Anthropic impacts in Mediterranean Marine Protected Areasrua.ua.es/dspace/bitstream/10045/18846/1/master_Luna.pdfDesprés voldria agrair als meus directors per haver-me donat l’oportunitat

Anthropic impacts in Mediterranean Marine Protected Areas

Beatriz Luna i Pérez

www.ua.es

www.eltallerdigital.com

AnthropicAnthropic

impactsimpacts in in

MediterraneanMediterranean

Marine Marine

ProtectedProtected

AreasAreas

Beatriz Luna i Pérez

Ph Thesis

UNIVERSITAT D’ALACANT

Departament de Ciències del Mar i Biología Aplicada

Departament de Ciències del Mar i Biologia Aplicada

Departamento de Ciencias del Mar y Biología Aplicada

“Athropic impacts in Mediterranean Marine Protected Areas”

“Impactes antròpics en Àrees Marines Protegides Mediterrànies”

Memòria presentada per optar al grau de Doctora, menció de

Doctora Europea, en la Universitat d’Alacant per

BEATRIZ LUNA i PÉREZ

ALACANT, Juny 2010



“Athropic impacts in Mediterranean Marine Protected Areas”

“Impactes antròpics en Àrees Marines Protegides Mediterrànies”

Memòria presentada per optar al grau de Doctora, menció de

Doctora Europea, en la Universitat d’Alacant per

BEATRIZ LUNA i PÉREZ

ALACANT, Juny 2010



TESI DOCTORAL


Impactes Antròpics en Àrees Marine Protegides Mediterrànies

Beatriz Luna i Pérez Juny 2010

VºBº Directors de la Tesi Doctoral

Dr. José Luis Sànchez Lizaso Dr. Carlos Valle Pérez Dr. Just T. Bayle Sempere



TESI DOCTORAL


Impactes Antròpics en Àrees Marine Protegides Mediterrànies

Beatriz Luna i Pérez 2010

VºBº Director del Departament de Ciències del Mar i Biologia Aplicada

Jose Luis Sánchez Lizaso

Els doctors JOSÉ LUIS SÁNCHEZ LIZASO, CARLOS VALLE PÉREZ I JUST T.

BAYLE SEMPERE, Professors del Área de Zoología de la Universitat d’Alacant,

CERTIFIQUEN:

Que la memòria de Tesi doctoral titulada “Anthropic impacts in Mediterranean Marine

Protected Areas”, presentada per BEATRIZ LUNA i PÉREZ, ha estat realitzada baix la

seva direcció al Departament de Ciències del Mar i Biologia Aplicada de la Universitat

d’Alacant. Per que conste als efectes oportuns, signen a Alacant el dos de Juny de l’any dos

mil deu.

Dr. José Luis Sánchez Lizaso Dr. Carlos Valle Pérez Fdo: Dr. Just T. Bayle Sempere



Agraïments - Acknowledgements

En primer lloc, voldria agrair a la meva família haver-me ensenyant des de sempre a tindre somnis i a

lluitar per ells, amés del seu suport incondicional durant tota la meva vida.

Després voldria agrair als meus directors per haver-me donat l’oportunitat de fer aquest tesi. A José

Luís Sànchez-Lizaso, per velar per mí durant aquestos set anys així com per haver cregut en mí des de

el principi, donant-me l’oportunitat d’entrar al departament durant la beca de col·laboració i en acabant

continuar amb el projecte BIOMEX. Gràcies també a Carlos Valle, per ensenyar-me pràcticament tot

el que sé de treball en la mar, d’anàlisis de dades, estadística, per ser el meu company de faena durant

tot aquest temps....i sobre tot, per la seva paciència en certs moments. A Just Bayle, per permetre’m

col·laborar amb el projecte EMPAFISH, amb el que vaig poder demanar la beca que ha finançat par

d’aquesta tesi i també per les “preses”, sense les quals supose que aquesta tesi encara no hi seria

escrita.

Gràcies també al grup de gent del projecte BIOMEX, especialment a Sandrine Polti per adentrar-me

en el món de la pesca esportiva i a Aitor Forcada per ensenyar-me a diferenciar els noms dels peixos

durant els embarcs amb els artesanals entre tantíssimes altres coses, a més del més important: pel seu

suport i amistat durant els darrers anys.

Molstisimes gràcies també a la gent de la Reserva marina de Tabarca, entre ells Felio Lozano i els

guardapesques que ens varem facilitar dades del pesca esportiva. Així també, voldría agrair al equip

del parc Natural de la Serra Gelada, especialment a Santa per tores les dades i a Eduardo Mínguez per

el seu suport durant els mostrejos. Als directors del Clubs Nautics d’Altea i Benidorm, per deixar-nos

un lloc sempre que ho necessitàvem. Agrair també a Rebeca, Curro i Manolo del centre de busseig

Celacanto, que ens han facilitat sempre la faena i ens han tractat com “de casa”. Gràcies també a Mari

i Javi, del centre de busseig Poseidon. Moltísimes gràcies a tots i totes els bussejadors i bussejadores,

pescadors i pescadores esportives que, de forma totalment desinteresada, han participat amb les seves

experiències en aquesta tesi, i dels que he tingut la sort d’ensenyar-me tantes coses.

A la gent de la Uni de Palermo, que tan bé me va cuidar durant la meva estança...sobre tot a Maria

Grazia, Manfredi di Lorenzo, Antonio di Franco, Giulio Granzita, Marco Milazzo i Renato Chemello.

Pasant a la gent de la casa, primer agrair-li a una persona que ja es part de la meva famila, a Yoana, el

seu carinyo, amistat i suport durant tots aquestos anys, estant ahí sempre per a tot...gràcies xiqueta!!.

Agrair també a quatre grans amigues que te tingut la sort de trobar-me pel camí, a Ana, Paqui, Carmen

i Mayte, gràcies per la vostra amistat, ànims i consells durant tot aquest temps. A Damià, per estar ahí

sempre, en els bons i roïns moments...i per la teva paciència immesurable amb el meu anglés-

macarrònic. A Mercedes i Marta, per ensenyar-me tot sobre eixa cosa “amorfa” que va resultar, era

una ascidia. A Alfonso Ramos, gràcies per presentarme a Halocynthia. A Yolanda, per fer-me

entendre la diferència entre un plagio i un orto... A tota la gent que son o han pasat pel departament en

algún moment, amb els que he tingut la sort de compartit tantes hores de faena i de bons moments fora

d’ella: Pablo Sanchez, Vero (tú casi ets del depar), Candela, Elena, Loya (per obrirme la porta...),

Pablo Arechavala, Tito, Aurora, Jose de la Osa, Emilio, Lute, Cristina, José Miguel, Celia, Viky,

Kilian, Esther, Marko, Juan Carlos, Antonio-medusas, Andrés, Irene, Pilar Hernández, Jose Jacobo,

José Vicente Guardiola, Macarena, Ana Nuevo, Agustín, ...i perdoneu-me si m’en he deixa algú més.

Gràcies també al meus amics i amigues que encara que no enteníeu que feia en la feina sempre heu

tingut unes paraules de ànim i un ratet per a compartir una birra amb mi. A les meves xiques (Ana,

Marta L, Marta Pomar i Vero), a la colla dels blocs, tot el grup de gent “Ruina” (quina sort haver-us

trobat sou genials), a les “tabarkines”, a Sergio, Angel, Ernesto, Esther, JuanFran....I, encara que sé

que vos he tingut un amica abandonades últimament, volia donar-li les gràcies de formar molt especial

a Manoli, Ana, Betty, Eva i Nuria per el vostre carinyo, per ajudar-me a superar tants moments difícils

que han anant sorgint en aquest temps i sobre tot per fer-me saber que podía comptar amb vosaltres,

espere que açò no canvie mai...

A tots i totes donar-vos les gràcies, perquè sense la vostra aportació professional o no, aquesta tesi no

haguera estat possible.

CONTENTS

GENERAL INTRODUCTION 1 1. Threat’s oceans: the payment for a service 3 2. The Mediterranean Sea: a threatened heritage 4 3. Marine Protected Areas as conservation and management tools 6 4. Anthropic impacts in multiple-use MPAs 8

4.1. Recreational fishing impact 9 4.2. SCUBA diving impact 10 4.3. Boat anchoring impact 11 4.4. Human trampling impact. 12

5. Justification 136. Objectives 14

6.1. Recreational fishing 14 6.2. SCUBA diving 14

CHAPTER 1 17Abstract 191. Introduction 172. Material and Methods 21

2.1. Study area 212.2. Data collection 21 2.3. Data processing 22 2.4. Analysis of data 23

3. Results 243.1. Long-term evolution of fishing effort (1997-2004) 24 3.2. Individual angler profiles 24 3.3. Effort 253.4. Catch 253.5. Catch per unit of effort 25 3.6. Estimation of total effort and total catch 28

4. Discusión 294.1. Anglers’ profile and motivations 29 4.2. Recreational fishing effects on fish assemblages 30 4.3. Socioeconomic conflicts 30 4.4. Recommendations 31


2.1. Study area 362.2. Recreational fisher interviews 37 2.3. Aerial survey 37 2.4. Recreational fishing competition 38 2.5. Estimation of total annual fishing effort and match 39

3. Results 393.1. Recreational fisher profile 39 3.2. Aerial survey 40

3.3. Recreational fishing competition 43 3.4. Estimation of total annual fishing effort and catch 44

4. Discussion 44

CHAPTER 3 49Abstract 511. Introduction 512. Material and methods 52

2.1. Study area 522.2. Sampling diver behaviour 53 2.3. Statistical analyses 54

3. Results 543.1. Divers profile 54 3.2. Effect of diver behaviour 54 3.3. Diver behaviour according to profile 56 3.4. Diver behaviour according to dive features 56

4. Discussion 584.1. Management of recreational scuba diving 60


2.1. Study area 672.2. Studied species 67 2.3. Sampling prcedures 68 2.4. Statistical design and analysis 69

3. Results 713.1. Effect of diving on the density 71 3.2. Changes in the structure of size class population 72 3.3. Changes in the degree of exposure 72

4. Discusión 744.1. Indicator species for detect the SCUBA diving effect on coralligenous assemblages 74 4.2. Response of Halocynthia papillosa to dive activity 75 4.3. Halocynthia papillosa as diving impact indicator 76


2.1. Study area 822.2. Studied species 83 2.3. Experiment set-up 83 2.4. Assessment of raised sediment 84 2.5. Statistical analyses 84

3. Results 863.1. Analysis of the raised sediment 86 3.2. Effect of flapping on Halocynthia papillosa 88 3.3. Effect of raised sediment on Halocynthia papillosa 89

4. Discusión 904.1. .Response of H. papillosa to SCUBA diving impact 90 4.2. Use of H. papillosa as SCUBA diving impact indicator 91

GENERAL DISCUSSION 93 1. Increased of human-uses in MPA: associated conflicts 95 2. Management strategies for recreational uses in MPA 100

2.1. Carrying capacity of recreational use and limit of acceptable change 100 2.2. Zonation as management tool for recreational activities 103 2.3. Marine education and awareness programmes 105 2.3. Long term monitoring programmes 107

CONCLUSIONS 109

GENERAL ABSTRACT 113 1. Introducció General 115 2. Impacte de la pesca recreativa en una Àrea Marina Protegida Mediterrània 117 3. Evaluant l’impacte de la pesca recreativa al Parc Natural de la Serra Gelada (SW Mediterràni): un estudi de base per a la seva futura regulació 118

4. Impacte sobre el bentos del busseig recreatiu en un Àrea Marine Protegida Mediterrània 119 5. Halocynthia papillosa (Linnaeus, 1767) com un indicador del impacte del busseig 121 6. Halocynthia papillosa com indicador de l’impacte del busseig: un experiment in situ 123

BIBLIOGRAPHY 125

Als meus pares, Amparo i Luis

i la meva germana, Jessica

1

___________________________________________________________General introduction

GENERAL INTRODUCTION

1. Threat’s oceans: the payment for a service

Although the ocean cover more than 70% of the Earth’s surface, today we know little about three level of biodiversity in the sea: genetic, species and ecosystem (McAllister, 1996; Ormond et al., 1997). However, we know that marine biodiversity is extremely valuable for the humankind since it accounts for over 60% of the economic value of the biosphere (Constanza et al., 1997). Specifically, sea provides five basic services to humans (Daily, 1997; Moberg and Folke, 1999):

� Ecosystem services: an example is the global “biological pump” that sequesters atmospheric carbon dioxide and transports carbon to the deep sea.

