Anthropogenic Disturbances and Shifts in Tropical Seagrass Ecosystems

Anthropogenic Disturbances and
Shifts in Tropical Seagrass Ecosystems
Johan S. Eklöf
Doctoral Thesis in Marine Ecotoxicology
Department of Systems Ecology
Stockholm University
Stockholm, Sweden
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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ABSTRACT
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 disturbances, 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 organisms (macroalgae, bivalves, burrowing shrimp, polychaetes, etc.). The formation of
novel feedback mechanisms can make these regimes resilient to disturbances like seagrass recovery and transplantation projects. Overall, this suggests that resource use activities 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: aquaculture; East Africa; ecosystem change; feedback mechanisms; Kenya;
management; overgrazing; regime shifts; resilience; seagrass; seaweed farming; sea urchins; 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 bottom 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 finished 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 Stadlinger: 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 tremendous 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 counting… 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 water, 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 sharing thoughts and opinions that have driven me further and tickled my research interest.
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
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INTRODUCTION
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SEAGRASS BEDS IN A CHANGING WORLD
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Seagrasses from species to seascape
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Seagrasses support vital services to society…
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…but are threatened by anthropogenic disturbances
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Seagrasses and disturbances in East Africa
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MAJOR AIMS OF THESIS
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METHODS
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GENERAL RESULTS AND DISCUSSION
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Effects of seaweed farming on seagrass ecosystems
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Cascading effects of fisheries in seagrass food webs
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Regime shifts in seagrass ecosystems
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MAJOR IMPLICATIONS
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Coastal management in an East African context
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Addressing complexity in seagrass management
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SAMMANFATTNING PÅ SVENSKA
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ACKNOWLEDGEMENTS
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REFERENCES
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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:
I
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.
II
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.
III
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.
IV
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.
V
Eklöf JS, Gullström M, Björk M, Asplund M, Dahlgren A, Hammar L,
Öhman MC. (Manuscript). Physical responses of two co-occurring seagrasses to different grazing regimes. In review for Aquatic Botany.
VI
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 development (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 effects on catches (e.g. Jiddawi and Ohman 2002; Stobutzki et al. 2006), removal
can cascade through food webs and cause major habitat changes like overgrazing of macrophytes and bioerosion of corals (see Pinnegar et al. 2000 for review).
When the resource outtake (e.g. fisheries) exceeds sustainable levels the pressure 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 unexpected 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 declining 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 anthropogenic 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 seagrass 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 angiosperms (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 distribution 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 feedback mechanisms as ‘ecosystem engineers’ (sensu Jones et al. 1994), most importantly 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, III),
III primarily due to an extraordinarily
high rate of primary production (Duarte and Chiscano 1999) supporting secondary production (Mateo et al. 2006); provision of a three-dimensional structure 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, between 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 exchange 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 importance 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, anthropogenic 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): although present in only 0.15% of the ocean surface, seagrasses (Smith 1981) and
their epiphytes (Moncreiff and Sullivan 2001) are highly important contributors to the primary production in the global oceans, which supports a substantial 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 sediment 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 valuable 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 disturbance, and alterations of natural disturbance regimes (Short and WyllieEcheverria 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 seagrass loss. The most important is decreased light penetration caused by increased 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. crustaceans) 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 eutrophication, 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 predation 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 disturbances, that synergistically affect ecosystem functioning (Lotze et al. 2006;
Orth et al. 2006). A growing body of literature suggest that anthropogenic disturbances (e.g. eutrophication) could cause shifts to alternative ‘regimes’,
‘phases’ or ‘stable states’ where other organisms like macroalgae dominate ecosystem functioning (Duarte 1995; Gunderson 2001; Munkes 2005; Valentine
and Duffy 2006). After such shifts, ‘undesirable’ feedback mechanisms can selffuel 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 therefore 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 TorreCastro 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 seagrass areas like Chwaka Bay (Zanzibar, Tanzania), this is illustrated by seagrass
beds being the most preferred fishing grounds, and seagrass-associated fish being 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 African 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 carrageenan (FAO 2002), a polysaccharide used as a thickening agent in food, cosmetics, 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 compete with other habitat-forming organisms for light and space, and also affect
community composition of associated organisms by attracting or deterring mobile species (Zemke-White and Smith 2006).
