Anthropogenic Disturbances and Shifts in Tropical Seagrass Ecosystems

Johan S. Eklöf

Doctoral Thesis in Marine Ecotoxicology

Department of Systems Ecology Stockholm University Stockholm, Sweden

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Doctoral dissertation 2008 Johan S. Eklöf Department of Systems Ecology Stockholm University SE – 106 91 Stockholm Sweden ©Johan S. Eklöf, Stockholm 2008 ISBN 978-91-7155-552-6 Printed in Sweden by Intellecta Docusys, Stockholm 2008 Distributor: Stockholm University Library Cover image: Jerker Lokrantz, Azote images (www.azote.se)

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To my parents

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ABSTRACTABSTRACTABSTRACTABSTRACT

Seagrasses constitute the basis for diverse and productive ecosystems worldwide. In East Africa, they provide important ecosystem services (e.g. fisheries) but are potentially threatened by increasing resource use and lack of enforced management regulations. The major aim of this PhD thesis was to investigate effects of anthropogenic distur-bances, primarily seaweed farming and coastal fishery, in East African seagrass beds. Seaweed farming, often depicted as a sustainable form of aquaculture, had short- and long-term effects on seagrass growth and abundance that cascaded up through the food web to the level of fishery catches. The coastal fishery, a major subsistence activity in the region, can by removing urchin predators indirectly increase densities of the sea urchin Tripneustes gratilla, which has overgrazed seagrasses in several areas. A study using simulated grazing showed that high magnitude leaf removal – typical of grazing urchins – affected seagrasses more than low magnitude removal, typical of fish grazing. Different responses in two co-occurring seagrass species furthermore indicate that high seagrass diversity in tropical seagrass beds could buffer overgrazing effects in the long run. Finally, a literature synthesis suggests that anthropogenic disturbances could drive shifts in seagrass ecosystems to an array of alternative regimes dominated by other or-ganisms (macroalgae, bivalves, burrowing shrimp, polychaetes, etc.). The formation of novel feedback mechanisms can make these regimes resilient to disturbances like sea-grass recovery and transplantation projects. Overall, this suggests that resource use ac-tivities linked to seagrasses can have large-scale implications if the scale exceeds critical levels. This emphasizes the need for holistic and adaptive management at the seascape level, specifically involving improved techniques for seaweed farming and fisheries, protection of keystone species, and ecosystem-based management approaches. Keywords: Keywords: Keywords: Keywords: aquaculture; East Africa; ecosystem change; feedback mechanisms; Kenya; management; overgrazing; regime shifts; resilience; seagrass; seaweed farming; sea ur-chins; Tanzania; Tripneustes gratilla; trophic cascades; Zanzibar

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PERSONAL REFLECTIONS

So here I sit with an almost finished doctoral thesis – a new, strange and very pleasant feeling! This small book marks the end of an interesting, rewarding and joyful four-year journey, and at the same time the beginning of a new one into uncharted waters. To the many that have followed me along the way I would like to say thank you from the bot-tom of my heart.

There are, however, some who deserve special credit: My main supervisor Nils Kautsky: without your support, starting when I was just a wee

degree project student, and ranging from reading manuscripts and filling my pockets with extra cash to explaining the fine arts of supervision and cooking Cataplana, I would not have been where I am today. In the future I will do my best to spread the same contagious enthusiasm that you infect us with everyday.

My associate supervisor Patrik Rönnbäck: even though often physically distant, you’ve always been there when needed, and we’ve had our fair share of good times. I still look forward to do some actual field work together with you in the future! And since it’s a free world we of course Keep on Rocking…

Maricela de la Torre-Castro: for being one of my dearest colleagues, coauthors and friends, for unending interest in my work, and all good times on Zanzibar and in Sweden. This is just the beginning…

Martin Gullström: for being a great colleague and friend. I hope that once we’ve fin-ished everything we’ve started, we’ll have many more opportunities to continue our exploration of the underwater world…

Mats Björk: for being a great mentor, both when it comes to seagrass physiology as well as ways of addressing marine science from a development perspective. Hope to see more of you in the future, both in Sweden and on Zanzibar!

Mr Mcha Mzee Manzi with family, Daudi and Rashidi, and the rest of Chwaka village: for showing me the true face of Zanzibar and the wonderful Swahili culture. The memories of our times together are forever with me.

My degree project students Malin Andersson, Camilla Nilsson, Rebecka Henriksson, Maria Asplund, Annika Dahlgren, Sara Fröcklin, Annika Lindvall and Nadja Stadlin-ger: for indirectly teaching me supervision in the most direct of ways; for all those days, weeks and months you spent in the field, and for putting up with me, malaria and the rice-and-fish diet. Without you this thesis would undoubtedly have been a lot thinner…

My ‘African’ coauthors Narriman S. Jiddawi, Jacqueline N. Uku and Tim R. McClanahan: for hands-on knowledge and interest in my work that has been a tre-mendous help over the years. Ahsanteni!

Åsa Forss: my fellow seaweed farmer, and a great “bollplank” who opened my eyes to the wider aspects of aquaculture. I really look forward to reading your thesis!

Albert Norström and Jerker Lokrantz: my fellow musketeers for 9.5 years and count-ing… All those hours of fun at KÖL, BIG, Skäggvik, Gula Villan, Mercuries, Africa

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House, Chumbe, and last but not least in room 242, made this journey endurable. You guys rock!

Paul Lavery, Kathryn McMahon, and all the other guys at ECU: for inviting me, and for making my Ozzie experience such a smooth and fun one! Hope to see more of you in the future!

Past and present course assistants, course leaders and students at ‘Marine Biology’ and ‘Marine Ecology’, and the staff at TMBL: for summer weeks at Tjärnö filled with wa-ter, sun and fun.

Clare Bradshaw, Ian Bryceson, Ragnar Elmgren, Tomas Elmqvist, Klemens Britas Eriksson, Carl Folke, Hasse Kautsky, Lena Kautsky, Mats Lindegarth, Jon Norberg, Magnus Nyström, Moks, Sara Sjöling, Sofia Wikström, and Marcus Öhman: for shar-ing thoughts and opinions that have driven me further and tickled my research inter-est.

Colleagues at the department of Systems Ecology, in particular all past and present members of the Ekotox group, Elin E, Erik A, Gustaf A, Marc, Stephan, Jakob vH, Henrik E, Sara B, Ninni, Bea, Anders W, Matilda, Antonia, Lisa A and Sussie Q: for a very nice and inspiring atmosphere that I’m very glad to be part of.

Friends outside of Academia, especially Ola K & Johanna, Pella L with family, Klas, Pelle P-Boy, Pelle King, Nettan, Susanna, Nicke J with family, Mia B, Jakob, Oskar, Lotta, Anna and Karin N: for being there and keeping my mind on the important things in life.

My family and relatives Farmor and Stickan, my uncles Per B and Per E and aunt Cissi with families, sister Anna & her Erik, and last but certainly not least my parents Karin and Sven: for endless support and encouragement that helped me get here.

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TABLE OF CONTENTS

LIST OF PAPERS 9

INTRODUCTION 10

SEAGRASS BEDS IN A CHANGING WORLD 11

Seagrasses from species to seascape 11

Seagrasses support vital services to society… 12

…but are threatened by anthropogenic disturbances 12

Seagrasses and disturbances in East Africa 14

MAJOR AIMS OF THESIS 18

METHODS 19

GENERAL RESULTS AND DISCUSSION 23

Effects of seaweed farming on seagrass ecosystems 23

Cascading effects of fisheries in seagrass food webs 26

Regime shifts in seagrass ecosystems 30

MAJOR IMPLICATIONS 34

Coastal management in an East African context 34

Addressing complexity in seagrass management 37

SAMMANFATTNING PÅ SVENSKA 39

ACKNOWLEDGEMENTS 43

REFERENCES 44

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LIST OF PAPERS

This thesis is based on the following six papers, referred to in the text by their Roman numerals: IIII Eklöf JS, de la Torre Castro M, Adelsköld L, Kautsky N, Jiddawi NS.

(2005). Differences in macrofaunal and seagrass assemblages in seagrass beds with and without seaweed farms. Estuarine, Coastal and Shelf Science 63(3): 385-396.

IIIIIIII Eklöf JS, Henriksson R, Kautsky N. (2006). Effects of tropical open-water

seaweed farming on seagrass ecosystem structure and function. Marine Ecology Progress Series 325: 73-84.

IIIIIIIIIIII Eklöf JS, de la Torre Castro M, Nilsson C, Rönnbäck P. (2006). How do

seaweed farms influence fishery catches in a seagrass-dominated setting in Chwaka Bay, Zanzibar? Aquatic Living Resources 19(2): 137-147.

IVIVIVIV Eklöf JS, Fröcklin S, Stadlinger N, Dahlberg A, Kimathi P, Uku J,

McClanahan TR. (Manuscript). Fishing, trophic cascades, and overgrazing of Kenyan seagrass beds. Submitted to Ecological Applications.

VVVV Eklöf JS, Gullström M, Björk M, Asplund M, Dahlgren A, Hammar L,

Öhman MC. (Manuscript). Physical responses of two co-occurring sea-grasses to different grazing regimes. In review for Aquatic Botany.

VIVIVIVI Eklöf JS, de la Torre-Castro M. (Manuscript). Seagrass loss and feedback

mechanisms: multiple regimes in seagrass ecosystems. The published papers are reprinted with the kind permission of the publishers.

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INTRODUCTION

Humankind is utterly dependent on the continued flow of ecosystem services (Costanza et al. 1997; Daily 1997), but currently transform the biosphere in an unprecedented manner (Vitousek et al. 1997; IPCC 2007) which threatens the flow of these very services (Loreau et al. 2002). This situation is particularly severe in coastal zones (<100 km inland) because these naturally dynamic areas suffers from extreme over-population (Shi and Singh 2003) and rapid develop-ment (Hinrichsen 1995). Outtake of coastal resources like fish and aquatic invertebrates is a strong contributor, especially in developing areas, where they constitute the cheapest and most accessible form of protein. Besides direct ef-fects on catches (e.g. Jiddawi and Ohman 2002; Stobutzki et al. 2006), removal can cascade through food webs and cause major habitat changes like overgraz-ing of macrophytes and bioerosion of corals (see Pinnegar et al. 2000 for re-view).