� Food: around 20% of the animal protein consumed by humans is provided from marine fisheries.

� Minerals and chemicals: examples include abiotic resources (fossil fuels, manganese nodules, table salt, etc.), as well as chemicals derived from organisms (such as chitin from crustaceans, used in medical and technological applications).

� Medicines: Marine organisms are increasingly found to contain biomedically active compounds, including antitumoral agents.

� Recreation and ecotourism: marine life, especially certain megafauna

(marine mammals, etc), has traditionally inspired humanity. For instance, recreational use of corals supports many regional economies.

Despite their immense value, it is not sufficiently recognized and marine ecosystems are rapidly deteriorating due to human activities. During more time the ocean plunder has happened out of control because it was thought to be so vast that it was judged inconceivable that human activities might significantly alter the structure and functioning of marine ecosystems. However, it is now obvious that seas reflect the footprint of human uses; from industries such as tourism and transportation, overfishing, waste disposal, excess of nutrients loads from agricultural runoff and the introduction of exotic species (reviews by Norse, 1993; Peterson and Estes, 2001; Steneck and Carlton, 2001). Moreover, it is important to keep in mind that anthropogenic threats to marine biodiversity are ubiquitous, cumulative and synergistic (Hughes, 1994). The capacity of oceans to recover from global perturbations and thus, to maintain ecosystem goods and services is weakening (Worm et al., 2006). Additionally, these problems are likely to increase in the future due to climate change. Consequently, as early as the past century, international organisms involved in ocean conservation and protection declared the urgent necessity of implementing new and drastic management measures, between which the establishment of marine protected areas was suggested (IUCN, 1988; Plan Development Team, 1990)

3

Anthropic impacts in Mediterranean MPAs__________________________________________

2. The Mediterranean Sea: a threatened heritage

The Mediterranean Sea is a biodiversity hotspot due to the unique variability of life hosted in its waters (Shi et al., 2005). Compared to others regions of the world and considering its small dimension (less than 0.8 % of the world’s ocean area, Farrugio et al., 1993), the Mediterranean is regarded as one of the world’s priority areas for conservations due to its elevated number of threatened and endemic species; its high biodiversity, that greatly changes among numerous different ecosystems and also because of the escalating human pressure that has been going on over centuries (Myers et al., 2000; Mittermeier, 2004; Shi et al., 2005; Bianchi, 2007; Briand and Giuliano, 2007; Abdulla et al., 2008). In the Mediterranean around 12.000 species (7% of all known marine species) have been recorded of which 25-30% are endemism (Bianchi, 2007). Although species richness is lower compared to tropical seas, Mediterranean food webs are comparable in their complexity (Sala, 2004).

The following are cited some examples that highlight of the Mediterranean role in marine biological diversity. For instance, the main spawning grounds of the Atlantic bluefin tuna, Thunnusthynnus, are in the Balearic archipelago, southern Tyrrhenian Sea, Levantine Sea and south of Turkey (Medina et al., 2007). Increasingly numbers of loggerhead, Caretta caretta, and green turtles, Chelonia mydas, annually nest

in the Mediterranean (Broderick et al., 2002). Carcharodon carcharias, white shark, species classified as endangered in the Mediterranean by the IUCN Species Survival Commission, has a unique breeding zone the Sicilian Channel (Tudela, 2004) and the area of Corso-Ligurian basin hosts a rich cetacean fauna (around 3500 individuals) of fin whale, Balaenoptera physalus (Notarbartolo di Sciara et al., 2003).

Specifically, seagrasess are the first hotspot habitat of the Mediterranean, since it provides habitat for an elevated number of vegetal and animal species (both vertebrates as invertebrates) which live, feed and reproduce closely associated to their leaves and rhizomes (Harlin, 1975; Den Hartog, 1983; Gambi et al., 2006). Consequently, P.oceanica and other seagrasses are key species which provide oxygen and nutrients as well as a natural coastal protection (Blanc, 1974; McRoy and McMillan, 1977; Boudouresque and Jeudy De Grissac, 1983; Sánchez Lizaso, 1993; Duffy, 2006). In addition, one of the most beautiful, diverse and productive ecosystem of the Mediterranean is the coralligenous assemblages, a bioconstruction which age may range from 600 to 7000 B.P. (Sartoretto at al. 1996). This highly diverse and heterogeneous biocenosis is built by an elevated number of species of algae, sponges, gorgonians, corals, bryozoans and tunicates and host many other taxa such as crustaceans, molluscs or fishes (Ballesteros, 2006). Other bioconstructions like vermetid reefs are within the most important of intertidal Mediterranean shores and it helds a very

4


diverse community (Molinier and Picard, 1953). These biogenic formations are built by the sessile

gastropods Dendropoma petreum and Vermetus triquetrus, which are endemic of the Mediterranean.

However, there are several threats that affect this unique environment., One of the most critical is coastal occupation since, presently , resident population along the Mediterranean coastal region is 143 million and during the summer months it rises up to ca. 320 due to tourist activities (Blue Plan, 2005). Consequently, the associated human impact has altered original landscapes and has produced deleterious effects on the marine environment (IUCN, 2007). The most visible effects of this pressure are the physical loss of characteristic habitats like Mediterranean. Seagrass meadows, a significant habitat used for refuge, reproduction and feeding of the 25% of Mediterranean flora and fauna species (Delbaere, 1998), that has showed an important regression due coastal physical modifications, pollution and bottom trawling. (Pérès and Picard, 1975; Zieman, 1976; Meinesz and Lefevre, 1978; Ardizzone and Pelusi, 1984; Sánchez Lizaso et al., 1990; Martín et al., 1997; Ruiz, 2000; Gonzalez-Correa et al., 2005). In addition, the coralligenous communities are highly impacted by pollution, trawling and SCUBA diving (Milazzo et al., 2002a; Ballesteros, 2006), therefore generating the deterioration of this singular biocenosis. The natural and characteristics environments are being replaced by artificial constructions used by tourism, coastal urbanizations and other economic activities. It has been estimated that if actual coastal occupation continues at the same rate, 50% of the Mediterranean coast may be transformed in a continuous metropolis in the next years producing an

irreversible change of marine costal environment (Blue Plan, 2005).

During the last century, fishing has rapidly increased in the Mediterranean and nowadays is an unsustainable industrial exploitation of natural resources (Goñi et al., 2000). As early as 2006 FAO assured that the majority of Mediterranean commercial fish stocks are over-exploited. In 2007, 58% of wild stocks were in a situation of full exploitation; 19% were overexploited, 8% were depleted and 1% was recovering from depletion which sums 86% of traditional stocks threatened to different degrees. Captures of wild fish have remained stable since the 80’s despite technical improvements, indicating that fish stocks are being exploited at their maximum levels (FAO 2009). Negative effects of fishing are not only limited to targeted species since some fishing activities produce impacts on the habitat (i.e. trawling on seaweeds) (González-Correa et al., 2005) and other non-targeted species like turtles, sharks and cetaceans are threatened by incidental by-catch (Tudela et al., 2005).

Moreover, the semi-enclosed status of the Mediterrenean Sea involves a high vulnerability to pollution because the total water renewal through the Strait of Gibraltar lasts around 70 seventy years (Agardy, 2003). All marine organisms are

5


under the effects of pollution due to industrial and agricultural waste, heavy metals and persistent organic and solid material which are discharged to sea. Discharged hydrocarbons (during shipping operations or accidents) may reach more critical levels since its impact can range from genetic alteration to direct poisoning of marine organisms (GESAMP, 1993).

Other important threat to the Mediterranean Sea is the introduction of non-indigenous species. The main vectors of species introduction are the waterways of Suez Canal (lessepsian species of Red Sea), hull fouling and ballast water associated with shipping and aquaculture activities (Flagella and Abdulla, 2005). Nowadays at least 137 molluscs, 99 fishes, 63 crustaceans and 9 macrophytes species have been recorded as non-native to the Mediterranean (Boudouresque and Verlaque, 2002; CIESM, 2002, 2004). The impacts are mainly ecologic and economic when the introduced species displaces native species, as has happened with the invasion of two species of green algae of the genus Caulerpa which compete with seagrass species in many Mediterranean regions (Galil, 2007). The success of native species depends of the healthy status of habitat to colonize, so much higher the degradation state of habitat will be more easily colonized by invasive species (Galil, 2000).

Furthermore, climatic change has been recognized as one of the greatest threats worldwide and their effects are visible now on all scales of marine ecosystem processes. The expected consequences of climate change on the oceans include water acidification and warming, rise of sea level due to polar thaw and the consequent

changes of atmospheric circulations and oceanic currents. In the Mediterranean, from 1980, an anomalous increase of surface sea temperature (SST) and deep water has been recorded (Benthoux et al., 1998; Lelieveld, 2002; Diaz-Almela, et al., 2007). As a consequence, species compositions may change in space and time because warmer-water species distributions ranges may expand and cold-water species ranges may shrink. Mass mortality events which have already happened in the north-western Mediterranean (Garrabou et al., 2001) would be more frequents. Increment of CO2 levels reduces the ocean pH (process known as acidification) and carbonate ion concentrations (Bates et al., 2008). This may affect marine organism that present carbonate or aragonite dependence, such as corals, plankton, coralline algae, etc (Hall-Spencer et al., 2008).

3. Marine Protected Areas as conservation and management tools

Among other management tools (i.e. sustainable exploitation and development, pollution control, barrier front non-indigenous species), Marine Protected Areas (MPAs) have been proposed as the one of the most effective conservation and management tool to face up this unprecedented alteration of marine ecosystems and mitigate its effects (Lubchenco et al, 2003). Positive effects of MPAs have proven useful even beyond their boundaries (for a review of MPA effects see Halpern

6


and Warner, 2002; Halpern, 2003; Gell and Roberts, 2003; PSICO, 2007; Claudet et al., 2008). MPA supplies refuge for threatened species, prevents degradation habitat and allows the development of natural biological communities. MPA may allow the spillover of adults and juveniles that can re-colonise adjacent areas, help to recover depleted fish stocks or restore degraded environments.

