in UV-sensitive species Increase mortality of early life stages Reduction of growth rates and biomass in UV-sensitive species and of nutritional value to zooplankton Changes in community composition Increased mortality of early life stages and reduction of growth rates in UV-sensitive species Increase in mortality during bleaching events through synergistic effects with temperature

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Observations elsewhere or experimental evidence Experimental evidence40 Experimental evidence41

Experimental evidence42

Experimental evidence43

Experimental and modelling evidence44 Evidence from modelling work45 Experimental evidence46 Experimental evidence47 Experimental evidence48 Experimental evidence49 Evidence from field and laboratory experiments50

Evidence from laboratory experiments51 Evidence from laboratory experiments52 (continued on next page)






Demersal and pelagic fish

Increase in frequency or intensity of severe storms and extreme rainfall events and a decrease in average rainfall

Mangroves

Expected climate impact in Australia Increase mortality of early life stages and reduction of growth rates Damage to epidermis and ocular components in pelagic species and increased mortality in egg and larval stages in shallow water and upper ocean Shifts in community abundance as coastal salinity regimes are altered and nutrient and sediment loading changes

Seagrasses

Destruction of seagrass beds

Kelp communities and subtidal macroalgae

Shifts in community abundance and increased local mass mortality events associated with storms and flood events

Benthic macrofauna

Shifts in community abundance and increased local mass mortality events associated with storms and flood events

Alteration of peak timing of life cycle events

Coral reefs

Mass mortality events associated with storms and flood events

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Observations elsewhere or experimental evidence Evidence from laboratory experiments53 Evidence from laboratory experiments54

Increase in mangrove area in southeast Australia may be indirectly linked to changes in rainfall although changes in land use likely to be overriding factor55 Loss of >1000 km2 in Harvey Bay after severe storms and flooding56 Switch from canopyforming macroalgae to turf-forming algae in South Australia linked to enhanced nutrient supply from coastal runoff58 Mass mortality of grazing urchins after freshwater pulse60

High rainfall may decrease salinity in estuaries so triggering prawn emigration in northern Australia62 Mass mortality of corals on Great Barrier Reef after cyclones and flood events64

Large-scale destruction in United States after cyclones57 Range shifts of macroalgae in New Zealand and California associated with storms and wave exposure59 Field experiments revealed shift in community composition with increased sedimentation61 High rainfall may decrease salinity in estuaries so triggering prawn emigration in the United States63 Mass mortality of corals in Caribbean after cyclones65






Phytoplankton


Rise in sea level

Mangroves

Seagrass

Seabirds




Community structure influenced by rainfall regime and runoff

Lower coral diversity on Great Barrier Reef in wet tropics66

Diatoms may decline with decreasing average runoff and nutrient input while dinoflagellates (including harmful algae) may profit from stormassociated runoff and humic substances in coastal waters Increased mortality events

Alteration of hydrological or tidal regimes leads to mortality of mangroves Mangrove retreat with rising sea level Reduction in growth of seagrass and distributional shifts

Loss of breeding sites for species that nest on low-lying coastal areas through increased flooding and erosion Loss of breeding and haul-out sites for species through increased flooding and erosion


Evidence from field experiment and time series67

High mortalities of turtles and seal pups associated with cyclones and storms68 Mangroves in Africa and Asia69 Caribbean70 50 cm rise in sea level expected to result in 30–40% reduction of seagrass growth71 Evidence from modelling work72

50 cm rise in sea level expected to lead to a 32% loss of turtle nesting beaches in the Caribbean73

Notes: 1Walker & Prince 1987; 2West & Larkum 1979, Cambridge & Hocking 1997, Inglis & Smith 1998; 3Seddon et al. 2000; 4Barry et al. 1995, Southward et al. 1995, Sagarin et al. 1999, Zacherl et al. 2003, Mieszkowska et al. 2005, Rivadeneira & Fernandez 2005, Simkanin et al. 2005, Smith et al. 2006; 5Valentine & Johnson 2004, Womersley & Edwards 1958; 6Tsuchiya 1983, Perez et al. 2000; 7Edyvane 2003, Edgar et al. 2005; 8Dayton & Tegner 1984, Zimmerman & Robertson 1985, Dayton et al. 1998, 1999, Adey & Steneck 2001; 9Blackburn & Creswell 1993, Blackburn 2005, G. Hallegraef pers. com.; 10M. Edwards 2005; 11Richardson & Schoeman 2004; 12Edwards & Richardson 2004; 13Edwards et al. 2006; 14Beaugrand et al. 2002, Bonnet et al. 2005; 15Richardson & Schoeman 2004; 16Greve et al. 2004, Edwards & Richardson 2004, Kirby et al. 2007; 17Hoegh-Guldberg 1999, Wilkinson 2004; 18Hoegh-Guldberg 1999, Knowlton 2001; 19Dulvy et al. 2003; 20Welsford & Lyle 2003, P. Last pers. com.; 21Beare et al. 2004, Byrkjedal et al. 2004, Perry et al. 2005, (continued on next page)