When the resource outtake (e.g. fisheries) exceeds sustainable levels the pres-sure on the resource base often becomes self-fuelling because the underlying socio-economic drivers are themselves prone to modification by change in the supply of services (Kremer and Crossland 2002). Together with changes in environmental conditions and natural disturbance regimes this can cause unex-pected and often ‘catastrophic’ shifts to alternative ecosystem regimes (Scheffer et al. 2001), which can be more or less permanent due to ‘hysteresis’ effects i.e. that thresholds for reverse shifts are different from those of initial shifts (Scheffer et al. 2001).

One of the suggested approaches to deal with overfishing is aquaculture, which aids the production of low-cost food (e.g. FAO 1994; Tacon 2001; FAO 2003) especially in tropical developing countries (Hasan 2001; Tacon 2001). Many forms are however resource-inefficient monocultures feeding first-world consumers (Naylor et al. 2000) at the expense of habitat destruction and declin-ing fish stocks in production areas often situated in developing countries (Rönnbäck 2001). Hence, there is a need for alternative forms that can ensure a sustainable flow of food and income (Rönnbäck et al. 2002).

In light of this background, this thesis deals with effects of some common an-thropogenic disturbances – primarily seaweed farming and fisheries - in one of the most important but least studied of coastal ecosystems: tropical seagrass beds. In the following sections I provide the reader with a background to sea-grass ecosystems from a biological to an anthropocentric point of view, and present the rationale behind the specific cases and questions I have addressed. Following a general overview and discussion of the results, I conclude with the major implications of my work.

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SEAGRASS BEDS IN A CHANGING WORLD Seagrasses from species to seascape

Seagrasses are a polyphyletic group of ca. 60 species of marine clonal angio-sperms (Green and Short 2003) that form beds or meadows along all continents except Antarctica (Robertson and Mann 1984). Most species require sediment bottoms, high light influx and oligotrophic conditions, limiting general distri-bution to shallow (ca. 0 to 10 m depth), more or less sheltered and well-lit areas (with some discrepancy between species and populations). In addition to other abiotic factors such as temperature, salinity, and exposure, seagrass distribution is regulated by biotic interactions such as grazing (Valentine and Duffy 2006), intra- and inter-specific competition (Williams 1987; Davis and Fourqurean 2001) and facilitation (Williams 1990; Reusch et al. 1994). Another highly important aspect is that seagrasses self-regulate their distribution by biotic feed-back mechanisms as ‘ecosystem engineers’ (sensu Jones et al. 1994), most im-portantly by stabilizing sediments which decreases turbidity (de Boer 2007).

The species diversity of seagrasses is generally low compared to that of other habitat-forming organism (e.g. macroalgae or corals), but community diversity is high, with representatives from all major phyla on a global scale (Hemminga and Duarte 2000). Also the abundance of associated organisms is generally higher than in unvegetated areas (Pihl 1986; Boström and Bonsdorff 1997; Arrivillaga and Baltz 1999: paper I, IIIpaper I, IIIpaper I, IIIpaper I, III), primarily due to an extraordinarily high rate of primary production (Duarte and Chiscano 1999) supporting sec-ondary production (Mateo et al. 2006); provision of a three-dimensional struc-ture in the water column (Bologna and Heck 1999; Salita et al. 2003), and the creation of calm microclimates (Hemminga and Duarte 2000).

On a landscape level seagrass beds often constitute a network of patches, be-tween which interactions (e.g. spread of organisms) are regulated by factors such as species-specific growth rates and major means of dispersal (Bell et al. 2006), patch size, degree of fragmentation and species identity (Bostrom et al. 2006). On the ‘seascape’ level, seagrass beds are open systems connected through ex-change of organic material, nutrients and movement of species with other coastal and terrestrial systems (see e.g. Ogden 1988; Moberg and Ronnback 2003; Harborne et al. 2006). A growing number of studies highlight the impor-tance of such cross-system (or habitat) interactions, primarily in terms of how the presence of and distance to other systems affect landscape dynamics. In the tropical literature much focus has been on interactions between seagrass beds and coral reefs, primarily in terms of fish community structure (Nagelkerken et al. 2000; Dorenbosch et al. 2005a; Grober-Dunsmore et al. 2007) but also

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processes like herbivory and predation (e.g. Ogden et al. 1973; Ogden 1988; Valentine et al. 2007).

At the highest scale, we are just starting to acknowledge seagrass beds as part of integrated social-ecological systems (de la Torre-Castro 2006), where another set of interactions (resource extraction, provision of goods and services, anthro-pogenic disturbances, etc.) between system components are regulated by abiotic and biotic, as well as social, economic and political factors. In an increasingly globalized world, this suggests that seagrass beds are affected by various factors such as spread of invasive species, ocean currents, climate change, international trade, political change, and spatially span across global regions. Seagrasses support vital services to society…

Seagrasses either directly or indirectly provide a range of ecosystem services to coastal societies (Duarte 2000; de la Torre-Castro and Rönnbäck 2004): al-though present in only 0.15% of the ocean surface, seagrasses (Smith 1981) and their epiphytes (Moncreiff and Sullivan 2001) are highly important contribu-tors to the primary production in the global oceans, which supports a substan-tial secondary production of in many cases economically important taxa like fish and crustaceans (Erftemeijer and Middleburg 1993; Jackson et al. 2001; de la Torre-Castro and Rönnbäck 2004). Furthermore, the leaf canopy reduces water flow velocity (Koch 1996), which increases settling of particles and sedi-ment organic matter content within meadows (Smith 1981; Gacia et al. 1999). Together with seagrass roots and rhizomes stabilizing sediments (Fonseca 1989), this reduces turbidity (Bulthuis et al. 1984) and coastal erosion (Almasi et al. 1987). Altogether, these services makes seagrass beds one of the most valu-able systems on a global scale (Costanza et al. 1997). …but are threatened by anthropogenic disturbances

Disturbance is an intrinsic process in ecosystems that regulates diversity and production and drives evolution. Seagrasses evolved under disturbance in the form of intensive grazing by megaherbivores (Domning 2001; Valentine and Duffy 2006), but in current-day systems humans are the dominating agent of disturbance: seagrasses currently experience a global crisis (Orth et al. 2006) caused by pollution, excessive removal or organisms, direct mechanical distur-bance, and alterations of natural disturbance regimes (Short and Wyllie-Echeverria 1996; Duffy 2006). Because of the disproportional importance of seagrasses, this affects the structure of associated communities, basic processes driving ecosystem functioning, and ultimately the flow of ecosystem services

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(Duarte 1995; Deegan et al. 2002). At the same time, seagrass recovery is often very slow, to the verge of non-existing (Larkum et al. 1989; Holmquist 1997), partially because of naturally slow recovery of many species, and that seagrasses to such an extent create their living-conditions (Duarte 1995; van der Heide et al. 2007). The importance of indirect effects

Indirect effects of anthropogenic disturbance are a major factor explaining sea-grass loss. The most important is decreased light penetration caused by in-creased sedimentation from land runoff and eutrophication-induced blooms of macro- and micro-algae (Short and Wyllie-Echeverria 1996). Another possible mechanism behind algal blooms currently receiving considerable attention is cascading effects of fishing of top predators (e.g. Williams and Heck Jr 2001; Valentine and Duffy 2006): in natural abundances mesograzers (e.g. crusta-ceans) can control algal growth and even buffer effects of eutrophication (Hughes et al. 2004), suggesting that seagrass food webs could be sensitive to removal of top predators. In fact, overfishing of top predators, and not eutro-phication, was recently suggested to be the major culprit behind habitat loss in shallow benthic systems like seagrass beds (Heck Jr and Valentine 2007).

Overfishing could also indirectly release various seagrass grazers from preda-tion control, e.g. urchins (Peterson et al. 2002; Alcoverro and Mariani 2004: paper IV) and fish (Valentine et al. 2007; Prado et al. unpublished), ultimately resulting in seagrass loss (Peterson et al. 2002). However, other factors like cross-system energy subsidies (Valentine and Heck 2005; Valentine et al. 2007), habitat size (Prado et al. unpublished) and eutrophication (Tewfik et al. 2005) also influence macrophyte-grazer interactions, and must be taken into account when evaluating the potential effects of fishing. Are there regime shifts in seagrass systems?

Seagrass beds are often subjected to multiple anthropogenic and natural dis-turbances, that synergistically affect ecosystem functioning (Lotze et al. 2006; Orth et al. 2006). A growing body of literature suggest that anthropogenic dis-turbances (e.g. eutrophication) could cause shifts to alternative ‘regimes’, ‘phases’ or ‘stable states’ where other organisms like macroalgae dominate eco-system functioning (Duarte 1995; Gunderson 2001; Munkes 2005; Valentine and Duffy 2006). After such shifts, ‘undesirable’ feedback mechanisms can self-fuel the dominance of these organisms which prevent seagrass recovery and the success of management approaches like pollution control (e.g. Munkes 2005).

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This is so far a relatively small area in seagrass research, but the demonstration of regime shifts could help to explain the often slow or non-existing recovery as well as the low success rate of restoration projects (Campbell 2002), and there-fore be highly important from a management point of view.

The global ‘seagrass crisis’ emphasizes the importance of seagrass management. At the turn of the century this was the least explored area in seagrass research (Duarte 1999), and despite much work during recent years (see e.g. Wood and Lavery 2000; Kirkman and Kirkman 2002; Orth et al. 2002; Thom et al. 2005; de la Torre-Castro 2006), a recent review concluded that seagrass management is still inadequate on a global scale (Walker et al. 2006). The problem is partly a lack of public appreciation of the values being lost (Orth et al. 2006), but also the lack of synthesis of the dynamics of seagrass loss and recovery (Duarte 1999; Walker et al. 2006).