Several definitions of Marine Protected Area have been formulated and applied in different conservation and management contexts. The international definitions that have been used are:

� “Any area of intertidal or subtidal terrain, together with its overlying water and associated flora, fauna, historical and cultural features, which has been reserved by law or other effective means to protect part or all of the enclosed environment”. (Resolution 17.38 of the IUCN General Assembly, 1988, reaffirmed in Resolution 19.46, 1994).

� “Any defined area within or adjacent to the marine environment, together with its overlying waters and associated flora, fauna, and historical and cultural features, which has been reserved by legislation or other effective means, including custom, with the effect that its marine and/ or coastal biodiversity enjoys a higher level of protection than its surroundings” (Convention on Biological Diversity, 2003).

� A new IUCN definition was presented at the World Conservation Congress, Barcelona October 2008 “A clearly defined geographical space, recognised, dedicated and managed, through legal or

other effective means, to achieve the longterm conservation of nature with associated ecosystem services and cultural values”.

There are just over 5,000 marine protected areas covering more than 3.1 million km2 (only 0.7% earth’s surface). The world smallest MPA is Echo Bay Provincial Park in Canada with a documented area of 0.4 ha and the largest MPA is the Phoenix Islands Protected Area in Kiribati with a documented area of 41,050,000 ha (UNEP-WCMC, 2009). The oldest designated MPA is the San Juan County/Cypress Island Marine Biological Reserve in the United States which was designated in 1923. In the Mediterranean context, the protected and managed marine area covers 97,410 km², approximately 4% of the Mediterranean. Excluding the Pelagos Sanctuary (87,500 km²), the area covered by coastal MPAs amount to only 9,910 km², which is 0.4% of the total surface of the Mediterranean Sea. According to the criteria defined per UICN, 94 MPAs exist at the present in the Mediterranean Sea. The establishment of the Mljet National Park (Croatia) in 1960 and Port-Cros National park (France) in 1964 created the two first MPAs. The rate of creation of new MPAs increased rapidly in the beginning of 1990s.

In general, the type of protection applied within MPAs in the Mediterranean is very diverse and reflects the cultural and political differences existing among countries

7


(Wood et al., 2008). The majority have been classified as multiple-use marine areas (Harmelin, 2000; Badalamenti et al., 2000; Francour et al., 2001). Multiple-use marine protected areas seek a balance between biodiversity protection and continued human use. Historically, also in the Mediterranean, designation was primarily driven by the presence of charismatic species, unique features or opportunity more than by a holistic ecological approach (Francour et al., 2001; Fraschetti et al., 2002, 2005).

4. Anthropic impacts in multiple-use MPAs

Tourism is now the largest single economic sector in the world and its impact on coastal environments is considerable and extremely difficult to manage and limit (Davenport and Davenport, 2006). In recent years there has been a growth of tourism in coastal areas (Boudouresque and Ribera, 1993; Davis and Tisdell, 1995; Ribera, 1991), with around 70% of European holiday-goers choosing the Mediterranean coast as their vacations destiny (European Commission, 2008).

Mass tourism implies the development of a range of leisure activities, such as: recreational fishing (Francour et al., 2001; Lynch et al., 2004; Pawson et al., 2007), diving (Asafu-Adjaye and Tapsuwan, 2008; Barker and Roberts, 2004; Hawkins et al., 2005) or boat anchoring (Creed and Amado Filho, 1999; Francour et al. 1999; Francour et al., 2001; Milazzo et al., 2004; Montefalcone et al., 2006). All these activities can damage the natural ecosystems, leading to changes in communities and biodiversity (Sala et al.,

1996; Zabala, 1996; Brown and Taylor, 1999; Martínez et al., 1999), in clear conflict with the conservation objectives of the MPAs. Furthermore, the supposed aesthetic value of these marine protected areas (MPAs) has led to mass tourism in those areas (Ribera, 1991, 1992; Richez, 1991, 1992, 1993; Capellà et al., 1998; Badalamenti et al., 2000; Milazzo et al., 2002a) which may results in real “leisure underwater parks”, such as Medes Islands Reserve (Ribera-Siguan, 1992). Increased human presence and ancillary infrastructures can exceed the environmental carrying capacity, producing conflicts difficult to solve and non considered effects. In this regard, declared MPAs without a proper planning and management taking into account all stakeholders, are likely to be victims of their own success because of “overfrequentation” (Ramos-Esplà et al., 2004).

On the other hand, the incomes of those leisure activities in the financial support of MPA had been showed (Arin and Kramer, 2002; Green and Donnelly, 2003; Depondt and Green, 2006; Rocin et al., 2008). A lack of financial support for the cost of managing MPA is frequently cited as the central behind the failure of many MPAs globally (Depondt and Green, 2006). The growing “marine tourism” market has been identified as one way in which MPAs may be financed (Fabinyi, 2008). This phenomenon, besides the negative ecological impacts in sensu, may generate a cycle in local

8


economic benefits provide by the MPA; protection generates more visitors, who in turn damage the quality of the marine ecosystem, which after a while results in the decreased appeal of the area to visitors, and hence a diminishing number of visits (Dixon et al., 1993; Davis and Tisdell, 1996).

The study of human impacts on MPAs has interested to a high number of authors and the results highlight how substantial these effects can be, potentially threatening the diversity of marine communities (Brown and Taylor, 1999). Most of these effects have been studied both tropical and in non-Mediterranean temperate seas. Probably, the relatively young age of Mediterranean MPAs is the major reason for the lack of these types of studies (Millazzo et al., 2002a).

In this thesis, a total of 146 published studies were revised in order to address the current knowledge about these impacts and to localize the gaps concerning Mediterranean MPAs. These studies were carried out world-wide and focused in the effects of anthropic activities on marine communities as recreational fishing, SCUBA diving, anchoring and trampling.

4.1. Recreational fishing impact

Recreational fishing is one of the most popular leisure activities in the Mediterranean, it involves a large number of users and consequently high levels of effort (Cowx, 2002; Pitcher and Hollingworth, 2002; Westera et al., 2003; Cooke and Cowx, 2006), which in some cases may equal or exceed the effort and catch of commercial fishery (Pollock, 1980; Dunn et al., 1989; Steffe et al., 1996;

Beal et al., 1998; Murray-Jones and Steffe, 2000). Commercial and recreational fishing have similar demographic and ecological effects on some fish species populations and can provoke equally serious ecological and economic consequences (Post et al., 2002; Coleman et al., 2004). For this reason, both recreational and commercial fishing should be taken into account to ensure effective fishery management, especially in MPAs. However, the impact of recreational fishing on fish populations has not been scientifically discussed to the same extent as commercial fishing (Cooke and Cowx, 2004, 2006). Biological consequences on the target populations can include truncation of the natural age and size structure, drops in fish abundance and biomass and changes in fish composition (Coll et al., 1999; McPhee et al., 2002; Westera et al., 2003; Coleman et al., 2004; Cooke and Cowx, 2004, 2006; Lewin et al., 2006).

Recreational fishing is as an important leisure activity which involves a high active industry mainly based on trade and tourism. Pointing at this, of all recreational fishing studies revised (72 studies), a great proportion (41.6%) is related with economics (Glover, 1980; Morey et al.,1991; Fisher and Ditton, 1993; Storey and Allen, 1993; Keith et al., 1996; Mike and Cowx, 1996; Schoeman, 1996; Baker and Pierce, 1997; Loomis et al., 1999; Smith, 1999; Roth et al., 2001; Shrestha et al., 2002; Toivonen et al., 2004) and

9


sociological aspects of recreational fishing (Hudgins, 1984; Chipman and Helfrich, 1988; Vigliano and Lippolt, 1991; Bohn and Roth, 1997; Vigliano et al., 2000; Ferrer Montaño et al., 2005).

Other studies (15% of recreational fishing works revised) are focused on the parameters that influence fishing mortality due to so-called "catch-and-release" (Feathers and Enable, 1983; Muoneke and Childress, 1994; Kwak and Henry, 1995; Diggles and Ernst, 1997; Broadhurst et al., 1999; Taylor et al., 2001; Davis et al., 2001; Ayvaizian et al., 2002; Cooke et al., 2003; Thorstad et al., 2004; Broadhurst et al., 2005). This activity is widespread in continental environments in Australia, North America and European countries but is little employed in Mediterreanean marine recreational fisheries.

Only eight of these recreational fishing studies were carried out in MPAs (Bennet and Attwood, 1991, 1993; Harmelin et al., 1995; Coll et al., 1999; Ferrer-Montaño et al., 2005; Lockwood et. al., 2001; Westera et al, 2003; Lynch, 2006) and only the studies of Harmelin et al. (1995) and Colleta al (1999), were done in Mediterranean MPAs (2.7%). Harmelin et al. (1995), showed an evident impact from amateur fishery in Port Cross Park, utilizing as indicator two fish species, Coris julis and Serranus cabrilla. In the Cabrera Marine Park, Coll et al., (1999) recorded, by means of visual census, a recovery of Epinephelus marginatus population as consequence of the prohibition of recreational fishing. The other recreational fishing studies in the Mediterranean were not performed in the MPAs. Coll et al., (2004) also studied the most affected species by spear-fishing

employing long time series of competition dates. Morales-Nin et al., (2005) studied all recreational fishing modalities by means of telephone and in situ surveys, fishing logbook and spearfishing competition dates around the Majorca Island for a long time period. On both studies a deleterious effect on the fishery resources was shown.

In the Mediterranean, recreational fishing also presents economic, social and cultural roles, and it can be considered by some as a competitive sector against commercial fishing, which is mainly dominated by small-scale fleet operating in coastal areas (‘artisanal fisheries’). This work has shown evidence that additional studies are needed, especially in Mediterranean MPAs, to reach a better understanding of the impact of recreational fishing. Sound management of fishery resources is essential in order to determine which are the most affected species, how much biomass is extracted from the ecosystem, as well as identify the profile of recreational fishers, their habits and motivations in order to clearly recognize the impacts and measure the ecological consequences.

4.2. SCUBA diving impact

SCUBA diving may result in the deterioration of benthic communities, as divers can easily damage marine organisms through physical contact with their hands, body, equipment

10


and fins (Talge, 1990, 1992; Rouphael and Inglis, 1995, 1997; Tratalos and Austin, 2001; Zackai and Chadwick-Furman, 2002; Pulfrich et al., 2003; Uyarra and Côté, 2007). Although the damage produced by individuals is usually quite minor, there is some evidence that the cumulative effects of these disturbances can cause significant localised destruction of sensitive organisms (Garrabou et al., 1998; Hawkins et al., 1999; Plathong et al., 2000). There is a bigger problem when the diving activity focuses in marine protected areas. In some cases, the effects of a massive number of divers in a few places in a marine protected area can be contrary to the main objectives of the creation of the MPA (Davis and Tisdell, 1995, 1996; Coma et al., 2004; Hawkins et al., 2005).