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Table 1 (continued) Expected and observed impacts of climate change on Australian marine life and field or experimental evidence from outside Australia Notes (continued): Rose 2005a, 2005b; 22Sims et al. 2001; 23Dunlop & Wooller 1986, Dunlop et al. 2001, Bancroft et al. 2004; 24Beaumont et al. 2006; 25Mason 1995, Crick et al. 1997, Archaux 2003, Both et al. 2004, Lehikoinen et al. 2004, Both et al. 2005, Marra et al. 2005, Jonzén et al. 2006, Moller et al. 2006; 26Dunlop & Wooller 1986, Chambers 2004; 27Robinson et al. 2005, MacLeod et al. 2005, McMahon & Hays 2006; 28Weishampel et al. 2004; 29Yntema & Mrosovsky 1982, Godfrey et al. 1999, Booth & Astill 2001, Glen & Mrosovsky 2004; 30Harris et al. 1991; 31Polovina et al. 1994; 32Thresher et al. 1989; 33King et al. 1992; 34Walker & Prince 1987; 35Griffiths 2003; 36Johnson et al. 2005; 37Blackburn & Cresswell 1993, Blackburn 2005; 38Venrick et al. 1987, Polovina et al. 1994, 1995; 39Polovina et al. 1995, Roemmich & McGowan 1995, Farnsworth et al. 1996, Ainsworth & Long 2005; 40Invers et al. 1997, 2002, Zimmerman et al. 1997; 41Gao et al. 1993, Kurihara et al. 2004, Michaelidis et al. 2005, Berge et al. 2006; 42Riebesell et al. 2000, Antia et al. 2001, Tortell et al. 2002, Engel et al. 2005; 43Orr et al. 2005; 44See Hoegh-Guldberg 2004; 45Guinotte et al. 2006, Raven et al. 2005; 46Dawson & Dennison 1996; 47Moorthy & Kathiresan 1997, 1998; 48Graham 1996, Rijstenbil et al. 2000, Cordi et al. 2001, Lesser et al. 2003, Przeslawski et al. 2004, 2005, Bonaventura et al. 2006; 49Graham 1996, Bischof et al. 1998, Swanson & Druehl 2000, Wiencke et al. 2006; 50Behrenfeld et al. 1993, Keller et al. 1997, Wilhelm et al. 1997, Wängberg et al. 1999, Garde & Cailliau 2000, Barbieri et al. 2002, Litchman & Neale 2005; 51Karanas et al. 1979, Damkaer & Dey 1983; 52Lesser 1996, 1997, Baruch et al. 2005, Drohan et al. 2005; 53Shick et al. 1996, Wellington & Fitt 2003; 54Hunter et al. 1982, Keller et al. 1997, Zagarese & Williamson 2001, Markkula et al. 2005; 55Saintilan & Williams 1999, Harty 2004, Rogers et al. 2006; 56Preen et al. 1995; 57Thomas et al. 1961; 58Gorgula & Connell 2004; 59Graham 1997, Cole et al. 2001; 60Andrew 1991; 61Norkko et al. 2002, Thrush et al. 2003a, 2003b, Lohrer et al. 2004; 62Staples 1980, Vance et al. 1985, Staples & Vance 1986, Vance et al. 1998; 63Zein-Eldin & Renaud 1986; 64Alongi & Robertson 1995, Alongi & MacKinnon 2005; 65Porter & Meier 1992, Gardner et al. 2005; 66De Vantier et al. 2006; 67Carlsson et al. 1995, Goffart et al. 2002; 68Limpus & Reed 1985, Pemberton & Gale 2004; 69Blasco et al. 1996; 70Ellison 1993, Parkinson et al. 1994; 71Short & Neckles 1999; 72Galbraith et al. 2002, Smart & Gill 2003; 73Fish et al. 2005.