While improved seagrass management strategies are needed globally, I and others suggest that they are especially important in tropical developing countries where (1) seagrasses are rarely included in management plans (de la Torre-Castro 2006); (2) local coastal communities are often more or less dependent on seagrass-associated services (Gell 1999; de la Torre-Castro and Rönnbäck 2004; de la Torre Castro and Jiddawi 2005); and (3) our general knowledge on seagrass dynamics, especially regarding disturbance and recovery, is generally scarce (Green and Short 2003). Seagrasses and disturbances in East Africa

The field work for this thesis was conducted in coastal areas of Eastern Africa. This is a seagrass ‘hot-spot’ in the western side of the Tropical Indo-Pacific seagrass bioregion, the largest and most diverse (Short et al. 2007) but least studied (Duarte 1999) of the five global seagrass bioregions.

Seagrass beds in the area support a high primary production (Kamermans et al. 2000), which together with the high structural complexity of many species makes seagrass beds an important economic resource through production of fish and shellfish (Gullström et al. 2002; de la Torre-Castro 2006). In major sea-grass areas like Chwaka Bay (Zanzibar, Tanzania), this is illustrated by seagrass beds being the most preferred fishing grounds, and seagrass-associated fish be-ing the most important market species (de la Torre-Castro and Rönnbäck 2004).

Despite low industrialization, rapid coastal development threatens seagrasses due to dredging, clearing for tourism and pollution (Ochieng and Erftemeijer 2003). In addition, unsustainable extraction of coastal resources to feed a rap-

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idly growing coastal population constitutes one of the key threats to East Afri-can coastal zones, including seagrass beds (Payet and Obura 2004). Is seaweed farming really a sustainable aquaculture?

Open-water farming of macroalgae (‘seaweeds’) was initiated in East Africa and Tanzania in the late 1970s (Mshigeni 1976), and achieved its breakthrough in the late 1980s when Philippine strains of two red algae (Rhodophyta), Eucheuma denticulatum (N.L. Burman) F.S.Collins & Hervey and Kappaphycus alvarezii Doty, were introduced to Unguja Island, Zanzibar (Lirasan and Twide 1993). The algae are farmed in shallow coastal areas for extraction of carra-geenan (FAO 2002), a polysaccharide used as a thickening agent in food, cos-metics, and pharmaceuticals (Philexport, 1996; FMC Biopolymer, 2003).

Seaweed farming is often depicted as ‘one of the most sustainable forms of aquaculture’ since (1) no feed, fertilizers or pesticides are used, (2) farming is claimed not to alter the physical environment in any major way (Johnstone and Ólafsson 1995), and (3) the new income to farmers (primarily women) can boost local economies (Pettersson-Löfquist 1995; Semesi 2002). Despite these benefits, seaweed farming de facto introduces macroalgae in habitats where they normally do not occur. This suggests that in large quantities, they could com-pete with other habitat-forming organisms for light and space, and also affect community composition of associated organisms by attracting or deterring mo-bile species (Zemke-White and Smith 2006).

Seaweed farms are placed in seagrass beds where vegetation-free areas are lack-ing or where farmers believe that seagrasses fertilize seaweeds (de la Torre-Castro and Rönnbäck 2004). Some farmers initially remove seagrasses to sim-plify farming (Collén et al. 1995; de la Torre-Castro and Rönnbäck 2004), and trampling (e.g. Eckrich and Holmquist 2000) and boat moorings (Walker et al. 1989) could also negatively affect seagrasses. In addition, the seaweeds could negatively affect seagrasses and associated organisms through shading, in the same way as blooms of free-floating macroalgae (Hauxwell et al. 2001; McGlaherty 2001). Scaled up to the system level, this could indirectly affect fish catches and sediment stabilization, and cause a trade-off in ecosystem ser-vices to coastal societies. Is there a link between seagrass overgrazing and coastal small-scale fisheries?

Artisanal fishing is the main subsistence activity in East Africa (Jiddawi and Ohman 2002). Most of the fishing is small-scale and inshore, using simple methods like drag nets, stationary basket traps (i.e. madema on Zanzibar),

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hook-and-line or spear from small dug-out canoes (i.e. mashua) and sail vessels (e.g ngalawa). With the introduction of nylon drag nets with small mesh size, outboard engines propelling larger boats, and changes in informal fishing insti-tutions (de la Torre-Castro 2006), fishing intensity has increased greatly during the last decades (Jiddawi and Ohman 2002). This negatively affects fish density, individual size and catch sizes of key target species in most areas (Jiddawi and Ohman 2002; McClanahan and Mangi 2004). Due to ‘poverty traps’ and poor management this feeds a negative spiral of continued fishing and evermore de-creasing fish stocks (Cinner et al. 2007).

Besides such direct effects, fishing indirectly affects coastal ecosystems. Illegal methods like drag nets (Mangi and Roberts 2006) and dynamite fishing (Obura 2001) degrades habitat structure, but more importantly, there is strong evidence for cascading effects in coastal food webs. In coral reefs, excessive outtake of urchin predators like triggerfish (Balistidae) and wrasse (Labridae) releases sea urchins from predation control, resulting in reduced habitat complexity and subsequent changes in fish abundance (McClanahan and Muthiga 1989; McClanahan and Obura 1995). Because of the slow growth rate and limited dispersal of these predators (Kaunda-Arara and Rose 2004), ecosystem recovery takes decades (McClanahan and Graham 2005).

In present-day Tanzanian and Kenyan seagrass beds the generalist sea urchin Tripneustes gratilla is, together with parrotfish (Gullström et al. unpublished), the most common seagrass macrograzer in fished areas (Alcoverro and Mariani 2004). During the last decade, hyperabundant populations of T. gratilla have been observed to overgraze complete seagrass beds of primarily Thalassodendron ciliatum in at least three areas along the Kenyan coast; Mombasa (Alcoverro and Mariani 2002), Watamu (Zanre and Kithi 2004), and Diani (Uku et al. in prep.). So far, no studies have directly investigated the direct causes to these overgrazing events, but the fact that urchin grazing is generally more common (with some exceptions) than fish grazing in fished areas (Alcoverro and Mariani 2004), clearly suggest that cascading effects of overfishing could be a major factor. There is, however, a clear need for experimental studies assessing whether fishing by reducing predation control on T. gratilla indirectly contrib-utes to increases in urchin abundance, since overgrazing within marine parks without fishing (Watamu, Mombasa and Chumbe) indicate that other factors, e.g. eutrophication (Tewfik et al. 2005), distance to coral reefs (Ogden et al. 1973) and the presence of shelter (Heck and Valentine 1995), could have over-riding influence on urchin populations.

Seagrass overgrazing is undeniably the strongest outcome of the interaction between seagrasses and grazers. The exact effect of grazing depends on factors such as grazing intensity, species- and population-specific sensitivity to grazing

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(Cebrian et al. 1998), the presence of other disturbances like shading (Macia 2000), and seasonal changes in light and temperature (Valentine and Heck 1991). Another grazing-related factor that has received little attention in sea-grass ecology, but has a major influence on growth of terrestrial grasses (e.g. Turner et al. 1993) is the frequency of grazing and the ‘grazing history’ (when did previous grazing occur, and what was the magnitude). This is because re-duced levels of stored carbohydrates, used to compensate for loss of biomass from grazing, will greatly affect the possibility to respond to additional grazing (Dyer et al. 1993). Furthermore, there is virtually no knowledge on the poten-tial interaction between the intensity and magnitude of grazing on seagrass beds in general.

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MAJOR AIM OF THESIS

The major aim of this thesis was to investigate direct and indirect effects of anthropogenic disturbances on tropical seagrass ecosystem structure and func-tion, and what this implies for coastal management. The thesis consists of three parts; two case studies conducted in East Africa on (1) open-water seaweed farming and (2) overgrazing and coastal fisheries, and (3) a synthesis on regime shifts in seagrass systems on a global scale. A conceptual overview of the thesis and respective papers is presented in Figure 1.

The following questions were addressed for respective part: (1) How, why, and to what degree does open-water seaweed farming affect tropical seagrass ecosystems, and are effects strong enough to cause trade-offs i.e. loss of ecosystem services? (2) Are there indirect effects of fisheries on sea urchin-seagrass interactions in tropical areas, and how do changes in grazing regimes affect seagrasses? (3) Can anthropogenic-induced changes in environmental conditions and simplification of seagrass food webs drive regime shifts in seagrass beds, and if so, what are the management implications?

Seagrass

regimeAlternative

regime

Regime shifts (VI)

Disturbances

Ecosystem services (III)

Desirable

feedbacks (IV, VI)

Undesirable

feedbacks (VI)

Feedbacks

(VI)

Resource extraction

& management

Overgrazing

and fisheries (IV, V)

Seaweed

farming (I, II, III)

Managing

seagrass loss (VI)

Fig 1. Conceptual model of thesis, highlighting the topics of the different papers.

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METHODS Study areas

The field work was conducted in two major areas: (1) Chwaka Bay (Zanzibar, Tanzania) and (2) the southern Kenyan coast (Fig. 2). The Kenyan and Tanza-nian coastline (600 and 800 km, respectively) has a narrow continental shelf, characterized by fringing coral reefs, lagoons with extensive seagrass and algal beds, limestone cliffs, mangrove forests, sand dunes and beaches (UNEP 1998; UNEP 2001). The seagrass flora comprises c. 12 species, with Thalassodendron ciliatum (Forskål) den Hartog. dominating and Enhalus acoroides (L.f.) Royle, Thalassia hemprichii (Ehrenberg) Ascherson, Cymodocea serrulata (R. Brown) Ascherson and C. rotundata Ehrenberg & Hemprich ex Ascherson also forming mixed and monospecific meadows (Ochieng and Erftemeijer 2003). The tidal regime is semidiurnal with two peaks and lows per day, and an amplitude rang-ing from roughly 1 to 3.5 m in neap and spring tides, respectively (Cederlöf et al. 1995). For more detailed descriptions of study areas, see respective papers.

Kenya

Tanzania Zanzibar

A

B

C DN

4 km

200 km

42˚E

4˚S

Fig 2. Map over study areas. (A) Africa with Kenya and Tanzania highlighted, (B) the Kenyan and Tanzanian coastlines highlighting Zanzibar, (C) Unguja Island (Zanzibar, Tanzania) and (D) Chwaka Bay (East coast of Unguja Island). Grey areas represent land, white is water, filled black is mangroves and leaves are seagrasses.