Seaweed farms are placed in seagrass beds where vegetation-free areas are lacking or where farmers believe that seagrasses fertilize seaweeds (de la TorreCastro and Rönnbäck 2004). Some farmers initially remove seagrasses to simplify 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 services 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),
15
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 institutions (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 decreasing 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 contributes 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 overriding 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
16
(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 seagrass 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 reduced 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 potential interaction between the intensity and magnitude of grazing on seagrass beds
in general.
17
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 function, 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?
Resource extraction
& management
ser
vic
es
(III
)
Managing
seagrass loss (VI)
Eco
sys
tem
Dis
t
Overgrazing
and fisheries (IV, V)
urb
an
c
es
Seaweed
farming (I, II, III)
Seagrass
regime
Desirable
feedbacks (IV, VI)
Regime shifts (VI)
Feedbacks
(VI)
Alternative
regime
Undesirable
feedbacks (VI)
Fig 1. Conceptual model of thesis, highlighting the topics of the different papers.
18
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 Tanzanian 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 ranging 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.
A
C
D
N
4 km
B
Kenya
4˚S
Tanzania
Zanzibar
200 km
42˚E
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.
19
Effects of seaweed farming
In the first study (paper
paper I)
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 studies on effects of seaweed farming at this time (2004) were based on comparisons
between farms and control sites (including paper I and III),
III there was an immediate need to experimentally validate previously observed patterns and identify mechanism(s) behind effects. In the second study (paper
paper II),
II the effects of
seaweed farming on a mixed seagrass community were experimentally investigated over 11 wks in replicated plots in three treatments: seaweed farms, controls, 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, sampled 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
paper I and II),
II farming effects could cascade to fish communities (Bergman et al. 2001) and ultimately fishery catches. In the third
study (paper
paper III)
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
paper IV).
IV The choice of study area was
based on the presence of (1) comparable seagrass beds, (2) several MPAs interspersed 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 investigated 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) focused on effects and interactions of three factors presumed to affect T. gratilla:
20
(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 modified 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). Finally, urchin grazing pressure on seagrasses was assessed by a natural herbivore
assay (Alcoverro and Mariani 2004), in which shoots of two dominating seagrass 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 Mariani 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 suggest 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 however 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 investigated how different combinations of grazing intensity and frequency (using
leaf clipping) affected shoot growth and rhizome carbohydrates in two cooccurring seagrass species, in this case T. hemprichii and Enhalus acoroides (p
paper V).
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 constraints 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.
21
Regime shifts in seagrass beds (paper
paper VI)
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 function, 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 discuss 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.
22
GENERAL RESULTS AND DISCUSSION
Effects of seaweed farming
The results of paper 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 generally disappear after a few months of farming (de la Torre-Castro and Rönnbäck
2004), we attributed these differences to effects of farming. This was corroborated by the experimental farming (paper
paper II),
II which reduced aboveground seagrass 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 seagrass 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 II)
II caused the observed patterns. Given
that seagrasses underneath real farms are also subjected to other farmingassociated 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 II contributed to
the overall differences observed in paper I , 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 remain within farms after a decade (15-20% cover, although mostly between farm
plots, paper I)
I indicates that a total seagrass loss is probably not likely when
farm intensity is kept at moderate levels. However, even with seagrasses remaining 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, fragmentation 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).
23
a) Invertebrate infauna
b) Fish catches
Fig 3. Community structure of (1) invertebrate infauna (biomass of major taxa,
paper I)
I and (b) fish catches (biomass per species, paper II),
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.
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
paper I).
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 (p
paper II).
II For
macroscopic epifauna (>2cm, paper II)
II and fish (based on catches using dema
traps, paper III)
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 (p
pa per II),
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
paper I, III).