Relationship among diver behaviour and impacts on marine communities, affected species and the extension of injuries has been widely studied around the world, mainly in coral reef areas such as Caribbean Sea, Red Sea or Australian Great Barrier Reef. Of a total 33 revised diving impact studies, 24 were done in MPAs but only 6 (18%) were carried out within Mediterranean MPAs. Effects of SCUBA diving on coral reefs has been extensively studied since 42% of the revised works evaluated the damage produced on hard and/or soft corals (Hawkins and Roberts, 1992; Medio et al., 1996; Harriot et al., 1997; Rouphael and Inglis, 1997; Hawkins et al., 1999; Scheleyer and Tomalin, 2000; Rouphael and Inglis, 2001; Walters and Samways, 2001; Rouphael and Inglis, 2002; Zackai and Chadwick-Furman, 2002;Wielgus et al., 2004; Barker ans Roberts, 2004; Hawkins et al., 2005), compared to the only two studies that evaluated the effect

on fish (Shackley, 1998; Milazzo et al., 2005). Eight of the revised works investigated diver’s underwater behaviour by means of directly following and diving profile and experience (Medio et al., 1996; Harriot et al., 1997; Rouphael and Inglis, 1997; Schaleyer and Tomalin, 2000; Walters and Samway, 2001; Zackai and Chadwick-Furman, 2002; Barker and Roberts, 2004; Dearden et al., 2006).

Regarding the analyzed species, of the six Mediterranean studies which analysed diving impacts on MPAs, two evaluated the diving effect on bryozoans (Sala et al., 1996; Calvisi et al., 2003), two on gorgonians (Garrabou et al., 1998; Linares et al., 2005), one evaluated the effect of feeding on the fish assemblage (Milazzo et al., 2005) and the last proposed a vulnerability index to SCUBA diving impact for some Mediterranean species (Lloret et al., 2006). However, none of these studies took into consideration the diver behaviour, profile, the type or rate of injury, or the effect on other species besides of these cited groups. The lack of this type of study highlights the need for efforts to increase present knowledge of SCUBA diving effect on Mediterraenan, especially in the case of MPAs.

4.3. Boat anchoring impact

In most cases “overfrequentation” by visitors in MPAs implies an increase of the number of boats and, in consequence, a rise of potential

11


anchoring events. Boat anchoring has been proven to be a severe threat to seagrass meadow conservation and tropical reef health (Davis, 1977; Walker et al., 1989; Creed and Amado-Filho, 1999; Dinsdale and Harriot, 2004; Saphier and Hoffman, 2005). The extent of anchor damage that has been observed depends on the number of boats and their sizes, weather conditions and substrate firmness (Walker et al., 1989; Francour et al., 1999).

In the Mediterranean Sea, the communities which may be affected by damage anchor are coralligenous habitats, seagrass meadows and, to a lesser extent, the infralitoral algal assemblages (Milazzo et al., 2002a). Of a total of 20 anchoring effect studies revised, eight (40%) were carried out in Meditteranean regions. Many research efforts assessed the consequences of anchoring on P. oceanica meadows, showing a direct adverse effect on meadow cover and/or shoot density (García-Charton et., 1993; Francour, 1994; Poulain, 1996; Francour et al., 1999; Ganteanume et al., 2004; Milazzo et al., 2004; Leriche et al., 2006; Montefalcone et al., 2006). Two of these works applied an experimental anchoring to corroborate the cause-effect relation between the anchoring and the seagrass damaged (Francour et al., 1999; Milazzo et al., 2004). Only García-Charton et al. (1993) evaluated the anchoring effect on the meadow-associated fish assemblage and Gauteaume et al. (2004) the possible anchoring impact on Pinna nobilis. At the present, the effects on other Mediterranean assemblages, such as coralligenous, remains unknown.

4.4. Human trampling impact.

Many marine organism (i.e. sessile invertebrates, seagrasses and erect macroalgae) have revealed a high vulnerability to human trampling (people walking over intertidal and upper infralittoral areas) in a wide array of environmental conditions and habitats (Liddle, 1991; Povey and Keough, 1991; Brosnan and Crumrine, 1994; Keough and Quinn, 1998; Milazzo et al., 2002a, 2004; Casu et al., 2006).

Nowadays, few data about the trampling effect exist for the Mediterranean Sea. Publications’ revision revealed that only three out of 15 studies (Milazzo et al., 2002b, 2004; Casu et al., 2006) where carried out in this environment. In most studies, the trampling effect was studied by means of experimental trampling and in some cases combining this with a comparison between natural sites with different number of visitors (Povey and Quinn, 1991; Brosnan and Crumrine, 1994; Keough and Quinn, 1998; Brown and Taylor, 1999; Schiel and Taylor, 1999; Eckrich and Holmquist, 2000; Milazzo et al, 2002b; Milazzo et al, 2004; Casu et al., 2006).

Vulnerability to this type of impact seems to mainly depend on the nature and morphology of marine algae and level use (Povey and Keough, 1991; Brosnan and Cumine, 1994). Erect foliose algae are severely damaged by trampling (Keough and Quinn, 1998; Schiel and Taylor, 1999; Milazzo et al., 2002a), while algal turfs or low

12


caespitose forms are more resistant (Brosnan and Crumrine, 1994; Schiel and Taylor, 1999). In other hand, the effects of human trampling on invertebrate species seems to be very species-dependent and, in the case of mobile associated fauna (i.e. gasteropods, cryptic fishes, crustaceans), lasting in the short term (Povey and Keough, 1991; Brosnan and Crumrine, 1994; Brown and Taylor, 1999; Milazzo et al., 2004; Casu et al., 2006).

As has been evidenced on previous studies, human trampling may cause the physical removal of a species (or a group of species) thus affecting population dynamics and indirectly the diversity of the entire community (Brosnan and Crumrine, 1994). These impacts may be inconsistent with conservation goals of MPAs, therefore more efforts to study the affected species, the extension of the impacts as well as possible management measures should be carried out.

5. Justification

Although the basic purpose of MPAs is not economic development but ecosystem conservation, the incomes of ‘marine tourism’ associated to it may be an opportunity for finance the management cost (Fabinyi, 2008). However, it should be taken into account the negative effect that an increased number of visitors may produce on the marine environment when their activities inside the MPA are not properly managed (Davis et el., 1995). The balance between environmental cost and economic benefits is frequently difficult to combine but necessary for a successful MPA in terms of sustainability and conservation.

Based on the lack of knowledge revealed by the revised literature about several human impacts on Mediterranean MPAs and their socio-economic importance, recreational fishing (an extractive leisure activity) and SCUBA diving (a non-extractive leisure activity) were the topics chosen as the most important to analyze and selected for research throughout this study. In addition to cited ecological and biological impacts, these two effects may create complex conflicts between users difficult to solve; on one hand, the conservationist point of view of divers and on the other, the supporters of a ‘traditional and harmless amateur activity’(Fabinyi, 2008). In the particular case of recreational fishing, these conflicts are related with artisanal fishery which in most MPAs are not allowed and is considered an important stakeholder for the effective running of MPAs. Moreover, many Mediterranean MPAs are also Fishery Reserves where the recovery of fish stocks and exportation of biomass to the adjacent fisheries are the main objectives for its establishment (Planes, 2005). Recreational fishing directly acts on the resource to be protected, consequently, there is a clear need to identify the magnitude of the potential impacts in order to properly manage fisheries resources.

SCUBA diving is a leisure activity but also an important economic resource that can benefit local economy and finance the MPA management cost (Arin and Kramer,

13


2002; Green and Donnelly, 2003; Depondt and Green, 2006). The lack of financial support for the cost of managing is frequently a cause of failure of many MPAs (Depondt and Green, 2006). In this sense, SCUBA diving has been identified as the main income source in some MPAs, where sometimes benefits exceed the management cost (Rocin et al., 2008). In MPAs where multiple uses are involved, one of the main objectives is the adequate organization of the different activities in order to allow them in a sustainable manner. Given the increasing importance of travel industry focused on diving, it is needed to clearly recognize the effects in Mediterranean MPAs, to avoid the reported impacts in other regions without renouncing to the economical development of local communities under sustainable management.

This work is structured in seven parts, the first corresponding to this general introduction, followed by five chapters (two of them dealing with recreational fishing and the last three examining SCUBA diving impact). Finally, a general discussion summarizes all the gathered information.

6. Objectives

6.1. Recreational fishing

The study of recreational fishing was approached from two dimensions; first, the study of recreational fishing from land with line (angling) and, in the other hand, the impact of recreational fishing boats. Each type of impact has been addressed in different chapters and for each one of them the main objectives are detailed below:

� Chapter I. To quantify the pressure of angling in the Tabarca Marine Protected Area as well as characterising the profile of anglers. In addition, the direct effect on coastal fish communities is assessed. This is compared to artisanal fishing pressure. Therefore, the aims of this chapter is to contribute to the knowledge of the level of exploitation of fish resources by recreational fishing and to provide essential management data needed for integrated management strategies in coastal zones, particularly in MPAs.

� Chapter II. The Serra Gelada Marine Park is a recent established MPA, in where uses (including fishing activities) have not been regulated yet. It offers a good opportunity for study the recreational boat fishing impact and to recommend sustainable management strategies. Therefore, the objectives of this study are to provide an estimation of exploitation of fish resources by recreational boat fishing as well as useful data about this type of impact which help to sustainable manage this leisure activity.

6.2. SCUBA diving

To study the diving impact activity, the first step was to known the in situ diver’s underwater behaviour. The second step was to detect the existence of species that can be used as indicators of this impact and finally, to corroborate experimentally the usefulness of this indicator of

14


SCUBA diving impact. These goals were tackled into the following chapters with these specifics objectives:

� Chapter III. This study provides an estimate of the damage rate of divers on the benthic communities and shows the existing relationship between diver behaviour and profile, and the immersion characteristics. The most important damaging factors have been identified in order to propose educational tools aimed to users to help to reduce these negative effects.

� Chapter IV. The aim of this chapter is to assess the suitability of the tunicate Halocynthia papillosa as an indicator of SCUBA diving impact and its potential role as an effective tool in the detection of disturbance of the ecological system at broader scales.

� Chapter V. The objective of this section is to detect the possible cause-effect relationship between SCUBA diving impact and H. papillosa response by means of manipulative field experiment and thus deepen in the potential use of this species as SCUBA diving impact indicator in Mediterranean MPA

15

Publication

Luna-Pérez, B-, Valle, C., Sánchez-Lizaso, J.L. 2010. Recreational fishing impact on a

Mediterranean Marine Protected Area. Aquatic Conservation and Freshwater Ecosystems

(Submited).

________________________________________________________________Chapter 1

Recreational fishing impact on a Mediterranean Marine Protected Area

Abstract

A survey of coastal angling was conducted in the Tabarca MPA. We used the reports collected by the fishery warden from 1997 to 2004 and intensive access point survey from 1 September 2003 to 31 May 2005. An increasing trend of annual effort has been observed in recent years. A higher effort during weekends and summer seasons were recorded too. A total of 1,376 fish belonging to 28 species were caught. A high proportion of catches was carnivorous predator species) and belonged to moderate-very high vulnerability group species. The highest value of daily CPUE and BPUE was detected in summer and the lowest value in winter. We estimate that angling activity in the Tabarca MPA would be capable of extracting annually 4.9 tons of fish. The catch of species of great economic importance, such as D. dentex, M. helena, S. umbra, Diplodusspp and S. aurata, was still of considerable proportions in comparison with the total commercial catch. Considering the biological and socioeconomic implications, the implementation of a comprehensive management strategy which taken into account the recreational fishing is needed.

Keywords: angling impact, intrinsic vulnerability index, marine protected area, management, Mediterranean Sea, multiple-uses.