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Effects of seaweed farming

In the first study (paper Ipaper Ipaper Ipaper I) we investigated differences in seagrass, macrofauna (>0.5 mm) and sediment in three seagrass beds, two seaweed farms with Eucheuma denticulatum (established on seagrass beds in the mid 1990s), and a sand bank (included to control for the presence of vegetation). Since most stud-ies on effects of seaweed farming at this time (2004) were based on comparisons between farms and control sites (including paper I paper I paper I paper I and III III III III), there was an im-mediate need to experimentally validate previously observed patterns and iden-tify mechanism(s) behind effects. In the second study (paper IIpaper IIpaper IIpaper II), the effects of seaweed farming on a mixed seagrass community were experimentally investi-gated over 11 wks in replicated plots in three treatments: seaweed farms, con-trols, and procedural controls (with sticks and lines but without the algae). Variables included standard seagrass aboveground metrics sampled every 15 days, and SOM-content, seagrass epiphyte cover, epifauna community structure (>2cm), accumulation of seagrass detritus and algal shading of seagrasses, sam-pled at the end of the experiment.

Since seagrasses, which are key habitats for important fishery species in the study area (de la Torre-Castro and Rönnbäck 2004), seemed to be affected by seaweed farming (paper I and IIpaper I and IIpaper I and IIpaper I and II), farming effects could cascade to fish com-munities (Bergman et al. 2001) and ultimately fishery catches. In the third study (paper IIIpaper IIIpaper IIIpaper III) we investigated how a seaweed farm (and the farmed seaweeds in particular) influenced fish catches, using a local artisanal fishing method (dema basket traps). In the first of two field studies, fish catches from three sites (a seaweed farm, a seagrass bed and a sand bank) were compared over three neap tides. In a second study the particular influence of the farmed algae (E. denticulatum) was investigated within a seaweed farm over a five-day period.

Urchin overgrazing and indirect effects of fishing

Two field experiments were conducted in Kenya to assess the effect of fishing

on the urchin Tripneustes gratilla (paper IVpaper IVpaper IVpaper IV). The choice of study area was based on the presence of (1) comparable seagrass beds, (2) several MPAs inter-spersed between fished areas along a more or less homogenous coastline, and (3) at least three documented sea urchin overgrazing events during the last decade.

In the first of two studies the effects of fishing in time and space was investi-gated by replicated sampling of T. gratilla density at 16 occasions from 1987 to 2006 in seven protected and fished reefs situated along a 150 km stretch of coast. Based on these results, a second in-depth study (conducted in 2006) fo-cused on effects and interactions of three factors presumed to affect T. gratilla:

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(1) fishing (by sampling in two fished and two protected areas), (2) the distance to coral reefs (by sampling in two sites within each area: Close and Far from the reef), and (3) presence of shelter (by comparing an Unvegetated site with the Far vegetated site in each of the four areas). Variables included urchin density, diversity, size, and relative predation rate on T. gratilla, assessed using a modi-fied version of the tethering method (McClanahan and Muthiga 1989): five randomly chosen urchins were pierced and tied using a 0.5 m nylon fishing line at regular intervals on to replicated 7 m nylon filaments attached to the bottom. Based on the survival of tethered urchins every 24h for three days, a relative Predation Index was calculated. The responsible predators were assessed by inspection of remaining urchin tests, following a standard method developed in the study area by one of the co-authors (McClanahan and Muthiga 1989). Fi-nally, urchin grazing pressure on seagrasses was assessed by a natural herbivore assay (Alcoverro and Mariani 2004), in which shoots of two dominating sea-grass species in the area, T. ciliatum and Thalassia hemprichii, were collected in each of the vegetated sites (Close and Far from reefs). Leaf turnover rates, which can affect the number of grazing marks, was not measured since a previous study showed no difference between these same four areas (Alcoverro and Mari-ani 2004). The presence/absence of urchin bite marks was later noted for each leaf, and used to calculate a Grazing Index on a shoot basis.

Fishing generally seems to affect dominating grazers in Kenyan seagrass beds, with fish and urchins dominating in protected and fished areas, respectively (Alcoverro and Mariani 2004). Herbivore assays using T. hemprichii leaves sug-gest that sea urchins feed with a greater intensity (more leaf area removed) than fish (McClanahan et al. 1994), while fish (primarily parrotfish like Leptoscarus vaigiensis) feed regularly and sometimes in the same areas (Macia and Robinson 2005). In addition, grazing frequency is known to be important in terrestrial grasses (Turner et al. 1993) but has not been tested in seagrasses. We do how-ever know that co-occurring species often respond differently to grazing (e.g. Cebrian et al. 1998; Alcoverro and Mariani 2005). Based on this, we then in-vestigated how different combinations of grazing intensity and frequency (using leaf clipping) affected shoot growth and rhizome carbohydrates in two co-occurring seagrass species, in this case T. hemprichii and Enhalus acoroides (ppppa-a-a-a-per Vper Vper Vper V). The reason for not using T. ciliatum , which has been overgrazed by the sea urchin T. gratilla (e.g. Alcoverro and Mariani 2002), was logistical con-straints in finding a site where this species co-occurred with T. hemprichii. However, for the specific questions addressed in the study (is there a difference in response between co-occurring species’), the choice of species was regarded less important.

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Regime shifts in seagrass beds (paper VIpaper VIpaper VIpaper VI)

The final paper of the thesis is a literature synthesis on regime shifts in seagrass

beds. The idea sprung partly from the findings of the two case studies (seaweed farming and fisheries) about changes in seagrass ecosystem structure and func-tion, and the growing understanding about the fundamental role of feedback mechanisms in buffering disturbances or contributing to change in ecosystems (e.g. Mayer and Rietkerk 2004; de Boer 2007; van der Heide et al. 2007).

The major aims was to investigate the occurrence of regime shifts in seagrass beds, elucidate potential mechanisms causing and maintaining shifts, and dis-cuss the implications of regime shifts for seagrass management. To do this, we conducted a literature review of published articles (using ISI web of Science and ASFA), book chapters and reports on shifts in seagrass beds.

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GENERAL RESULTS AND DISCUSSION Effects of seaweed farming

The results of paper Ipaper Ipaper Ipaper I showed that the two seaweed farm sites generally had less seagrass (% cover, biomass, shoot density and canopy height) than the three seagrass sites. Based on claims from local seaweed farmers that seagrasses gener-ally disappear after a few months of farming (de la Torre-Castro and Rönnbäck 2004), we attributed these differences to effects of farming. This was corrobo-rated by the experimental farming (paper IIpaper IIpaper IIpaper II), which reduced aboveground sea-grass biomass (of primarily Enhalus acoroides) by 40% compared to controls. The lack of major effects on the second species Thalassia hemprichii could be due to species-specific differences in stress sensitivity, morphology and possibly also reduced interspecific competition from E. acoroides.

Although the mechanisms behind the effects were not explicitly tested, we suggest that a combination of shading (3.6% of surface light reached the sea-grass canopy underneath the algae), emergence stress (due to seagrass leaves becoming exposed during longer periods than normal), mechanical abrasion by the algae and potentially toxic algal exudates (even though this given current knowledge seems less likely, see paper IIpaper IIpaper IIpaper II) caused the observed patterns. Given that seagrasses underneath real farms are also subjected to other farming-associated disturbances (removal, trampling, boat moorings, etc.), and the great difference in scale – experimental plots covered 3.75 m2 for 11 wk, whereas farms cover km2 for decades – the magnitude of effects in real farms is probably much greater than shown in the experimental study.

If similar effects and mechanisms as those presented in paper IIpaper IIpaper IIpaper II contributed to the overall differences observed in paper paper paper paper IIII, the two studies together provide a short- and long-term assessment of farming effects. Some variables like shoot length and growth seem to be affected directly, but the fact that seagrasses re-main within farms after a decade (15-20% cover, although mostly between farm plots, paper Ipaper Ipaper Ipaper I) indicates that a total seagrass loss is probably not likely when farm intensity is kept at moderate levels. However, even with seagrasses remain-ing between plots, changes in sediment structure (e.g. SOM-content and grain size) underneath farms may prohibit regrowth of rhizomes or settlement of new shoots even at local (< 1m) scales (Creed and Amado 1999). In addition, frag-mentation of the meadows caused by the farming could increase the risk of seagrass loss due to natural disturbances like strong waves (Fonseca and Bell 1998).

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a) Invertebrate infauna b) Fish catches

Effects on associated organisms

Since seagrasses constitute both the energetic and structural base of the system, changes in seagrass abundance are generally reflected in associated organisms. This also seems to be the case with seaweed farming: invertebrate macrofauna (>0.5 mm) were less abundant and had a lower total biomass in the seaweed farms than in the seagrass beds, but higher or similar densities and biomass as in the sand flat (paper Ipaper Ipaper Ipaper I). Also the cover of epiphytic algae was 25% lower on seagrasses within farming plats than in controls, which was probably caused by the decrease in shoot length, mechanical abrasion or shading (ppppaper IIaper IIaper IIaper II). For macroscopic epifauna (>2cm, paper IIpaper IIpaper IIpaper II) and fish (based on catches using dema traps, paper IIIpaper IIIpaper IIIpaper III) there was no major difference in either abundance or diversity between seaweed farms and seagrass beds. This is probably because structural complexity – provided more or less by seagrass as well as farmed algae – alone is considered one of the most important structuring factors for near shore mobile fauna (Wheeler 1980; Jenkins and Wheatley 1998).

In terms of the structure of the associated fauna community, a pattern seen in infauna and fish catches (Fig 3), and possibly also epifauna (ppppaaaaper IIper IIper IIper II), was that the farms seemed to harbor an associated community ‘intermediate’ to those found in the seagrass beds and in the unvegetated area (paper I, IIIpaper I, IIIpaper I, IIIpaper I, III). A similar community shift due to farming has also been observed in meiofauna (Ólafsson et al. 1995) and fish communities (Bergman et al. 2001), and is probably caused by taxa adapted to either seagrass or bare sand not being found within the seaweed farms, while more generalist taxa (e.g. the rabbit fish Siganus sutor [paper IIIpaper IIIpaper IIIpaper III] and the sea urchin Echinometra mathei [paper IIpaper IIpaper IIpaper II]) are attracted to

Fig 3. Community structure of (1) invertebrate infauna (biomass of major taxa, paper Ipaper Ipaper Ipaper I) and (b) fish catches (biomass per species, paper IIpaper IIpaper IIpaper II), visualized using MDS plots. Black points are samples in seagrass beds, white are samples within seaweed farms, and crosses are samples within a sand bank without vegetation.