above Scott Reef off northwest Australia for several months, resulting in an almost total bleaching of these offshore reefs and mortality of corals down to 30 m depth. The recovery of Scott Reef has been very slow (Wilkinson 2004). By the middle of this century, temperature thresholds for coral bleaching will be exceeded every year in Australia if sea temperatures increase as projected by global climate models (HoeghGuldberg 1999). Based on the current responses of corals, it is estimated that an increase of 2°C in tropical and subtropical Australia would result in annual bleaching and quite possibly regular, large-scale mortalities (Hoegh-Guldberg 1999, 2004, Lough 2000). A geographic analysis of risk to the Great Barrier Reef associated with these changes in sea temperature indicated that the projected succession of devastating mass coral bleaching events will severely compromise the ability of reefs to recover, no matter where they are found along the Queensland coastline (Done et al. 2003). This analysis indicated that deterioration of coral populations is likely in most of the scenarios examined and this is reinforced by findings from other studies (Hoegh-Guldberg 1999, Donner et al. 2005). For large, mobile animals that may be transient visitors to coastal waters, oceanic warming may impact particular life stages such as juveniles or embryos. For example, gender in all turtles is determined by ambient nest temperatures during embryonic development (Mrosovsky et al. 1992, Godfrey et al. 1999, Hewavisenthi & Parmenter 2002a). Small changes in temperature close to the pivotal temperature at which a 50:50 sex ratio is produced (~29°C for marine turtles) skew the sex ratio of hatchlings, with warmer temperatures producing more females (Yntema & Mrosovsky 1982, Godfrey et al. 1999, Booth & Astill 2001, Glen & Mrosovsky 2004). Many nesting beaches around the world, including most Australian beaches, already have a strong female bias (Limpus

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1992, Loop et al. 1995, Godfrey et al. 1996, Binckley et al. 1998, Hewavisenthi & Parmenter 2002b, Hays et al. 2003, Glen & Mrosovsky 2004) so if temperatures rise, the proportion of eggs developing as males may be further reduced. However, light-coloured (thus cooler) beaches within nesting regions produce more males (Hays et al. 2003). In Queensland beaches on offshore coral cays and islands have lighter-coloured sand than mainland beaches, thus maintaining sex ratios (Environment Australia 1998). Therefore if temperatures warm on these beaches, the gross skewing in sex bias may have serious implications for local breeding population persistence. On a global scale outbreaks of disease have increased over the last three decades in many marine groups including corals, echinoderms, mammals, molluscs and turtles (Ward & Lafferty 2004). Causes for increases in diseases of many groups remain uncertain, although temperature is one factor that has been implicated in corals, molluscs and turtles (Harvell et al. 2002). Previously unseen diseases have also emerged in new areas through shifts in distribution of hosts or pathogens, many of these shifts are in response to climate change (Harvell et al. 1999). A consequence of climate-mediated physiological stress is that host resistance to pathogens or parasites can be compromised (Scheibling & Hennigar 1997, Garrabou et al. 2001, Lee et al. 2001, Harvell et al. 2002, Mouritsen et al. 2005). Temperature-induced disease outbreaks in corals on the Great Barrier Reef have occurred at the same time as bleaching events, resulting in increased coral mortality rates (Jones et al. 2004). A large-scale mortality of greenlip abalone, Haliotis laevigata, along the south Australian coast in 1985 and 1986 due to infection by Perkinsus parasites may have been aggravated by warmer water temperatures predisposing the abalone to this disease (Goggin & Lester 1995). Population declines due to temperature-related disease susceptibility have also been reported in several Californian abalone species through both observational and experimental studies (Davis et al. 1996, Vilchis et al. 2005). Fibropapillomatosis, a disease that causes tumours, is now common in green turtles Chelonia mydas and olive ridley turtles Lepidochelys olivacea (Adnyana et al. 1997, Jones 2004). This disease was first documented in the 1930s and was rare until the early 1980s but has since reached epidemic proportions in many turtle populations worldwide (Jones 2004). The prevalence of the tumours in young turtles suggests prolonged exposure to anthropogenic pollutants may be responsible (Adnyana et al. 1997, Herbst et al. 2004, Jones 2004, Ene et al. 2005, Foley et al. 2005). However, the increase of this disease in recent decades coincides with rapidly rising temperatures so it may also be indirectly related to climate change (Robinson et al. 2005). Distribution and abundance Temperature influences the abundance and distribution of coastal marine life such as macroalgae, seagrasses and molluscs (McMillan 1984, Walker & Prince 1987, Jernakoff et al. 1996, Steneck et al. 2002, Hiscock et al. 2004). Fluctuations in species abundances and community composition have been linked to variations in temperature (Southward et al. 1995, Tegner et al. 1996, Dayton et al. 1999, Grove et al. 2002, M.S. Edwards 2004, Schiel et al. 2004, Smith et al. 2006). Shifts in species distributions associated with ocean warming are documented from rocky shores in Europe, the United States and South America (Barry et al. 1995, Sagarin et al. 1999, Zacherl et al. 2003, Mieszkowska et al. 2005, Rivadeneira & Fernandez 2005, Simkanin et al. 2005). For example, a recent comprehensive resurvey of rocky intertidal shores around the United Kingdom found range extensions in the northern (high-latitude) limits of some warm-water species over the past 50 yr and a retraction in the southern limits of fewer cold-water species although rates of recession were not as fast as rates of advancement in warm-water species (Mieszkowska et al. 2005). The high levels of endemism along Australia’s southern coastline could increase vulnerability to temperature increases compared to temperate rocky shores elsewhere; many endemic species may have more stringent temperature limits and so may be particularly susceptible to warming (Beardall et al. 1998).