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
paper III]
paper II])
III and the sea urchin Echinometra mathei [paper
II are attracted to
24
the algae either as a food source or shelter (Neish 2003). In addition, mechanical disturbance in farms could also deter organisms. Overall, the effects of seaweed 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 effects on functions performed by associated organisms (e.g. grazing, predation,
etc.). For instance, lucinid bivalves, the single taxon most affected by the presence of seaweed farms (paper
paper I),
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 ecosystem 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
paper II),
II and is probably even more greatly
reduced when seagrass cover reaches the 15-20% seen after >10 yrs of farming
(paper
paper I).
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
paper I).
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 seagrass loss (N.S. Jiddawi, pers. comm.), but this requires more investigation
before taken as a fact. Third, the results of paper 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
paper III).
III There are however a
25
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 presence 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 influence 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) eutrophication stimulating urchin recruitment and feeding, of which only eutrophication 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 regulation of salt marsh production, and it seems likely that grazing has a similarly
important role in seagrass beds (Valentine and Duffy 2006).
26
Results of our 20-year survey in protected and fished Kenyan reefs (experiment 1, paper IV),
IV as well as a more in-depth study in four of these areas (experiment 2, paper IV)
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 resemblances to effects of fishing observed in hard-bottom systems (see table 6 in
paper IV ).
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 shelter 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 distribution 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 urchin density in the absence of shelter) was found in Watamu Marine Park. Together with the ongoing sea urchin overgrazing in Watamu, this suggests that
27
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 important 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 tethered urchins (including non-predation related mortality) was lower in unvegetated 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 Thalassodendron 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 seagrasses still being present and not overgrazed). Furthermore, a previous study
has suggested that differences in sensitivity to intensive grazing between Thalassodendron 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 continue to investigate some of these aspects, including (1) the distribution, diversity and collection of invertebrate urchin predators like starfish, (2) patterns of
urchin recruitment in time and space, (3) the effects of eutrophication on grazing rates and seagrass responses to grazing, and (4) with what success local managers 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
paper V)
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
28
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 seagrass roots and rhizomes, and are known to be exhausted by various disturbances 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% reduction 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 intensively grazed T. hemprichii could be less resistant to additional disturbances such
as grazing or shading.
120
7
15
21
% of initial shoot density
100
80
60
40
20
0
Control (UC)
Leaf grazing (UL) Rhizome grazing (UR)
Control (SC)
Unshaded
Leaf grazing (SL)
Rhizome grazing (SR)
Shaded
Treatments
Fig 4. Results of a 3-wk field experiment on effects of carbohydrate loss due to shading (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 rhizomes were disturbed (SR) than in controls and unshaded one’s.
29
The second species Enhalus acoroides showed no response in growth and carbohydrates to changes in either intensity or frequency of grazing. This difference 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 diversity have been previously demonstrated in temperate mono-specific beds
(Hughes and Stachowicz 2004; Reusch et al. 2005) but have not yet been experimentally 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
paper VI)
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 dominating species are less able to support services like fish production than the former dominating one (e.g. Montefalcone et al. 2006).
Second, we identify alternative regimes (vs. seagrass) under constant environmental conditions: burrowing worms and shrimp. Examples from Europe,
USA, the Caribbean and South Africa suggests that these bioturbating organisms can exclude seagrasses from areas where they are abundant by (1) destabilizing 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 organisms with strong engineering traits form feedbacks that benefit them directly by
30
altering abiotic conditions, and indirectly by excluding other organisms thereby
reducing interspecific space competition.
Third, we identify shifts between seagrasses and alternative regimes (macroand micro-algae and bivalves) driven by changes in anthropogenic disturbances
(changes in environmental conditions, introduction of novel species, and alterations 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 alternative 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 algal 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 dependent 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 invading Musculista senhousia in California, USA (Williams 2007) and the introduced 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.
31
General mechanism for shifts between
seagrass and alternative regime
(-)
(+)
(-)
Direct and
High seagrass
indirect disturbance
biomass
Low seagrass
biomass
(+)
(-)
(-)
Low abundance
of competing
organims
Change in environmental conditions
(+)
Seagrass
regime
(-)
High abundance
of competing
organims
(+)
Alternative
regime
Fig 5. Conceptual model of regime shifts from dominance of seagrass to other organisms (paper
paper VI).