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the algae either as a food source or shelter (Neish 2003). In addition, mechani-cal disturbance in farms could also deter organisms. Overall, the effects of sea-weed farming on associated communities bears close resemblance to those of drift macroalgae in seagrass beds (Deegan et al. 2002; Adams et al. 2004), even though the mechanism behind the relative shift in dominating vegetation type (seagrass to algae) are fundamentally different.

A highly important aspect not addressed in my studies is potential indirect ef-fects on functions performed by associated organisms (e.g. grazing, predation, etc.). For instance, lucinid bivalves, the single taxon most affected by the pres-ence of seaweed farms (paper I)paper I)paper I)paper I), benefit seagrasses by reducing levels of toxic sulphide, while the seagrass leaves provide protection from predators (Barnes and Hickman 1990; Reynolds et al. 2007). Since sulphide stress seems to be an important factor in seagrass decline (Borum et al. 2005), it is possible that the decline of lucinids could accentuate the loss of seagrass underneath farms. Until tested, this however remains an interesting hypothesis. Trade-offs in seaweed farming?

Some of the dominating forms of aquaculture can result in a trade-off of eco-system services to local communities (Rönnbäck 1999; Naylor et al. 2000). My results indicate that this could also be the case in seaweed farming, depending on the intensity and scale of farming. First, seagrass production decreased by 30% over the 11 wks of farming (paper IIpaper IIpaper IIpaper II), and is probably even more greatly reduced when seagrass cover reaches the 15-20% seen after >10 yrs of farming (paper Ipaper Ipaper Ipaper I). Even though the total production is probably greater in farms due to the rapid growth of the farmed algae (Zemke-White and Smith 2006), most of this removed from the system through harvest. Second, the loss of seagrass cover undoubtedly reduces sediment erosion control, even though the presence of the seaweeds probably will dampen wave energy to some extent, and the loss of grain-forming Halimeda algae actually decreased mean grain size (paper Ipaper Ipaper Ipaper I). In areas like Paje and Jambiani (Zanzibar East coast) where farming was originally introduced, drift sand banks in farming areas could be an indirect result of sea-grass loss (N.S. Jiddawi, pers. comm.), but this requires more investigation before taken as a fact. Third, the results of paper IIIpaper IIIpaper IIIpaper III indicate that seaweed farms probably influences fish catches. For some fishery species, e.g. the seagrass rabbitfish Siganus sutor that often feeds on the farmed algae (Russell 1983; Bergman et al. 2001), the loss of seagrasses seems to be compensated for by the presence of the farmed seaweeds. Hence, seaweed farms could actually increase fish catches in vegetation-free areas, should possible effects on biodiversity and ecosystem functioning be carefully addressed (paper IIIpaper IIIpaper IIIpaper III). There are however a

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number of other aspects that must be taken into consideration: (1) the location of the farms, which prohibits fisheries during part of the tidal cycle, (2) the loss of important meio-, macro-fauna and seagrass epiphytes constitute a food source to many commercial fish species, (3) the fluctuating quality and size of the habitat (seaweeds) due to harvest, (4) the fact that nets cannot be used within farms, (5) the lack of knowledge about the importance of seagrass pres-ence in the landscape for fish catches in farms, and (6) property right issues and conflicts between seaweed farmers and fishermen (de la Torre-Castro 2006). Overall, we suggests that seaweed farms, at least in their current state, are not comparable to seagrass beds as fishing grounds, and are not suitable as fishing grounds per se. Cascading effects of fisheries in seagrass food webs

From the literature we see that overgrazing of submerged macrophytes by sea urchins, observed in e.g. coral reefs (McClanahan and Shafir 1990), temperate macro-algal reefs (Sala and Zabala 1996; Shears and Babcock 2002), temperate kelp beds (Steneck et al. 2002) and seagrass beds (Rose et al. 1999), has often been attributed to excessive removal of urchin predators. However, also other factors such as eutrophication, disease, presence of shelter, etc. have great influ-ence on urchin populations (Sala et al. 1998).

A review of the knowledge of causes, consequences and management of sea urchin overgrazing of seagrasses on a global scale (Eklöf et al. submitted) showed that while many studies discuss causes of overgrazing (e.g. overfishing), few explicitly investigate them. The three major categories of potential drivers were (1) increased recruitment due to changes in abiotic variables (e.g. water temperature), (2) reduced top-down control due to overfishing, and (3) eutro-phication stimulating urchin recruitment and feeding, of which only eutrophi-cation has been experimentally demonstrated to induce overgrazing (Tewfik et al. 2005).

In a seminal paper, Strong (1992) argued that trophic cascades are restricted to aquatic hard-bottom systems dominated by simple and poorly defended plants (macroalgae). Similarly, Pinnegar et al. (2000) suggested that cascading effects of fisheries are uncommon in soft-bottom systems because destructive fishing methods (e.g. trawls) mask indirect effects. At the time of these reviews, however, virtually no studies had investigated the presence of trophic cascades or cascading effects of fishing in vegetated soft-bottom systems. Since then, Silliman and Bertness (2002) demonstrated the importance of top-down regula-tion of salt marsh production, and it seems likely that grazing has a similarly important role in seagrass beds (Valentine and Duffy 2006).

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Results of our 20-year survey in protected and fished Kenyan reefs (experi-ment 1, paper IV paper IV paper IV paper IV), as well as a more in-depth study in four of these areas (ex-periment 2, paper IV paper IV paper IV paper IV) showed higher densities of T. gratilla in fished than protected areas. Since predation rates on tethered urchins was 1/3 as high in fished as in protected areas, we suggest that removal of urchin predators, by reducing predation control on urchins, contributes to increasing abundances of seagrass-feeding urchins in Kenyan seagrass beds. To our knowledge, this is the first study to confirm this pattern in seagrass systems, which has strong resem-blances to effects of fishing observed in hard-bottom systems (see table 6 in paper paper paper paper IIIIVVVV).

The major predators were surprisingly not finfish but asteroids, which could be due to the average large size of the urchins encountered and tethered (93% were larger than 50mm in test diameter), and the presence of seagrasses as shel-ter from visual predators such as fish (see below). There are currently no density estimates of these asteroid predators in the areas, but several of the dominating predatory species (e.g. Protoreaster linki) are collected for ornamental trade and use as bait in trap fisheries (Gossling et al. 2004), which could affect their dis-tribution outside protected areas.

Cross-habitat interactions seem to have a strong influence on seagrass food webs (Dorenbosch et al. 2005b; Valentine et al. 2007; Vanderklift et al. 2007), but we found no major effects of distance to patch reefs on any of the sampled variables. Supported by results of a recent study, demonstrating the overriding influence of shelter for predation by finfish (Vanderklift et al. 2007), we suggest that the presence of seagrass leaves in both Far and Close sites provided the tethered urchins with shelter, despite that densities of reef-associated predators probably decreases with increasing distance into seagrass beds (Dorenbosch et al. 2005b).

Shelter from predators is a highly important function of seagrasses, and Heck and Valentine (1995) indicated that sea urchin overgrazing of seagrasses may be self-regulated through a negative feedback loop where seagrass loss indirectly increases predation rates on urchins. The feedback should however only be valid when predators are functionally present, and therefore not control urchins in fished areas where predators are less abundant. Our results showed that the effect of shelter (seagrass) was dependent on an interaction between all three factors (fishing, presence of seagrass, and area). While the pattern in Mombasa Marine Park and the fished Diani and Bamburi viewed together confirmed the original hypothesis - shelter is important for decreasing predation in protected but not in fished areas - the opposite pattern (lower predation and higher ur-chin density in the absence of shelter) was found in Watamu Marine Park. To-gether with the ongoing sea urchin overgrazing in Watamu, this suggests that

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other factors like increased predation on intermediate predators in the absence of seagrass, the larger urchin size in Watamu, bottom topography, and possible eutrophication (T. R. McClanahan, pers. comm.) in this area are more impor-tant than predation control.

Another important aspect regarding the role of shelter was that in the two fished areas Diani and Ras Iwatine, the total urchin density and survival of teth-ered urchins (including non-predation related mortality) was lower in unvege-tated than vegetated sites despite no difference in predation rates. This suggests that seagrasses are important as shelter from other stressors as well (e.g. strong sunlight), and that a general ‘stressor feedback loop’ dependent on site-specific effects of seagrass loss on urchins, potentially regulates urchin overgrazing in protected as well as fished areas.

The link between fishing and urchin grazing on the two seagrass species Tha-lassodendron ciliatum and Thalassia hemprichii was more elusive, since no major effect of fishing was found (despite that the highest grazing index was found in fished areas). This could be due to high presence of urchins also within the protected Watamu area, but also the inadequacy of the sampling method in capturing the full effect of overgrazing (since the assays are dependent on sea-grasses still being present and not overgrazed). Furthermore, a previous study has suggested that differences in sensitivity to intensive grazing between Thalas-sodendron ciliatum and Thalassia hemprichii and (Alcoverro and Mariani 2005) results in T. ciliatum dominating in protected areas (with less urchins) and T. hemprichii dominating in fished areas with more urchins (McClanahan et al. 1994).

These results highlight some very interesting aspects of how fishing influences tropical seagrass systems, but also how little we currently know to draw any certain conclusions. Within a MASMA/WIOMSA-funded research program I, together with Swedish, Kenyan, Tanzanian and Mozambiqan colleagues con-tinue to investigate some of these aspects, including (1) the distribution, diver-sity and collection of invertebrate urchin predators like starfish, (2) patterns of urchin recruitment in time and space, (3) the effects of eutrophication on graz-ing rates and seagrass responses to grazing, and (4) with what success local man-agers deal with overgrazing by e.g. urchin removal. Effects of grazing regime and seagrass species on responses to grazing

The results of our simulated grazing study (paper Vpaper Vpaper Vpaper V) showed that in Thalassia hemprichii, the intensity and not frequency of grazing seemed to be important for growth responses. Even though there was no clear effect compared to the ungrazed controls (which confirms that T. hemprichii is relatively resistant to

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grazing (e.g. Alcoverro and Mariani 2005)), the difference in growth between the two intensities suggests that different grazing regimes still could have an effect on seagrass growth that could cascade through the food web.