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There are interactive effects between the impacts of warming and availability of nutrients on distribution and abundance of macroalgae. Declines of giant kelp forest communities in Tasmanian coastal waters have been associated with thermal and nutrient stress (Edyvane 2003, Edgar et al. 2005). Macrocystis kelp forests in Australia are found predominantly in the southeast where water conditions are cool and relatively nutrient rich. There has been a considerable decline in Tasmanian kelp forests over the past 50 yr associated with rising temperatures (Edyvane 2003). Further, an unusual dieback of the shallow sublittoral brown macroalga Phyllospora comosa along the east coast of Tasmania in 2001 has also been attributed to above-average seawater temperatures coupled with nutrient stress (Valentine & Johnson 2004). If the EAC strengthens as projected by climate models, warm, nutrient-poor water will impinge more frequently on Tasmanian giant kelp communities, potentially leading to local extinction and a shift of macroalgal communities to understoreydominated forms (Kennelly 1987a,b, Dayton et al. 1999). Globally, mangrove distribution is generally constrained by the 20°C winter sea isotherm; there are a few exceptions, such as the more southerly distribution of mangroves in eastern Australia (Duke 1992). It has been suggested that this distribution is the result of small-scale extensions of warmer currents, such as the EAC, or that the southern populations are a relict representing refuges of more poleward distributions in the past (Duke 1992). As mangrove species show considerable variation in their sensitivity to temperature, species composition of mangrove forests will alter as temperatures rise and species distributions are expected to shift poleward (Field 1995). Evidence suggests that some benthic and demersal fish species may be able to move as oceans warm, regardless of whether there is a shift in associated habitats such as coral reefs, kelp forests or rocky reef communities. Certain fishes associated with coral reefs appear to be able to populate reefs that do not have corals, as shown by the appearance of coral reef fishes in southern New South Wales and Victoria during the summer (Hoegh-Guldberg 2004). These fishes recruit into coastal areas and grow for several months, disappearing when cold conditions return. Many coral reef fish may be able to move southward as oceans warm, although obligate corallivorous species would presumably be missing (Hoegh-Guldberg 2004). This has already been observed in other parts of the world such as California, where the composition of near-shore rocky reef fish communities shifted in dominance from cold-water northern species to warm-water southern species as temperatures warmed (Holbrook et al. 1997). However, coral bleaching has already led to local extinctions of a few coral-associated fish (Dulvy et al. 2003) and doubtless many more could disappear as coral bleaching episodes increase. Other mobile groups such as seabirds and marine mammals may be able to rapidly shift their distributions with climate change, although many are restricted to coastal habitats during breeding seasons. Warmer waters may allow marine turtles and dugongs to extend their foraging distributions in Australian inshore waters further south. However, green turtles Chelonia mydas and dugongs Dugong dugon selectively feed on seagrasses while hawksbill turtles Eretmochelys imbricata forage on coral reefs, so their ability to shift distributions are likely to be limited by changes in the distribution of their food sources. Range expansions have already been observed in seabird species along the west coast of Australia, with tropical species extending their breeding and foraging ranges southward (Dunlop & Wooller 1986, Dunlop et al. 2001). The recent growth of nesting colonies of wedge-tailed shearwaters Puffinus pacificus in southwestern Australia may be due to a southerly movement from more northerly colonies as temperatures rise (Bancroft et al. 2004). Wedge-tailed shearwaters are found only over waters with surface temperatures exceeding 20°C (Surman & Wooller 2000). The population of Australasian gannets Morus serrator that breed in southeast Australia has increased by approximately 6% per year since 1980, with new breeding sites being established as nesting space becomes limited (Bunce et al. 2002). This increase appears to be associated with a long-term