1. Introduction

Tourism is now the largest single economic sector in the world, and its impact on coastal environments is considerable and extremely difficult to manage and limit (Davenport and Davenport, 2006). In recent years there has been a growth of tourism in coastal areas (Ribera 1991; Boudouresque and Ribera 1993; Davis and Tisdell, 1995), with around 70% of European holiday-goers choosing the Mediterranean coast for their vacations (European Commission, 2008). In 2008, of the 180 million tourists that visited Europe, about 40 million chose the Spanish Mediterranean coast as their destination, with leisure activities cited as being the main motivation (Spanish Ministry of Industry, Tourism and Trade, 2008).

Mass tourism implies the development of a range of leisure activities, such as: recreational fishing (Francour et al., 2001; Lynch et al., 2004; Pawson et al., 2007), diving (Barker and Roberts, 2004; Hawkins et al., 2005; Asafu-Adjaye and Tapsuwan, 2008) or boat anchoring (Francour et al. 1999; Creed and Amado Filho, 1999; Francour et al., 2001; Milazzo et al., 2004; Montefalcone et al., 2006). Furthermore, the supposed aesthetic value of marine protected areas (MPAs) has led to mass tourism in those areas (Ribera, 1991, 1992; Richez 1991, 1992, 1993; Capellà et al., 1998; Badalamenti et al., 2000; Milazzo et al., 2002). All these activities could damage the natural ecosystems, leading to changes in communities, and biodiversity (Sala et al., 1996; Zabala,

19

Antrhopic impacts in Mediterraneans MPAs_________________________________________________

1996; Martínez et al., 1999; Brown and Taylor, 1999), in clear conflict with the conservation objectives of the MPAs.

Recreational fishing is one of the most popular leisure activities, and it involves large numbers of people and consequently high levels of effort (Cowx, 2002; Pitcher and Hollingworth, 2002; Westera et al., 2003; Cooke and Cowx, 2006), which in some cases may equal or exceed the effort and catch of commercial fishery (Pollock, 1980; Dunn et al., 1989; Steffe et al., 1996; Beal et al., 1998; Murray-Jones and Steffe, 2000). Commercial and recreational fishing have similar demographic and ecological effects on fished populations, and can have equally serious ecological and economic consequences (Post et al., 2002; Coleman et al., 2004). For this reason, both recreational and commercial fishing need should be taken into account to ensure effective fishery management, especially in MPAs. However, the impact of angling on fish populations has not been scientifically discussed to the same extent as commercial fishing (Cooke and Cowx, 2004, 2006).

This leisure activity can have several biological consequences on the target populations, such as truncation of the natural age and size structure, a drop in fish abundance and biomass, and changes in fish composition (Coll et al., 1999; McPhee et al., 2002; Westera etal., 2003; Coleman et al., 2004; Cooke and Cowx, 2004, 2006; Lewin et al.,

2006, Rius, 2007). In the Mediterranean, recreational fishing also has economic, social and cultural roles, because commercial fishing is largely the domain of small-scale concerns operating in coastal areas (‘artisanal fisheries’).

In multiple-use marine protected areas, all permitted activities must be managed in order to carry out the conservation and protection objectives of the ecosystem, to minimize conflict and allow all users to continue their activities in an ecologically suitable manner. Additional studies are needed to reach a better understanding of the impact of recreational fishing, to determine which are the most affected species and how much biomass is extracted from the ecosystem, as well as to identify the profile of recreational anglers, and their habits and motivations.

The number of licences issued could provide a measure of the amount of fishing effort, but many anglers not in possession of official licenses or fishing illegally have been detected and consequently reliable data regarding effort and yields remain unknown. Very few studies exist on this activity in the Mediterranean (Coll et al., 2004; Morales-Nin et al., 2005; Lloret et al., 2008). This study is not only a long-term analysis of fishing activity, but it is also the first empirical investigation that quantifies in situ the pressure of angling, the direct effect on coastal fish

20

________________________________________________________________Chapter 1

Figure 1. Map of the study area showing the boundaries of the marine protected area of Tabarca and their different levels of protection. I: integral reserve; II: buffer area; III: transitional area.

community, comparing this with artisanal fishing as well as characterising the profile of fishing anglers. The aim of this study is to contribute to the knowledge of exploitation of fished resources by recreational fishing and to provide essential management data needed for integrated management strategies in coastal zones, particularly in MPAs.

2. Material and Methods

2.1. Study area

This study was carried out in the Tabarca MPA, in the western Mediterranean Sea, in southeast Spain (Fig 1.). This MPA was created in 1986 to protect breeding stocks, improve recruitment to neighbouring areas and restock marine species of commercial

interest. The Tabarca marine reserve lies 4 km off the coast, measures 1,400 ha and consists mainly of an island and other islets and small reefs. It is divided into three management zones featuring different levels of protection (Ramos Esplá, 1985; Fig. 1): 1) the integral reserve area (100 ha), where only scientific activities are allowed; 2) the buffer area (630 ha), where some selective fishing methods are allowed; and 3) the transitional area (670 ha), where a number of leisure activities are permitted (angling, SCUBA diving, anchoring). Recreational fishing has always been allowed in this transitional area, but in January 2006 angling was forbidden throughout the entire MPA.

2.2. Data collection

21


Two approaches were adopted for the data-collection process: one to obtain data over a long timescale (effort data) and one to obtain data over a shorter timescale (catch and effort data). For the long-term effort data, we used the reports collected by the fishery warden from 1997 to 2004, obtaining long-term trends in recreational fishing inside the MPA. Intensive surveys were also conducted, from 1 September 2003 to 31 May 2005, (21 months in total). In each season, 30 days were randomly chosen for sampling and were made up of 15 weekend days and 15 week days, making for 120 sampled days in total. The surveys were conducted during the 16 most active hours of the day for fishing, from 8 a.m. to midnight. All the anglers encountered were interviewed to obtain specific details of their fishing effort, catch and socioeconomic data.

The direct interview is one of the most frequently employed techniques in recreational fisheries (Pollock et al. 1997; Lockwood 1999; Reid and Montgomery, 2005), because each catch can be assessed objectively by an observer, unlike when gathering data using telephone, postal or logbook surveys (Malvestuto, 1983; Hayne, 1991; Robson, 1991; Pollock et al., 1994; Sullivan, 1999). The anglers were interviewed at the end of their day fishing (“access point survey”), as this method has the advantage of providing information on both the catch and total effort, although it is restricted to fisheries with only one access point, or imply the need to sample more people at the same time. (Imber et al., 1991; Lockwood 1997; Reid and Montgomery, 2005). In our case, this

method was chosen because the recreational fishery access point for the Tabarca MPA is clearly defined in restricted to the harbour.

During the on-site interview, the following data were recorded for each angler: time spent fishing, number of lines and hooks used, species caught, catch by number and size (total length) of each individual. Information on the following topics was also asked for: gender, age, place of residence, fishing frequency for each season of the year, hours spent fishing per day and preference for weekend or weekday. They were also asked to give their annual expenditure on angling activities, and whether or not they had a fishing licence.

2.3. Data processing

The daily fishing effort data was derived by calculating the average of the total amount of individual effort on any given day. Thus, for example, the effort in day-k is the total amount of the number of lines of angler-j on day-k multiplied by the number of hours spent fishing (Palermo and Thompson, 2000).

The biomass was obtained by using the proven weight-length relationships for this area (Valle et al. 2002; Froese and Pauly, 2004). Daily catch per unit of effort was calculated in terms of biomass (BPUE) and abundance (CPUE), expressed as weight or abundance of the total catch for each species during one day, divided by the total effort (lines/hours) on that particular day.

22

________________________________________________________________Chapter 1

Intrinsic vulnerability of each species was obtained from Cheung et al. (2007), which is defined as the risk of local extinction associated with the life history and ecological characteristics of a species (Cheung et al., 2005). When species information was not available, intrinsic genus vulnerability was used. The most vulnerable fish are deemed to be long-lived and slow-growing species with low reproductive potential and a narrow geographic range. The index values vary from 1 to 100, with 100 as the most vulnerable, and four categories of vulnerability: very high (up to 80), high (up to 60), moderate (up to 40) and low (up to 20) (Cheoung et al., 2005). The proportion of the catch for each category was then calculated. The average intrinsic vulnerability index of the catch was also calculated from the arithmetic mean of the intrinsic vulnerability index of the fishes, weighted by their abundance in the catch.

The trophic category (Bell and Harmelin, 1983) of the species caught was also studied, as well as the catch proportions for each level. The trophic category has four levels and places each species in one category according to their position within the food web based on the prey items preferred.

2.4. Analysis of data

Three univariant analyses were conducted. One two-way analysis of variance (ANOVA) was used to test differences in the effort of fishing between seasons over a number of years (1997-2004) and study the existence of

trends in effort over the years (Underwood, 1981). For this purpose, the experimental design had two factors: Year, a fixed factor with eight levels; and Season, a fixed factor with four levels. A total of 72 reports were entered into the analyses as replicates. A second two-way variance analysis (ANOVA) was used to test the existence of differences in effort distribution for a particular season or type of day. The first factor considered in this analysis was Season, a fixed factor with four levels. The second factor was Type of day, previously crossed with two levels (weekend/weekday). For each combination of factors, 15 replicates were recorded, with a total of 120 replicates. Another one-way analysis of variance was used to identify differences in the catch for the season (four levels). The variables analysed were the mean daily CPUE (catch abundance per unit of effort) and BPUE (catch biomass per unit of effort). Prior to ANOVA, homogeneity variance was verified using Cochran’s test (Cochran, 1951), and transformations carried out when needed (Underwood, 1997).

PERMANOVA was conducted to determine differences among seasons in terms of the specific composition of the catches (Anderson, 2001). Species composition in each catch (in biomass and abundance) was analysed using the PRIMER software package, to identify which species contributed to the differences (SIMPER) (Clarke and Warwick, 2001).

23


Table.1 Two-way analyse of the variance (ANOVA) (Y:Year, S:Season) for the effort (lines) df : degree of freedom; M.S.: Mean square; F: F real; level of signification: ns, non significant; *=P<0.05; **=P<0.01; ***=P<0.001.

Source of variation df M. S. F F versus

Year 7 486.31 6.59** Residual

Season 3 6027.26 81.63** Residual

Y X S 21 188.92 2.56** Residual

Residual 2272 73.83

3. Results

3.1. Long-term evolution of fishing effort (1997-2004).

In order to obtain the angling effort data, 2,920 reports were analysed which were grouped by season. The total effort measured during this period was 18,321 lines, with a highest value of 2,956 lines during 2004, and a lowest value of 1,807 lines in 1998. An increasing trend of annual effort has been observed in recent years. With the exception of the warm spring of 2003, effort was always greatest during the summer periods, as detected statistically by the ANOVA (Table 1).

3.2. Individual angler profiles

Recreational fishing in the Tabarca marine reserve is an overwhelmingly male activity, with 263 males (97.8%) and 6 women (who were usually accompanied a male relative). Ages varied between 22 and 74 years old, with a mean of 46.8 ± 0.63 (mean ± standard error), but anglers between 40 and 60 years old made up the largest

segment (66.2% of the total). These middle-aged men usually go fishing alone (29.2%) or in pairs (49.3%), with few groups of three or more encountered (3.5%). Although anglers who fish all year around were the greatest proportion (54.3%), the percentage of anglers who prefer fishing during the warmest season was also high (38.6%). Similarly, many anglers choose the weekend as their favourite time to go fishing (62.8%), though many also go fishing on weekdays (37%). Only 8 anglers were residents of Tabarca island, with the rest from elsewhere, most commonly Alicante city (40%). The mean expenditure per year per person amounted to €902, on goods and services directly related to their fishing activities in the Tabarca Marine Reserve.