Carbohydrates (sugar and starch) are stored as energy reserves primarily in sea-grass roots and rhizomes, and are known to be exhausted by various distur-bances such as changes in light climate (Alcoverro et al. 2001) and cropping (Dawes et al. 1979). The results of our simulated grazing showed that rhizome carbohydrate content in T. hemprichii was similarly affected mostly by grazing intensity, even though there seemed to be an interaction between intensity and frequency for starch. Results of an unpublished study, conducted together with Australian colleagues (Eklöf et al. unpublished) demonstrate that a 50% reduc-tion in carbohydrate levels severely affected responses to simulated grazing in another tropical/subtropical seagrass species (Halophila ovalis [R. Brown] Hooker f.) (Fig 4). Due to the fundamental role of carbohydrates in seagrasses as energy reserves (Touchette and Burkholder 2000), this suggests that inten-sively grazed T. hemprichii could be less resistant to additional disturbances such as grazing or shading.

7 15 21

% of initial shoot density

Control (UC) Leaf grazing (UL) Rhizome grazing (UR) Control (SC) Leaf grazing (SL) Rhizome grazing (SR)

Treatments

0

20

40

60

80

100

120

Unshaded Shaded

Fig 4. Results of a 3-wk field experiment on effects of carbohydrate loss due to shad-ing (4d with 20 % light), on the recovery of Halophila ovalis after two different forms of grazing: leaf removal (L) and rhizome disturbance (R), measured after 7, 15 and 21 days (mean + 1 S.E.). In short, recovery rate was lower in shaded plots where rhi-zomes were disturbed (SR) than in controls and unshaded one’s.

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The second species Enhalus acoroides showed no response in growth and car-

bohydrates to changes in either intensity or frequency of grazing. This differ-ence in comparison to the response in T. hemprichii could be due to differences in size, carbohydrate storage and translocation capabilities, and highlights an interesting aspect of seagrass species diversity. Differences between co-occurring species in responses to a similar disturbance, e.g. grazing in Kenyan seagrasses (Alcoverro and Mariani 2005), could indicate a higher ‘response diversity’ (Elmqvist et al. 2003) in these multispecific beds compared to monospecific ones. While some species are affected by grazing, others are not, which over time could buffer the total effects of overgrazing. Similar effects of genetic di-versity have been previously demonstrated in temperate mono-specific beds (Hughes and Stachowicz 2004; Reusch et al. 2005) but have not yet been ex-perimentally addressed at the species level. Due to the short duration of our study it is important to emphasize the need for future studies that increase the temporal and spatial scale to explore the applicability of the results. Regime shifts in seagrass ecosystems

Based on a literature review (paper VIpaper VIpaper VIpaper VI) we identify three major categories of shifts in seagrass systems. First, seagrass species shifts (from one dominating species or set of species to another) have been observed on a global scale, and mostly constitute shifts from typical ‘climax’ to ‘pioneer’-type species, driven by changes in environmental conditions like temperature, nutrient, light or grazing regime. Strictly speaking, these are community shifts but not true regime shifts, since the regime is still seagrass-dominated, and the change in environmental conditions - and not feedback mechanisms - often is the main factor preventing recovery. Nevertheless, they can affect ecosystem functioning if the new domi-nating species are less able to support services like fish production than the for-mer dominating one (e.g. Montefalcone et al. 2006).

Second, we identify alternative regimes (vs. seagrass) under constant environ-mental conditions: burrowing worms and shrimp. Examples from Europe, USA, the Caribbean and South Africa suggests that these bioturbating organ-isms can exclude seagrasses from areas where they are abundant by (1) destabi-lizing sediments, (2) mechanically disturbing seagrass roots and rhizomes, and (3) smothering seagrass leaves. At the same time, densities of these organisms are usually low within well-established seagrass beds because the seagrass roots and rhizomes prevent their burrowing. This is a classic example of how organ-isms with strong engineering traits form feedbacks that benefit them directly by

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altering abiotic conditions, and indirectly by excluding other organisms thereby reducing interspecific space competition.

Third, we identify shifts between seagrasses and alternative regimes (macro- and micro-algae and bivalves) driven by changes in anthropogenic disturbances (changes in environmental conditions, introduction of novel species, and altera-tions of seagrass food webs). Initially, seagrass loss will not only benefit other organisms competing for limiting resources, but also increase the rate of seagrass loss due to a positive feedback where loss of sediment stabilization increases turbidity. This change in feedback alone has been proposed to form an alterna-tive regime: unvegetated or bare sediment (van der Heide et al. 2007). It is however possible that this regime could be populated by benthic animals such as bivalves (see below).

The most common shift described in the literature is that from seagrass to al-gal dominance, driven either by (1) shifts from nutrient to light limitation due to eutrophication (Duarte 1995), (2) loss of grazer control of algae, possibly caused by overfishing of predatory fish (Heck Jr and Valentine 2007), and (3) invasion of non-native macroalgae, where the success of invaders is partly de-pendent on other stressors that weaken the competitive ability of seagrasses. These alternative algal regimes are often resilient to disturbances such as seagrass recovery and management interventions due to strong feedback mechanisms: increasing oxygen demand induces anoxia and sulfide stress on seagrasses (Borum et al. 2005), loss of habitat could negatively affect populations of mesograzers and top predators (Williams and Heck Jr 2001), and dissolved nutrients from decomposing seagrass and algal tissue fuels re-occurring algal blooms (Lavery and Mccomb 1991; Pihl et al. 1999; Troell et al. 2004).

Another potential shift is from seagrass to bivalves, with examples involving native blue mussels (Mytilus edulis) in Denmark (Fogh Vinther 2007), the in-vading Musculista senhousia in California, USA (Williams 2007) and the intro-duced oyster Crassostrea gigas in Willapa Bay, Washington (Buhle and Ruesink, unpublished) replacing Zostera marina L. These have so far not been discussed as regime shifts, but in all of the examples disturbances causing seagrass loss (eutrophication, fragmentation etc.) benefits dominance of bivalves (both by opening up space and reducing competition), which prevents seagrass recovery. Just as in the examples of ‘alternative stable states’ (see above), disturbances affecting one of the two dominating organisms (seagrass or bivalve) will in the long run benefit increasing dominance of the other.

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Direct and

indirect disturbance

Change in environ-

mental conditions

Low seagrass

biomass

High seagrass

biomass

Low abundance

of competing

organims

General mechanism for shifts betweenseagrass and alternative regime

(-)(-)

Seagrass

regime

Alternative

regime

High abundance

of competing

organims(+)

(+)

(+)

(+)

(-)

(-)

(-)

The general mechanism for all the shifts seems to be that disturbances or changes in environmental conditions that reduce seagrass abundance and fitness benefits associated organisms by reducing competition. If these organisms form biotic feedbacks that self-fuel their dominance and prevent seagrass recovery, the system can be argued to be in another regime (Fig 5).

Based on these observations, we pose three major questions. First, are these alternative communities actually true alternative regimes? The change in feed-back mechanisms is one key criterion that seems to be fulfilled, but we know very little about the spatial and temporal scale. One way to answer the question could be trying to force shifts from seagrass to an alternative regime in con-trolled experiments and monitor their persistence over time, which has been successfully done in forests and lakes (Scheffer and van Nes 2007). Second, do these alternative communities constitute a number of separate alternative re-gimes, or are they just example of a single ‘non-seagrass’ regime where the iden-tity of the community is a reflection of factors the site-specific conditions, dis-turbance, sequence of arrival, community identity prior to disturbance, etc.? A

Fig 5. Conceptual model of regime shifts from dominance of seagrass to other or-ganisms (paper VIpaper VIpaper VIpaper VI). Seagrasses normally dominate ecosystem functioning by con-trolling competing organisms and self-fuelling their dominance through positive feedback mechanisms. Disturbance and altered environmental conditions decreases seagrass abundance and fitness, which benefits competing organisms that gradually become dominating by the formation of novel feedbacks.

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simple form of evidence would be observations of shifts not only to but between these ‘alternative’ regimes. Regardless of whether they are true regimes or not, however, the third question is which factors drive shifts to specific alternative regimes or communities, and whether the answer can be used to predict or even prevent shifts? Based on the examples reviewed, we suggest that a combination of (1) site-specific conditions (e.g. nutrients, temperature, salinity, etc.), (2) com-munity identity (e.g. which seagrasses are present, presence of other engineering species like bioturbators or mussels, grazing sea urchins, etc.), (3) driver(s) (type and regime of disturbances, e.g. eutrophication, fishing or introduction of inva-sive species) and (4) the landscape (which systems are present in the landscape, in what regimes are they, and what is the degree of connectivity to the actual system of interest), all contribute to the risk for shifts to occur, and to what a sys-tem will shift.

The regime shift literature is rapidly growing, but there is currently a tendency to view regimes or states as either of two possible options, e.g. coral or macroal-gae on reefs, rooted macrophytes or phytoplankton in lakes, and seagrasses or macroalgae in seagrass beds, etc. (Walker and Meyers 2004). At the same time both theory and observations clearly suggest that ecosystems are complex adap-tive systems with a range of potential regimes (Vandermeer et al. 2004; Jasinski and Payette 2005; Scheffer and van Nes 2007). Besides that this is an interest-ing observation from a scientific point of view that highlights the extreme com-plexity involved, the existence of multiple regimes is highly important from a management perspective: by focusing on simply one potential shift, e.g. from seagrass to macroalgae, managers could simultaneously increase the risk for other shifts to occur, if we do not look for other types of shifts (see e.g. Fogh Vinther 2007).