428



warming trend and a concurrent increase in the abundance of small pelagic prey fish, principally pilchards Sardinops sagax. Phenology Water temperature and day length are the principal triggers or correlates for the timing of biological events such as breeding or migration in marine animals and flowering and seed germination in marine plants (Parmesan & Yohe 2003). Synchrony in reproduction of widely distributed seagrass beds and mangroves (Clarke & Myerscough 1991, Inglis & Smith 1998, DiazAlmela et al. 2006) suggests control by these environmental variables. Such synchronies of biological events in distant populations may be regulated by a large-scale independent factor such as temperature or day length. Regular flowering of the seagrass Posidonia australis occurs between April and June in southwestern Australia, probably induced by a seasonal decline in water temperatures (West & Larkum 1979, Cambridge & Hocking 1997). However, further north in Shark Bay P. australis meadows do not flower every year (Larkum 1976). Widespread flowering P. australis is also rare off central New South Wales on the east coast (Walker et al. 1988). Shark Bay and central New South Wales are near the northern limits for this temperate seagrass species so the threshold decline in water temperature required to trigger flowering may begin to occur less frequently. As a warming of coastal waters is projected, particularly off southeast Australia, episodes of flowering of P. australis may become even rarer in northern meadows. The deposition of seed banks after flowering is an important process that allows seagrass beds to recover rapidly from catastrophic disturbances such as storms or floods (Preen et al. 1995). Temperature has also been correlated with the timing of mass spawning in tropical reef corals on the Great Barrier Reef (Babcock et al. 1986) and on the tropical west coast (Simpson 1991). However, the physiological and evolutionary mechanisms that underlie the timing of reproduction in corals and in most marine invertebrates are far from clear; thus it is difficult to speculate on the consequences of any change in the timing of spawning. There is global evidence that climate change is influencing the phenology of larger marine fauna. Marine turtles in Florida in the United States are nesting earlier in response to warmer ocean temperatures (Weishampel et al. 2004). Warmer waters also reduce the interval length between the multiple clutches laid within a nesting season (Sato et al. 1998, Hays et al. 2002). Not all adult turtles will breed each year, but the relative numbers arriving annually at widely separated rookeries in Australia and the Indo-Pacific are similar, suggesting large-scale environmental forcing on reproductive success (Limpus & Nicholls 1988, Chaloupka 2001). Variation in winter sea-surface temperature anomalies partly explains internesting intervals of a Costa Rican population of green turtles Chelonia mydas, with 2-yr remigration probabilities increasing in warmer years (Solow et al. 2002). In Australia, interannual fluctuations in numbers of green turtles nesting at rookeries within the Great Barrier Reef are positively correlated with the Southern Oscillation Index, also with a 2-yr lag (Limpus & Nicholls 1988). Modelling studies suggest breeding intervals (time between nesting years) are determined by resource provisioning on adult feeding grounds and the 2-yr lag represents the time required for physiological provisioning for reproduction and migration (Hays 2000, Rivalan et al. 2005). Green turtles are herbivorous so are likely to be tightly coupled to productivity in coastal waters (Broderick et al. 2001). Mean egg-laying dates of many terrestrial bird species around the world have advanced considerably in response to increasing temperatures (Archaux 2003, Both et al. 2004, 2005, Moller et al. 2006). Migratory species are arriving earlier and leaving later (Mason 1995, Crick et al. 1997, Lehikoinen et al. 2004, Marra et al. 2005, Jonzén et al. 2006). Most evidence is from the Northern Hemisphere, but a similar pattern has recently been found in Australian migratory wetland birds such as the curlew sandpiper Calidris ferruginea and the double-banded plover Charadrius bicinctus (Beaumont et al. 2006). It is assumed that such changes are also occurring in Australian seabirds. Protracted breeding seasons observed in seabird species in Western Australia are likely to be a 429



response to changing climate (Dunlop & Wooller 1986, Chambers et al. 2005). Breeding success of little penguins Eudyptula minor in Bass Strait is correlated with sea temperatures and mean laying dates are earlier in warmer years (Chambers 2004). Pelagic systems Physiology All plankton are poikilothermic and thus physiological rate processes and rates of overall growth are highly sensitive to temperature (Eppley 1972, Peters 1983, Huntley & Lopez 1992), with many plankton having a Q10 between 2 and 3 (i.e., a doubling to tripling in the speed of rate processes for a 10°C temperature rise). Species have a thermal optimum where growth is maximal and thermal limits beyond which net growth ceases or becomes negative. Basal metabolic losses increase with increasing temperature so that zooplankton fitness and, subsequently, abundance and distribution may be adversely affected. Little information is available on temperature ranges for Australian plankton, and in most cases experiments have been carried out with temperate plankton strains. Culture studies do give some indication (e.g., Smayda 1976) and suggest that species with tropical and subtropical distributions have growth optima

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