With regard to compliance with fishing regulations, 39.1% of those interviewed did not have a fishing licence, and there was a large number of anglers who refused to respond this question (20%). We also observed that 14.5% of anglers were fishing illegally, as they had exceeded the number of lines allowed.

24

Figure 2. Mean diary effort (lines·hour) at weekend/weekday for each season at Tabarca MPA. Bars indicate the standard error.

3.3. Effort

The 120 days surveyed amounted to a total of 589 lines and 2,622 hours of fishing, with a total effort of 5,781 line hours. The mean daily effort varied widely both between weekends and weekdays and between seasons. The greatest daily effort was recorded during summer weekends (167.24 ± 209.57 line hours); by contrast, the lowest effort was recorded during autumn weekdays (1.33 ± 5.15 line hours) (Fig 2). Weekend effort was always significantly higher than weekday effort, in all seasons (F= 26.82; p<0.001). Differences between seasons were detected (F=3.46; p<0.05), since summer showed an effort significantly greater than the other seasons (SNK; Sumer>other seasons (p<0.01)).

3.4. Catch

A total of 1,376 fish were caught during the survey period, with a total biomass

of 1,428 kg. Catches included 28 fish species from 12 families, the most common being the Sparidae with 13 species recorded (Table 2).

In terms of biomass, most of the catch was made up of carnivorous predator species: 41.6% mesophagous (Diplodus spp, Symphodus tinca, Sparus aurata) and 50.6% macrophagous (Dicentrachus labrax, Dentex dentex, Labrus merula) (Fig. 3).

Average intrinsic vulnerability was 49.2 (out of 100), representing a moderate-high level. A high proportion (70.89%) of fish caught recorded intrinsic vulnerability values ranging from 40 (moderate vulnerability) to 90 (very high vulnerability) and only 28.11% found to be low-vulnerability species.

Figure 3. Percentages of each trophic level. CMM: microphagous carnivorous; CMS: mesophagous carnivorous; CMC: macrophagous carnivorous; HBV: herbivorous.

3.5. Catch per unit of effort

In terms of daily CPUE and BPUE, the highest value was detected in summer (a CPUE of 5.78 ± 0.89 individuals/line hours and BPUE of 420.9 ± 80.84 g/line hours) and the lowest value in winter (a CPUE of 1.29 ± 0.32 individuals/line

________________________________________________________________Chapter 1

25

Table 2. Abundance (Ab) and biomass (Bio) of each species catching during the study with their trophic category (Tr.Cat) and vulnerability (Vul).

Family Species Abun. Tr.Cat Vul. Bio.

Congridae Conger conger (Linnaeus, 1758) 14 CMC 60 28.52 Haemulidae Pomadasys incisus (Bowdich, 1825) 28 CMS 40 15.15 Labridae Labrus merula (Linnaeus, 1758) 105 CMC 37 141.45 Symphodus tinca (Linnaeus, 1758) 113 CMS 50 121.80Moronidae Dicentrarchus labrax (Linnaeus, 1758) 54 CMC 42 175.96Mugilidae Mugil spp 18 CMS 49 22.96 Mullidae Mullus surmuletus (Linnaeus, 1758) 8 CMS 44 8.92 Muraenidae Muraena helena (Linnaeus, 1758) 20 CMC 53 71.71 Pomacentridae Chromis chromis (Linnaeus, 1758) 15 CMM 10 1.87 Sciaenidae Sciaena umbra (Linnaeus, 1758) 74 CMC 48 73.79 Scorpaenidae Scorpaena notata (Rafinesqe, 1810) 19 CMC 61 10.49 Scorpaena scrofa (Linnaeus, 1758) 8 CMC 61 5.90 Serranidae Epinephelus marginatus (Lowe, 1834) 2 CMC 62 6.16 Serranus cabrilla (Linnaeus, 1758) 3 CMC 32 1.29 Serranus scriba (Linnaeus, 1758) 107 CMC 32 63.36 Sparidae Boops boops (Linnaeus, 1758) 64 CMM 45 8.48 Dentex dentex (Linnaeus, 1758) 34 CMC 52 146.21 Diplodus annularis (Rafinesqe, 1810) 72 CMS 21 19.40 Diplodus cervinus (Lowes, 1838) 1 CMS 50 1.28 Diplodus sargus (Linnaeus, 1758) 156 CMS 35 148.94 Diplodus vulgaris (E. Geoffrey Saint-Hilare, 1817) 145 CMS 22 126.17 Lithognathus mormyrus (Linnaeus, 1758) 17 CMS 53 8.24 Oblada melanura (Linnaeus, 1758) 92 CMM 33 61.06 Pagellus acarne (Risso, 1826) 24 CMS 45 6.74 Pagellus erythrinus (Linnaeus, 1758) 4 CMS 45 1.37 Sarpa salpa (Linnaeus, 1758) 79 HBV 50 38.90 Sparus aurata (Linnaeus, 1758) 99 CMS 41 111.66 Spondyliosoma cantharus (Linnaeus, 1758) 1 CMM 63 0.46

Figure 4. Mean diary CPUE (nºindividuals/lines·hour) and BPUE (grams/lines·hour) for each season at Tabarca MPA. Bars indicate the standard error.


26

Table 3. PERMANOVA analysis of variance among Seasons for CPUE (abundance/lines·hours) and BPUE (biomas/lines·hour). df.: degree of freedom; M.S.: mean square; F: F real; level of signification: ns, non significant; *=P<0.05; **=P<0.01; ***=P<0.001.

CPUE BPUE Source df M. S. F M. S. F Season 3 8375.00 2.852 *** 5454.53 1.657**

Residual 28 2936.04 3291.35 Total 31

Table 4. For both variables, CPUE and BPUE there are: Above diagonal: average dissimilarity between seasons; below diagonal: the first three species primarily account for the observed differences with their respective average dissimilarity within brackets. A:autumn; W:winter; Sp: spring; S: summer.

CPUE A W Sp S

A 77.7 79.9 83.0 W D.sargus(11.3) 79.6 83.9 S.tinca(8.3) D.vulgaris(8.2) Sp B.boops(9.5) S.tinca (8.5) 81.1 D.sargus(8.5) D. vulgaris(8.4) D.vulgaris(8.4) B.boops (8.3) S D.anularis(9.6) D.anularis(9.8) D.anularis(9.0) D.sargus(9.5) D.sargus(8.9) D.vulgaris(7.5)

D. vulgaris(7.8) S.tinca(7.0) B.boops (7.3)

BPUE

A W Sp S A 80.10 80.91 85.2 W S.tinca(10.5) 79.89 85.3 D.sargus(9.8) L.merula(7.6) Sp D.vulgaris(8.8) S.tinca(10.5) 82.9 D.sargus(8.3) D.vulgaris(9.1) S.aurata(7.3) S.aurata(7.3) S D.sargus(8.0) D.labrax(8.1) D.labrax(9.1) D.labrax(7.4) L.merula(7.7) D.vulgaris(7.0)

L.merula(7.0) D.sargus(7.5) D.sargus(6.6)

________________________________________________________________Chapter 1

27

hours and BPUE of 90.82 ± 20.55 g/line hours) (Fig 4). There were differences between seasons (CPUE, F=4.106 p<0.01 and BPUE, F=3.80 p<0.05), and SNK confirmed that these differences were the result of high catch rates in summer. However, 27 anglers (10%) did not catch any fish during their outing, with the highest number of “zero-fish anglers” per day recorded in winter, when BPUE and CPUE were at their lowest.

PERMANOVA analyses found statistically significant differences between seasons for two variables (Table 3), and the subsequent pair-wise comparison showed that for CPUE the difference was between the warm and cold seasons (spring/summer and autumn/winter). However, for BPUE the difference was only between summer and the other seasons.

SIMPER or analyses of dissimilarity between groups confirmed again this variability which was detected by the previous analysis and detected the species that most account for the observed differences (Table 4). For the CPUE variable, with the exception of S.tinca, the other species that showed differences between seasons were all Sparids. Presence of his species, together with D.labrax and L.merula, also marked differences between seasons for BPUE. The high BPUE of D.labrax in summer was responsible for the differences between this season and the others. For spring, was the high BPUE of D.vulgaris the responsible of marked these differences.

3.6. Estimation of total effort and total catch

Taking into account the seasonal and week effort distribution as well as the BPUE per season and species, we estimate that angling activity in the Tabarca Marine Reserve would be capable of extracting 4.9 tons of fish per year. The total annual catch estimated per species has been compared with data from landings in the Santa Pola, the nearest commercial harbour to MPA (Table 5).

Table 5. Annual landings per species in Santa Pola Harbour per the artisanal fishery, the catch estimations per specie extract for anglers annually calculated by the Tabarca marine reserve and the proportion between this.

Sta Pola MPA % C. conger 19836.88 74.92 0.37 D. annularis 8166.00 265.24 3.24 D. dentex 2610.70 259.34 9.93 D. labrax 2821.82 929.93 32.95 D. sargus 7842.94 450.47 5.74 D. vulgaris 1674.50 396.45 23.67 L. merula n.d. 352.74 L. mormyrus 12697.90 91.12 0.71 M. helena 3003.80 262.19 8.72 Mugilidae 15747.30 86.64 0.55 O. melanura 776.40 202.38 26.06 S. aurata 7895.69 314.89 3.98 S. salpa 3500.90 261.32 7.46 S. scriba 120.50 175.95 146.02 S. scrofa 4177.60 17.11 0.40 S. tinca 307.00 314.78 102.53 S. umbra 1536.82 134.11 8.72

The recreational fishing catch of S.tinca and S. scriba proved to be greater


28

than the artisanal fishing landing, considering the low commercial value. Regarding other species of greater economic importance, such as D.dentex, M. helena, S. umbra, Diplodusspp and S. aurata, the angling catch was not as great, but was still of considerable proportions in comparison with the total commercial catch.

4. Discussion

The effect that recreational fishing can have is often underestimated, as a single angler is thought to have a much lower impact on fish stocks compared with commercial fisheries (a trawler or purse seine, for example), but the cumulative impacts of thousands or millions of anglers should not be taken lightly. The European Parliament recently recognised (EP, T6-0255/2009) that recreational fishing may have a significant impact on fish stocks and that within two years of the date of the new regulation, “the Member States may estimate the impact of recreational fisheries conducted in their waters”.

The inherent difficulty of monitoring such highly diverse and disperse recreational fishing activities may contribute to this underestimation of the impact (Arlinghaus et al., 2002). In our study, we observed quantitative differences between our own data and information from the fishery guard, indicating how difficult it can be to obtain accurate data of this type. Some objective indicators could be useful, such as licence control. However, we observed an additional problem in this study, as a high proportion of anglers (39.1%) did not hold such a licence.

Therefore, the lack of more accurate long-term monitoring programmes, the lack of official data on actual activity (“angling effort”) and the fact that recreational fisheries are so widely dispersed make it difficult to obtain an accurate picture of the real impacts of angling.