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MAJOR IMPLICATIONS

The studies in this thesis address effects and mechanisms involved in different anthropogenic disturbances, and the results have direct implications for seagrass management. However, applied natural sciences can only aid managers to a certain extent: without s addressing the social and economic drivers that push systems into undesirable trajectories, the future will undoubtedly be bleak. Hence, with the risk of reaching outside my own area of expertise, I will also address some of the more important drivers in the case studies. Coastal management in an East African context

The results of the two case studies from East Africa viewed together suggest

that a shift from seagrass to farmed algae cascades up through the food web to fish catches (paper I, II, IIIpaper I, II, IIIpaper I, II, IIIpaper I, II, III), while removal of urchin predators cascades down through the food web and ultimately contributes to loss of seagrass production following overgrazing (paper IV, V, paper IV, V, paper IV, V, paper IV, V, Fig 5). Adding changes in basic environ-mental conditions such as nutrient, temperature, light and salinity regime, shifts in communities and feedbacks seems to drive shifts to alternative ecosystem regimes (paper VIpaper VIpaper VIpaper VI). This highlights that seagrass beds must be treated and managed as ecosystems where changes in important components or processes undoubtedly will influence the trajectory of the system as such. Increasing the sustainability of open-water seaweed farming

Without a doubt, seaweed farming is a preferred option to shrimp farming or dynamite fishing, but the questionable environmental aspects demonstrated here should still be addressed to improve the sustainability of the activity. Sea-weed farming must to a greater extent be incorporated into a holistic coastal management, as one out of many important subsistence activities. Practically, this could entail lessening impacts on particular benthic communities by re-stricting site choice and farming intensity; implementation of alternative farm-ing methods such as rafts (Lundsör 2002), floating long-lines (Hurtado and Agbayani 2002), and enclosed land-based polyculture systems (Qian et al. 1996; Wu et al. 2003) after the sustainability of such methods has been prop-erly investigated.

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Triggerfish,Wrasse

Starfish:

e.g. Protoreaster linki

Thalassodendron

ciliatum

HUMANS

?

(+) Reduced

predation

(-) Overgrazing

(-?)

Sea urchins:

e.g. Tripneustes gratilla

Seaweed

farming

Overgrazing

and fishing

Seagrass bed Farmed seaweeeds

FISHING AND COLLECTION

OF INVERTEBRATES

Invertebrate

in- and epi-fauna

Fish

(+) Reduced

predation

(-) Loss of

habitat

(-/+) Habitat

change

(-/+) Change in

fish catches

(-/+) Loss/gain

of fishing

grounds

(-)

(-) Shading

(-) Loss of

food

SEAWEED

FARMING

The socio-economic aspects of seaweed farming must also be addressed (see e.g. Sievanen et al. 2005), since many farmers are abandoning the activity due to low prices paid by the seaweed companies (Bryceson 2002; de la Torre-Castro and Rönnbäck 2004) and a recent attempt to increase prices by opening the market to competition seems to have failed (Forss et al. 2007). Metaphori-cally speaking, tropical seaweed farming on Zanzibar seems to be at a cross-roads: low prices indirectly diminish possible environmental side-effects, but

Fig 5. Conceptual model of how effects of seaweed farming and fishing cascade through the seagrass food web.

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brings low socio-economic sustainability. Irrespective of which direction sea-weed farming takes, there is a pressing need for holistic management. At the highest scale, this should entail addressing the socio-economic drivers behind seaweed farming: on one hand the needs for alternative subsistence activities aiding sustainable development, on the other the rapidly increasing need for cheap seaweeds to the global carrageenan industry. The different wants and needs of various stakeholders indicate a key question that I together with PhD candidate Åsa Forss from Södertörn University College currently address (Forss and Eklöf, unpublished): does seaweed farming contribute to sustainable devel-opment, or does it constitute another example of how extraction of a valuable natural resource brings more problems than it solves? We suggest that seaweed farming in some areas still has a long way to go, and currently does not fulfill the many promises associated with the activity. Managing overgrazing through fisheries

The results of the second case study suggests that fishing, besides directly af-fecting target populations, can cascade through seagrass food webs and ulti-mately affect seagrass production and distribution (paper IV, Vpaper IV, Vpaper IV, Vpaper IV, V). The much lower abundances of T. gratilla in protected vs. fished coral reefs further indi-cate that this also refers to fishing in adjacent systems, and that this issue must be viewed and managed on a seascape level.

One increasingly popular approach in fishery management is Marine Pro-tected Areas (MPAs), which come in various forms, shapes and degree of pro-tection. The results of paper IVpaper IVpaper IVpaper IV suggest that protection could reinforce preda-tion control on T. gratilla, but also demonstrate that MPAs with decades of protection still are vulnerable to overgrazing fronts of urchins. This could be the result of two mechanisms: either, these parks are too small to offer adequate protection, or other factors (e.g. eutrophication or increasing water tempera-ture) are the basic driver(s). First of all, the relative role of such factors must be thoroughly assessed, and this will partly be conducted within the above-mentioned MASMA project. If overfishing is indeed proven to be the major culprit, a simple ecological solution would be to extend the scale of protection. From a societal point of view this would of course be problematic, if not detri-mental, since fisheries provide income and food on the table for a large propor-tion of the coastal population. Since seagrass beds are part of a larger social-ecological system, MPA management must involve both ecological and social variables (Jentoft et al. 2007). One approach could be ‘bottom-up’-implemented reserves allowing some fisheries, that interestingly provide equally good protection as closed ‘top-down implemented’ parks (McClanahan et al.

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2006). It is however unknown to what extent they will protect seagrasses from overgrazing.

At the highest scale, successful management must also address the underlying socio-economic factors driving overfishing (e.g. coastal migration, increasing rates of unemployment, demand for fish from the rapidly expanding coastal population and tourist industry, etc.), to ensure any long-term and large-scale success. Addressing complexity in seagrass management

One of the explicit aims of our literature synthesis on regime shifts (ppppaaaaper per per per VIVIVIVI) was to address if and how some common seagrass management strategies – protection, pollution control, seagrass transplantation, and monitoring - take regime shifts into account.

In terms of protection using MPAs, seagrasses are included in many parks (Green and Short 2003), but local success seems to be ultimately dependent on site-specific conditions and the lack of diffuse disturbances on a regional scale, rather than enforcement (Marba et al. 2002). This suggests that while protec-tion can be useful to diminish effects of small-scale direct mechanical distur-bances, other more holistic approaches such as water quality control, legal pro-tection of seagrass habitats, fishery policies and control of invasive species, are ultimately needed to provide adequate protection.

Seagrass transplantation is in some areas like the USA and Australia a com-monly used strategy to restore seagrass beds, but has had a generally low suc-cess rate (Campbell 2002). An often overlooked but probably common prob-lem in restoration projects per se is that the ecosystem may in fact have shifted to an alternative regime, and that management in fact includes the breaking of undesirable feedback mechanisms and strengthening of desirable ones (Suding et al. 2004). From a seagrass perspective, we highlight the many difficulties regarding ecosystem restoration, where complex issues such as feedbacks and hysteresis effects must be addressed. For instance, poor water quality is often a major driver behind seagrass decline and shifts to macroalgae, and two exam-ples from Northern Europe highlight how water quality management failed to restore seagrass beds due to strong feedback mechanisms between unwanted organisms (Munkes 2005; Fogh Vinther 2007). When hysteresis effects occur, it must be investigated whether water control to adequate levels are economi-cally feasible. If the problem is cascading effects of fishing, it may be warranted to reintroduce to predators or remove intermediate predators, which has been successfully conducted in lakes (Meijer et al. 1999) but not in seagrass beds.

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Finally, monitoring of seagrass health is often focused on coarse variables such as depth limits, percent cover or standing biomass. While this may ade-quate to monitor change in water quality and general coastal ecosystem health (Dennison et al. 1993), there is a risk that once changes at this level occurs, a regime shift is already on its way and seagrass decline is inevitable. If monitor-ing should aid managers in detecting shifts in time to respond before seagrass decline occurs, more sensitive response variables (some seagrass-based ones are already being tested, see paper VIpaper VIpaper VIpaper VI) must be introduced, which will require a lot more detailed and costly monitoring. Since some of the feedbacks that seem to be important for a functional seagrass regime are dependent on other organ-isms, e.g. mesograzers controlling macroalgae (Hughes et al. 2004), top preda-tors controlling intermediate predators (Heck Jr and Valentine 2007) and sea urchin (paper paper paper paper IIIIVVVV), sponges (Porifera) controlling phytoplankton blooms (Peterson et al. 2007), evaluation of their respective abundances and popula-tion trends could in some cases provide an assessment of the sensitivity of the system.

In conclusion, changes in feedback mechanisms and shifts between dominat-ing species could be an important part of seagrass decline in some areas. This does not entail that linear seagrass decline due to changes in environmental conditions is less important, or that regime shifts will explain all cases of sea-grass loss, only that it is one potential explanation that due to its implications for management must be considered. Due to the high degree of uncertainty (not knowing if, when and to what a system will shift) and the high costs in-volved in restoring seagrass beds, some final key points emerge: (1) Seagrass research and management should adopt more of a holistic system view, where seagrasses are viewed as one component in social-ecological systems affected by variability and change in biotic, abiotic, social, economic and political drivers; (2) management must be adaptive by viewing management actions as tests in an ever-changing world, since rigid programs strictly targeting certain distur-bances could decrease the chance of managing future unexpected change; and (3) we must strive to maintain biodiversity from gene to landscape level as insurance for continued ecosystem functioning in the face of unexpected change.

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SAMMANFATTNING PÅ SVENSKA

Mänskliga störningar i sjögräsängar: Vad betyder de för kustzonsförvaltning?

En doktorsavhandling i Marin Ekotoxikologi

Johan S. Eklöf

Systemekologiska Institutionen, Stockholms universitet

Sjögräs är marina (havslevande) blomväxter som bildar mer eller mindre täta bestånd (sjögräsängar) i grunda kustområden världen över. Genom snabb tillväxt som ger föda åt växtätare och strukturell komplexitet som ger habtat och skydd, hyser sjögräsängar generellt en högre artrikedom och individtäthet av andra organismer (t ex alger, musslor, snäckor, maskar, kräftdjur och fiskar) än vegetationsfria bottnar. Eftersom många av de associerade organismerna (t ex fisk) har ett kommersiellt värde, har sjögräsängar ett indirekt mycket högt ekonomiskt värde helt klart jämförbart med det hos t ex tropiska korallrev.