VI Seagrasses normally dominate ecosystem functioning by controlling 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.
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 feedback 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 controlled 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 regimes, or are they just example of a single ‘non-seagrass’ regime where the identity of the community is a reflection of factors the site-specific conditions, disturbance, sequence of arrival, community identity prior to disturbance, etc.? A
32
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) community 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 invasive 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 system 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 macroalgae 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 adaptive 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 interesting observation from a scientific point of view that highlights the extreme complexity 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).
33
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
paper I, II, III),
III while removal of urchin predators cascades down
through the food web and ultimately contributes to loss of seagrass production
following overgrazing (paper
paper IV, V, Fig 5). Adding changes in basic environmental conditions such as nutrient, temperature, light and salinity regime, shifts
in communities and feedbacks seems to drive shifts to alternative ecosystem
regimes (paper
paper VI).
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. Seaweed 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 restricting site choice and farming intensity; implementation of alternative farming 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 properly investigated.
34
Seaweed
farming
Overgrazing
and fishing
HUMANS
FISHING AND COLLECTION
OF INVERTEBRATES
(-)
(-/+) Change in
fish catches
(-/+) Loss/gain
of fishing
grounds
Triggerfish, Wrasse
?
(+) Reduced
predation
(-?)
Starfish:
e.g. Protoreaster linki
Fish
(-/+) Habitat
change
(+) Reduced
predation
(-) Loss of
food
Sea urchins:
e.g. Tripneustes gratilla
Invertebrate
in- and epi-fauna
(-) Overgrazing
(-) Loss of
habitat
SEAWEED
FARMING
(-) Shading
Seagrass bed
Thalassodendron
ciliatum
Farmed seaweeeds
Fig 5. Conceptual model of how effects of seaweed farming and fishing cascade
through the seagrass food web.
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 TorreCastro and Rönnbäck 2004) and a recent attempt to increase prices by opening
the market to competition seems to have failed (Forss et al. 2007). Metaphorically speaking, tropical seaweed farming on Zanzibar seems to be at a crossroads: low prices indirectly diminish possible environmental side-effects, but
35
brings low socio-economic sustainability. Irrespective of which direction seaweed 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 development, 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 affecting target populations, can cascade through seagrass food webs and ultimately affect seagrass production and distribution (paper
paper IV, V).
V The much
lower abundances of T. gratilla in protected vs. fished coral reefs further indicate 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 Protected Areas (MPAs), which come in various forms, shapes and degree of protection. The results of paper IV suggest that protection could reinforce predation 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 temperature) 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 abovementioned 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 detrimental, since fisheries provide income and food on the table for a large proportion of the coastal population. Since seagrass beds are part of a larger socialecological 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.
36
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 (p
pa per
VI) was to address if and how some common seagrass management strategies –
VI
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 protection can be useful to diminish effects of small-scale direct mechanical disturbances, other more holistic approaches such as water quality control, legal protection 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 commonly used strategy to restore seagrass beds, but has had a generally low success rate (Campbell 2002). An often overlooked but probably common problem 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 examples 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 economically 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.
37
Finally, monitoring of seagrass health is often focused on coarse variables
such as depth limits, percent cover or standing biomass. While this may adequate 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 monitoring 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 VI)
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 organisms, e.g. mesograzers controlling macroalgae (Hughes et al. 2004), top predators controlling intermediate predators (Heck Jr and Valentine 2007) and sea
urchin (paper
paper I V ), sponges (Porifera) controlling phytoplankton blooms
(Peterson et al. 2007), evaluation of their respective abundances and population trends could in some cases provide an assessment of the sensitivity of the
system.
In conclusion, changes in feedback mechanisms and shifts between dominating 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 seagrass 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 involved 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 disturbances 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.
38
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
39
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.
40
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
41
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.
42
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 University, Tanzania), especially director Dr A. Dubi, Dr N.S. Jiddawi, Dr M. Kyewalyanga, 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 Centre (SMF), and the Minor Field Study (MFS) scholarships provided by Sida
(Swedish International Development Cooperation Agency).
43
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