4.1. Anglers’ profile and motivations

The angling effort has shown an increase in recent years. The pattern of angling activity, both in the long term series and the more thorough study, was marked by greater effort during holiday periods and at weekends which reflect a leisure-based motivation rather than an economic interest. This is a common pattern for anglers, who predominantly practice the recreational fishing for multitude of nonconsumptive objectives during leisure time, where catching fish for consumption is only one of many driving forces for practice this activity (Fedler and Ditton, 1994; Policansky, 2002; Cowx, 2002b; Arlinghaus and Mehner, 2004). Therefore it can be assumed that angling is not motivated by the same economic incentives that drive commercial fishery to overexploit fish stocks (see Myers et al., 1997). Nonetheless, the calculated 4.9 tons of fish extracted from the protected area reflects the fact that catching fish is the most important aspect which determine the main “product” of the angling experience, the satisfaction; hence catching fish matters for many anglers to a great extent (e.g., Graefe and Fedler, 1986; Petering et al., 1995; Connelly and Brown, 2000; Arlinghaus, 2006).

________________________________________________________________Chapter 1

29

4.2. Recreational fishing effects on fish assemblages

The angling target areas coincide with areas which provide refuge for fishing resources that are not accessible for professional fishing. These coastal areas are very important for many fish species (García-Rubies and McPherson, 1995; McPherson, 1998), as crucial processes in their life cycle (such as reproduction, shelter of larvae or juveniles, and feeding) and stock recruitment usually occur here, which is why any alteration of these processes could lead to harmful consequences for the fish population.

Our study revealed that a large proportion of catches were of mesophagous and macrophagous species, indicating that those species high up on the food chain are the most impacted by the recreational angling in Tabarca. Paine (1996) indicated that even a small-scale extraction of these high trophic levels from fish populations leads to a loss of diversity and marked changes in the food webs. Depending on role and dominance within the ecosystem, the removal of a substantial part of a population can significantly affect the trophic process and community structure (De Roos and Persson, 2002; Donald and Anderson, 2003; Ward and Myers, 2005; Worm et al., 2005), especially in aquatic ecosystems where the strength of top-down control is greater than in terrestrial ecosystems (Shurin et al., 2002), and the natural diversity at higher trophic levels is low (Petchey etal., 2004). Therefore, given the high trophic level of the species preferences of recreational fishers and the high

harvest rates observed in this study (a total of 4.9 tons), it can be considered that recreational fishing has the potential to affect the trophic structure of fish populations.

The large number of species of high intrinsic vulnerability that are caught indicates how harmful it may be to fish for these species, whose populations are at a high risk of collapse due to fishing (70% of catches are of species between moderate and very high vulnerability). The vulnerability index can provide valuable information to evaluate the robustness or capability of recovery of a fish community in view of the impact from angling. The fragility of the community studied showed the damage that fishing can cause and the danger that this fish population runs if angling continues over time.

4.3. Socioeconomic conflicts

In our study, an overlap between recreational and commercial fishing has been detected in catch composition, a fact that may produce tension and conflict between the two sectors, as has occurred in other areas where professional fishing was forbidden but recreational fishing was allowed (Westera et al., 2005). The ideas of protecting breeding stocks, improving recruitment to neighbouring areas and restocking marine species of commercial interest were the most important initial goals behind creating the Tabarca Marine Reserve in 1986 (Ramos-Esplá, 1985). Such protection was supposed to provide direct benefits for the adjacent fisheries by means of a spill-over process (Clanahan and


30

Mangi, 2001; Roberts et al., 2001), but the view of the local artisanal fishery is different. During the study period, we recorded opinions of the local artisanal sector, which perceives that the traditional fishing ground is now “private property” for recreational fishers. This dissatisfaction between users may lead to the rejection and failure of the marine protected area (Arlinghaus, 2005).

The economic importance of recreational fisheries on local and regional economies has been recognised in this study and in others (Cowx, 2002a; Cowx et al., 2004). In this regard, we sought the opinion of the owners of restaurants (25), hotels and guesthouses (12) and supermarkets (6) on Tabarca who earn a living mainly from tourism. During the period studied, fishing anglers spent a total of €242,800, mainly on vessel transport and accommodation. In the three months of summer, Tabarca receives around 3,500 visitors per day, many of whom come for the sun and beach, but for the rest of the year the prospect is rather different. Most of the visitors to Tabarca were recreational anglers who visited the island at the weekend. When angling is forbidden, as in the Tabarca case, managers must take into account those sectors that would be economically affected. The promotion of ecotourism by the management agency based on non-extractive activities may provide economic benefits for the local economy that could be used to compensate the cost incurred by restricting recreational fishing, as has happened in other MPAs (Forgaty et al., 2000; Dixon et al.,

2001). The development of ecotourism activities, such as sailing, kayaking and snorkelling, for example, could improve social acceptability by providing sources of income for the local population (Dixon et al., 1993; Attwood et al., 1997). Designing educational environmental and cultural programmes may also constitute two ways of offering a new appeal for tourists, while making the visible proof of the results of protection measures and other elements of natural and cultural heritage known to visitors.

4.4. Recommendations

As much in this study as in others around the world, the integration of psychological, sociological, and economic perspectives in recreational fisheries management is highly important (Cox and Walters, 2002; Post et al., 2002; Arlinghaus, 2005; Cury etal., 2005). To provide a more realistic picture of angling activity and to identify promising management strategies, more research is needed to explore the dynamics between anglers, their target species and the social and economic dimension.

Long-term monitoring programmes of the impact of recreational fishing, with standardised and comparable protocols using a social, economic and biological approach must be established to obtain comparable and more accurate data with which to understand all aspects of angling, so as to produce a coherent management plan according with the needs of each area.

________________________________________________________________Chapter 1

31

As explained above, recreational fishing produces similar negative effects on fish populations and other communities or habitats as the effects of commercial fishing, which must not been underestimated. Furthermore, very little information exists on the real situation of angling, in terms of both the effort and catch and the effect on the food chain or on other habitats and communities. Some studies have shown that partial closures of MPAs are ineffective as conservation tools for fish communities, and that only no-take marine protected areas should be created (Denny and Baboock, 2004). The fact of the matter is, MPA’s that

have commercial fishing bans in place may attract recreational fishers and produce a negative effect of attraction and increased effort, which will in turn lead to greater impact on the fish communities that have been protected. In view of these and previous arguments, the unknown and uncertain element nature of angling and specially in the MPA cases, management guided by precaution may be the most coherent solution, together with the prohibition of recreational fishing, as there is much uncertainty over the possible outcome and mistakes could lead to serious or irreversible damage.


32

1

____________________________________________________________________________Chapter 2

Assessing recreational fishing impact at Serra Gelada Marine Park (SW Mediterranean): a baseline study for its future regulation.

Abstract

Recreational fishing is an important leisure activity which has similar effects on exploited populations than commercial fishing and in some cases, can generate an effort and catch equal or higher than it. Nowadays, recreational fishing does not have any regulations in the Serra Gelada Marine Park and this study is a good opportunity for evaluate the associated impact and design future management strategies. Recreational boat fishing was studied employing different methods: aerial surveys, interviews and competition data. The relationships of recreational fishers with other users and other sociological aspects were studied too. Recreational fishing is an extensive leisure activity which is intensively practiced by local residents during the whole year. The recreational boat fishing effort is very high, mainly focused on the buffer zone and the special protection area.

In the buffer zone, three important areas were identified due to their highest effort density, two of them surrounding aquaculture installations. A total of 33 species were recorded as the most frequently caught by fishers in the study area. The competition data showed a mean catch per unit effort of 0.465 ± 0.082 kg boat-1 hour-1 and 0.323 ± 0.054 individuals boat-1 hour-1. The annual total catch of recreational boat fishing in the Serra Gelada Marine Park was estimated in 24.90 t. The importance of recreational fishing impact on fisheries resources and their sociological implications had been evidenced on this study. Recreational boat fishing should be taken into account in the management strategies on this and other MPA.

Keywords: Marine Protected Areas, coastal management, fishing effort, human activities, conservation, aerial survey.

1. Introduction

Recreational fishing is one of the most frequent leisure activities in coastal areas, and it involves large number of people and consequently high levels of fishing effort (Cowx, 2002; Pitcher and Hollingworth, 2002; Westera et al., 2003; Cooke and Cowx, 2006), which, in some cases, may be translate in an effort and catch equal or higher than in commercial fisheries (Pollock, 1980; Dunn et al., 1989; Steffe et al., 1996; Beal et al., 1998; Murray-Jones and

Steffe, 2000). Commercial and recreational fishing have similar demographic and ecological effects on fished populations, such as truncation of the natural age and size structure, decline of fish abundance and biomass and changes in fish assemblage composition (Coll et al., 1999; McPhee et al., 2002; Westera et al., 2003; Coleman et al., 2004; Cooke and Cowx, 2004, 2006; Lewin et al., 2006; Rius, 2007).

35

Anthropic impacts in Mediterranean MPAs_________________________________________________

In the Mediterranean Sea, recreational fishing is particularly important, representing more than 10% of total fisheries production in the area (EU, 2004). Recently, as a sign of this importance, the European Parliament recognised that recreational fishing may have a significant impact on fish stocks and that the Member States may estimate the impact of recreational fisheries conducted in their waters (EP, T6-0255/2009). Moreover, particularly in the Mediterranean, recreational fishing has important economic, social and cultural roles, because there is an overlap with commercial fishing since it is largely domain of small-scale concerns operating in coastal areas (Gordoa, 2009). Despite the importance of the recreational fishing, there are very few studies in the Mediterranean about the knowledge of this activity (Coll et al., 2004; Morales-Nin et al., 2005; Lloret et al., 2008; Gordoa 2009).

In the case of multiple use marine protected areas, all activities have to be managed in order to carry out the conservation and protection objectives, minimize conflicts and allow all users follow with their activities ecologically sustainable. The impact of recreational fisheries on coastal fauna could be particularly sizeable in and around marine protected areas (MPAs), where an increasing number of visitors fish for pleasure (Cooke et al., 2006).

Serra Gelada Marine Park is a recent established MPA, in where uses (including fishing activities) have not been regulated yet. It offers a good opportunity to study the recreational boat fishing impact and recommend

sustainable management strategies. Therefore, the objectives of this study were to contribute to the knowledge of exploitation of fish resources by the recreational boat fishing and to provide essential data needed for integrated management in coastal zones, particularly in MPAs. To reach these objectives we i) studied the recreational fishermen profile, their motivation and perceptions, ii) identified the most affected fish species, iii) studied the spatial and temporal recreational fishing effort distribution as well as iv) estimated the annual recreational fishing capture and effort.

2. Material and methods

2.1. Study area

The study area is located in Serra Gelada Marine Park (4 920 ha) (southwestern Mediterranean, Spain) (Fig.1). The coastal shelf is characterized by great environmental heterogeneity attributable to the co-occurrence of Posidonia oceanica and Cymodocea nodosa meadows, and sandy and rocky seabeds.