Trots sin betydelse är sjögräsängarna hotade över hela världen, och har minskat 15% i utbredning under de senaste 20 åren. De största hoten är muddring och övergödning, men på senare år har man även insett att intensivt fiske verkar kunna påverka dynamiken i systemen, om fiskarna som tas bort är funtionellt sett viktiga (t ex rovdjur eller betare) i näringsväven. På sikt gör detta att vi riskerar att inte bara utarma dessa värdefulla system, utan i slutändan förlora viktiga tjänster som fiske och kustzonsskydd. Detta är speciellt oroande i tropiska utvecklingsländer, eftersom sjögräsängar där indirekt förser fattiga människor med billig mat i form av fisk.

Mot denna bakgrund var huvudsyftet med min avhandling att undersöka hur nyttjandet av två olika resurser knutna till sjögräsängar i Östafrika – algodling och kustnära småskaligt fiske – påverkar sjögräsängar och produktionen av ekologiska varor och tjänster (t ex fiske), och i slutändan diskutera vad detta innebär för lokal kustzonsförvaltning.

Min första fallstudie handlar om algodling, som introducerades som en alternativ inkomstkälla i Tanzania (Östafrika) på slutet av 1980-talet. Två arter

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tropiska rödalger (Euchema och Kappaphycus) odlas i grunda havsvikar, torkas på land, och exporteras sedan till fabriker i Europa, USA och Asien. Där utvinner man sockerarten karragen som används som stabiliseringsmedel i mat, smink och läkemedel. T ex hittar man i svenska butiker karragen i hamburgerkött, läppstift och i glass.

Eftersom algodling inte kräver tillsats av gödnings- eller bekämpningsmedel utmålas aktiviteten ofta som en av de mest hållbara typerna av vattenbruk. Många algodlingar placeras dock på bottnar där sjögräs normalt växer, och detta skulle kunna påverka sjögräsen och därigenom hela ekosystemet. I min första studie (artikel I) jämförde vi utbredningen av sjögräs i områden med och utan algodlingar, och visade att det generellt fanns mindre sjögräs under algodlingarna. Därtill fanns det färre arter och individer av olika små ryggradslösa djur (maskar, kräftdjur och musslor) under odlingarna. Eftersom att vi inte kunde bevisa att detta berodde på odlingarna i sig byggde vi upp egna experimentella algodlingar i en sjögräsäng, och följde utvecklingen under 11 veckor (artikel II). Resultatet var entydigt: under algerna fick sjögräsen mindre solljus, växte långsammare, och minskade med 30% i utbredning.

Eftersoms sjögräsängarna i studieområdet utgör viktiga fiskeplatser misstänkte vi att algodlingen indirekt skulle kunna påverka fisket. M h a en lokal fiskemetod (betade fiskfällor) jämförde vi fiskfångster från fällor placerade inne i algodlingar med fångster från fällor placerade i sjögräsängar och på ren sand (artikel III). Resultaten visade lite oväntat att fångsterna från fällor i algodlingen var lika stora som i de från sjögräsängen, men större än i fällorna placerade på ren sand. Detta skulle kunna betyda att om även om sjögräs försvinner, så kan algerna attrahera fiskar och därigenom stödja ett visst fiske. Däremot fanns skillnader mellan sjögräsområdet och algodlingen i vilka fiskarter som fångades, eftersom algerna verkar attrahera vissa fiskar mer än andra. Dessutom förändras fiskarnas habitat (algerna) drastiskt när algerna skördas, och på många platser finns konflikter mellan algodlare och fiskare eftersom näten ofta trasslar in sig i algerna. Sammantaget tyder detta på att algodlingar i dagsläget inte kan ersätta sjögräsängar som fiskeområden, och att en expansion av algodling skulle leda till minskade fiskeområden.

Min andra fallstudie handlade om indirekta effekter av kustnära småskaligt fiske, som är den i särklass viktigaste inkomstkällan för kustbefolkningen i Östafrika. Brist på andra sysselsättningar och förbättrade fiskemetoder har under de senaste årtiondena lett till ett ökat fisketryck, vilket gett minskade fångster men samtidigt ökad efterfrågan på fisk. I brist på en fungerande fiskeförvaltning har detta lett till en ’ond’ spiral, där fiskbestånden minskar i allt snabbare takt.

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I flera områden längs Kenyas och Tanzanias kuster har man observerat hur onormalt täta bestånd av sjöborrar har betat ner hela sjögräsängar, vilket enligt lokalbefolkningen lett till minskade fiskfångster. Orsaken till den stora mängden sjöborrar skulle kunna vara ett intensivt fiske av rovfiskar som äter sjöborrar, men även övergödning och förändringar i vattentemperatur.

För att testa hypotesen om indirekta effekter av fiske undersökte vi först utbredningen av dessa sjöborrar i sju områden längs den Kenyanska kusten mellan 1987 och 2006 (artikel IV). Resultaten visade att antalet sjöborrar per yta var betydligt mycket högre i fiskade områden än inne i marina nationalparker där fiske är förbjudet, vilket tyder på att fisket är en betydande faktor. För att bekräfta denna hypotes utförde vi sedan en detaljstudie i två marina nationalparker och två fiskade områden. Resultaten bekräftade delvis vår hypotes: i de fiskade områdena åts en tredjedel så många sjöborrar upp av rovdjur (sjöstjärnor och fiskar) per tidsenhet som i de två skyddade områdena, vilket resulterade i att vi fann tre gånger fler sjöborrar där. Grovt sett återspeglades detta även i en skillnad i hur mycket sjöborrarna betade på de två olika sjögräsarterna. Dock fann vi att frånvaron av sjögräs – vilket på flera platser var ett resultat av det stora antalet betande sjöborrar – verkar vara en betydande faktor, eftersom sjögräsen inte bara utgör föda utan även skydd mot rovdjur. Förlusten av sjögräs verkar i sig kunna dämpa mängden sjöborrar och därigenom deras destruktiva betning på sjögräsen. Detta är ett exempel på en viktig s.k. återkopplingsmekanism, som ’buffrar’ effekten av störningar i ekosystemet och bidrar till att bibehålla sjögräs över långa tidsrymder.

Vi utförde även en fältstudie där olika typer av betning (hur ofta och hur mycket som betas) simulerades m h a klippning (artikel V). Resultaten visade att förändringar i ’betningsregim’ från lätt till intensiv betning (hur mycket som betas), men inte hur ofta betningen skedde, minskade tillväxt och inlagring av kolhydrater hos en av de dominerande sjögräsarterna, samtidigt som den andra arten inte påverkades. Detta tyder på att den relativt höga diversiteten av sjögräs som återfinns i dessa ängar kan ha en ’buffrande effekt’: även om vissa arter påverkas negativt av betningen, så kan andra mindre känsliga arter ta över deras roll.

Givet resultaten från dessa två fallstudier blev jag intresserad av hur människan påverkar sjögräsängar på en ekosystemnivå, framförallt i fråga om skiften från t ex sjögräs till odlade alger, eller från mycket sjögräs till mycket sjöborrar. Därför genomförde jag tillsammans med en kollega en litteraturstudie över skiften i sjögräsängar på en global skala (artikel VI). Vår sammanställning visade att sjögräs som försvinner pga mänskliga aktiviteter i vissa fall ersätts av andra arter som t ex invaderande alger, musslor och grävande kräftdjur. Detta sker delvis för att de är bättre anpassade till de nya förhållandena, men också pga mindre

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konkurrens från sjögräsen om t ex solljus och utrymme. Många av dessa nya organismer påverkar i sig levnadsförhållanden som vattenkvalité och sedimentstruktur, vilket skapar nya ’återkopplingsmekanismer’ som gynnar de nya arterna men samtidigt missgynnar sjögräsen. Detta kan i värsta fall få systemet att ”gå i baklås”, där den naturliga återväxten av sjögräs effektivt förhindas. Dessutom försvårar detta olika förvaltningsstrategier som sjögrästransplantering eller vattenrening, som annars skulle kunna få systemet på rätsida igen.

Sammanfattningsvis visar resultaten av mina studier att även småskaliga aktiviteter i utvecklingsländer som algodling och fiske kan leda till storskaliga förändringar i kustnära ekosystem som sjögräsängar, om skalan i nyttjandet överskrider kritiska gränser. Lösningen till denna problematik är troligtvis mångfacetterad, och innefattar bl a (1) en helhetstänkande samförvaltning av ekosystem på landskapsnivå, (2) teknologiska förändringar vad gäller odlingsmetoder, fiskeredskap och fiskekvoter, (3) en diversifiering av resursnyttjande hos lokalbefolkingen (dvs att man breddar sig och inte bara sysslar med fiske eller algodling) (4) nya övervakningsmetoder för att utvärdera ’ekosystemhälsa’ hos sjögräsängar baserat på arter med nyckelfuntioner, samt (5) reglering av nationella och internationella drivkrafter (t ex den globala algodlingsindustrin) som bidrar till att öka trycket på dessa viktiga ekosystem.

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ACKNOWLEDGEMENTS I wish to acknowledge the people of Chwaka and Marumbi villages in Chwaka Bay, Zanzibar (Tanzania), without whose hospitality, interest and assistance much of the field work for this thesis would not have been possible. The staff at the Institute of Marine Sciences (Zanzibar, Dar Es Salaam Univer-sity, Tanzania), especially director Dr A. Dubi, Dr N.S. Jiddawi, Dr M. Kye-walyanga, Mr M. M. Manzi, Mr S. Yahya, Mr U. A. Makame, Mr T. Buluda, Mr O. Amir and the technical staff Mr M. M. Mwadini, Mrs K. U. Said and Mr C. A. Mahawi, are deeply thanked for their full support during my stays. The Kenyan Marine Fisheries and Research Institute (KMFRI, Mombasa, Kenya), especially Dr J. Kazungu, Dr J. N. Uku and Mr A. Kimathi, and the Kenya Wildlife Service (KWS) are deeply thanked for institutional and practical support. Financial support was provided by the K & A Wallenberg, J.A. Letterstedt’s and A. Wilhelmina Memory stipend foundations, Stockholm Marine Research Cen-tre (SMF), and the Minor Field Study (MFS) scholarships provided by Sida (Swedish International Development Cooperation Agency).

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