The Serra Gelada Marine Park was created in 2005, however the regulation of the activities inside the MPA does not come into effect yet. It is subdivided into three management zones: i) Special Protection, which contain three marine areas, Benidorm Island, Mitjana Island and Mascarat-Morro Toix Barrier; ii) Special Use, which contain two aquaculture installations; and iii) Buffer Area (Generalitat Valenciana, 2005). Currently the recreational fishing is allowed in all the MPA except in the

36

____________________________________________________________________________Chapter 2

Figure 1. Location of study area with the protection areas

aquaculture installations. Recreational fishing regulations in the area (Generalitat Valenciana, 1998) set up a maximum daily catch (5 kg of total catch plus one piece by fisher) and a minimum catch size per species (1615/2005 Real Decreto del Ministerio de España). Moreover, a license of recreational fishing is required, the number of gears that can be used simultaneously is limited at two per license, and all type of nets, traps and long lines are forbidden.

.2.2. Recreational fisher interviews

Studies elsewhere using interview data show that this approach can produce a lot of information useful for assessing fisheries (Neis et al., 1999; Rocha et al., 2004; Forcada, 2010). Meetings with the recreational boat fishers were done in June 2007 in the two main harbours of the study area (Benidorm and Altea) where the total number of recreational fishing boats recorded in this year was

140. During these meetings the fishers were informed about the development of this project and were urged to take part actively with a personal interview. A questionnaire of 21 items was used to obtain information of their personal profile as recreational fisher (gender, age, place of residence, years of fishing experience, …), their perception of the Marine Park, their relationship with other Marine Park users and their fishing habits (how much time they fish per day and year, what seasons and hours of the day they prefer, usual fishing places and the species most frequently captured). In all, 84 questionnaires were conducted.

2.3. Aerial surveys

Aerial surveys have been frequently used to assess recreational fishing effort (Pollock et al., 1994), because it is an efficient method to estimate boats distribution over large areas (Malvestuto, 1983). Sampling was

37


stratified by Week Period and period during the day (Time Day). In each Week Period (weekday: Monday to Thursday; weekend: Friday to Sunday) were surveyed two sampling periods per day: morning (9:00-13:00) and afternoon (16:00-20-00). Three samples were randomly selected for each combination of the two factors. Thus, we conducted 12 aerial surveys from June to September 2007. This sampling design allows recognize patters on the temporal effort distribution (Tisman and Whitmore, 2006). Aerial sampling was limited to days with good visibility and flying conditions, and survey flights were conducted at around 100 m altitude. On each survey, we covered all the study area in an hour, and the geographical position of all boats sighted was recorded by means of a GPS. A vessel was considered a fishing boat actively engaged in fishing if a fishing rod o handline in use was observed on-board.

Instantaneous recreational boats fishing effort was obtained as number of boats fishing in the study area in one hour, and it was estimated for each combination of the two factors (Week Period and Time Day). The data of each survey were included in a geographic information system to obtain the spatial distribution of the recreational fishing effort density. The study area was divided using a grid of 500 × 500 m cell size. Effort density was calculated as the sum of the number of recreational fishing boats in each cell divided by the cell surface, and it was expressed as number of boats hour km-2.

Analysis of variance (ANOVA) was used to test for significant differences in recreational fishing effort density. The experimental design consisted of 3 factors: Week Period (2 levels: weekends and weekdays; fixed), Time Day (2 levels: morning and afternoon; fixed and orthogonal) and Protection Level (3 levels: Special Protection Area, Buffer Area and Non-protected Area.; fixed and orthogonal). Therefore, with n=3 sampling days, there was a total of 36 samples. When the ANOVA F-test was significant, post hoc analyses were conducted using Student-Newman-Keuls (SNK) multiple comparisons (Underwood, 1981). Before analysis, Cochran’s test (Cochran 1951) was used to test for homogeneity of variance. As the transformations (�(x + 1) and ln(x + 1)) did not remove the significant heterogeneity, analyse was performed on the untransformed data, but with the F-test p-value set at 0.01 because ANOVA is robust to departures from this assumption (Underwood, 1997).

2.4. Recreational fishing competition

The competition data used to undertaken this study were obtained from two recreational fishing clubs, which held several competitions in Serra Gelada Marine Park during the period studied (2007). The competitions consisted in catch the maximum number of individuals and weight of fishes in a period of 10 hours. The recreational fishing clubs were reluctant to give the detailed competition data; they omitted the names of boats, the catch by species, and the catch for each participant. A total of 25 event competition were used to obtain the mean catch per unit effort

38

____________________________________________________________________________Chapter 2

in the study area expressed in abundance (CPUE, individuals boat-1 hour-1) and biomass (BPUE, kg boat-1 hour-1), in order to estimate the total catch of the recreational fishery from boat in Serra Gelada Marine Park.

2.5. Estimations of total annual fishing effort and catch

Estimates of total annual fishing effort and total annual catch of the recreational boat fishery in Serra Gelada Marine Park were calculated by combining the data from the interviews (fishing seasonality), aerial surveys (instantaneous recreational boats fishing effort) and recreational fishing competition (CPUE, BPUE).

3. Results

3.1. Recreational fisher profile

Recreational boat fishing in Serra Gelada Marine Park is a totally male activity. The typical recreational boat fisherman is a mean of 58.1 ± 1.6 (± standard error) years old (Figure 2a), resides in a town located closer that 20 km from Serra Gelada Marine Park (Figure 2b) and has more than 21 years of fishing experience (Figure 2c). A great proportion of the fishers interviewed (57%) are not members of a fishing club.

.

Figure 2. Percentage of recreational fishers interviewed by (a) age, (b) Residence distance to MPA, (c) Fishing experience, (d) Boath length, (e) Number of recreational fishers by boat, (f) Influence of MPA existence for their decision of choose this fishing place.

0102030405060708090

Nothing Little Very much

Influence of MPA existence Boat length (m)

0102030405060708090

< 5 5 to 6.5 6.6 to 8 > 8

0102030405060708090

30-45 46-55 56-65 >65

Age (years)

0102030405060708090

0-20 21-50 51-100 > 100

Residence distance to MPA (km)Fishing experience (years)

0102030405060708090

1 to 5 6 to 11 to > 20

a) b) c)

d)

Number of recreational fishers by boat

0102030405060708090

1 2 3 >3

e) f)

Pe

rcen

tage

P

erce

ntag

e

39


The recreational fishing boats have a length that average 6.37 ± 0.17 m (Figure 2d), are manly taken up by at least two fishers (Figure 2e) and most of them (57.2%) have GPS and echosounder Most of the recreational fishers interviewed (69%) know the existence of Serra Gelada Marine Park, although for the great proportion (Figure 2f) the existence of the MPA does not have influence on their decision of fishing there.

The great proportion of recreational fishers affirms that have good terms with other recreational and professional fishers (Figure 3). In contrast, a high percentage of recreational fishers declare bad relationships with spear fishers, divers, aquatic sky enthusiasts and wind surfers. The most conflictive relationship is with the aquaculture and others, declared by about 23% of the fishers interviewed.

The recreational fishing activity in the Serra Gelada Marine Park is carried out

during the whole year (Figure 4a) and along all the day (Figure 4b). Recreational fishers fish an average of 117.14 ± 12.5 days year-1 and 4.91 ± 0.16 hours day-1. Altea’s Bay is the place most preferred by the recreational fishers interviewed to fish inside the MPA (Figure 4c).

According to the interview answers, in all, 32 species belonging to 13 families (most are fish but are two cephalopods) are the species most frequently caught by the recreational fishers using diverse fishing gears (angling, handlines, trolling lines). The species with highest frequency of occurrence are: Serioladumerili, Trachurus spp., Coryphaena hippurus, Pagellus erythrinus, Pagellus acarne, Scomber scombrus andEuthynnus alletteratus (Table 1).

3.2. Aerial surveys

During the aerial surveys, a total of 291 recreational fishing boat positions were recorded around the Serra Gelada Marine Park. Instantaneous recreational

Figure 3. Percentage of recreational fishers interviewed that have no, conflictive or good relationship with other MPA users.

0102030405060708090

100

Diving

Aquati

c Ski

Wind

Surfi

ng

Aquac

ulture

& ot

hers

Profess

ional

fishin

g

Recrea

tional

fishin

g

Spear f

ishing

No relationConflictiveGood

Perc

enta

ge

40

____________________________________________________________________________Chapter 2

Figure 4. Percentage of recreational fishers interviewed by (a) fishing annual habits, (b) favourite fishing place and (c) fishing diary habits.

Fishing annual habits

010203040506070

8090

Whole year Non-summer Only summer

a)

Favorite fishing place

0102030405060708090

Altea's

Bay

Benido

rm's B

ay

Mitj

ana I

sland

Benido

rm Is

land

Serra

Gela

da

Mas

carat

b)

Fishing diary habits

0102030405060708090

Whole day Morning Afternoon

c)

Perc

enta

ge

Perc

enta

ge

Perc

enta

ge

41


Table 1. Frequency of occurrence of the species on the total catch of the recreational fishers interviewed.

Class Family Species Frequency Cephalopoda Loliginidae Loligo spp. 1.58 Octopodidae Octopus vulgaris (Cuvier. 1797) 0.64 Osteichthyes Phycidae Phycis phycis (Linnaeus 1766) 0.59 Serranidae Serranus cabrilla (Linnaeus. 1758) 0.47 Serranus scriba (Linnaeus. 1758) 2.93 Moronidae Dicentrarchus labrax (Linnaeus. 1758) 0.59 Pomatomidae Pomatomus saltatrix (Linnaeus. 1766) a 1.00 Carangidae Seriola dumerili (Riso, 1810)a 10.78 Trachinotus ovatus (Linnaeus. 1758) a 0.12 Trachurus spp. 10.43 Corypahenidae Coryphaena hippurus (Linnaeus 1758) a, c 9.67 Sparidae Boops boops (Linnaeus. 1758) 0.70 Dentex dentex (Linnaeus. 1758) a 4.34 Diplodus annularis (Rafinesqe. 1810) 0.70 Diplodus puntazzo (Cetti. 1777) 0.47 Diplodus sargus (Linnaeus. 1758) 1.99 Diplodus vulgaris (E.Geoffrey Saint-Hilare. 1817) 2.17 Lithognathus mormyrus (Linnaeus. 1758) a 0.59 Oblada melanura (Linnaeus. 1758) 0.47 Pagellus acarne (Risso. 1826) 7.56 Pagellus bogaraveo (Brünich. 1768) a 2.52 Pagellus erythrinus (Linnaeus. 1758) a 9.14 Pagrus pagrus (Linnaeus. 1758)a, b 1.46 Sparus aurata (Linnaeus. 1758) 1.70 Centrachantidae Spicara maena (Linnaeus. 1758) 1.35 Labridae Xyrichtys novacula (Linnaeus. 1758) 4.69 Sphyraenidae Sphyraena sphyraena (Walbaum. 1792) a 0.94 Scombridae Auxis rochei (Riso. 1810) a, c 1.17 Euthynnus alletteratus (Rafinesque. 1810) a, c 6.50 Sarda sarda (Blonch. 1763) a, c 3.46 Scomber scombrus (Linnaeus. 1758) a, c 6.56 Thunnus thynnus (Linnaeus. 1758) a, b, c, d, e 1.17

a Species considered vulnerably (Cheung et al., 2007) b Species included in the Red List of the IUCN c Species included in the Annex I of the UNCLOS d Species included in the OSPAR convention

e Species included in the BARCOM convention

42

____________________________________________________________________________Chapter 2

Table 2. Mean instantaneous recreational boats fishing effort standard error (boats hour) during each time day and week period.

Weekday Weekend Morning 25.33 ± 3.56 37.33 ± 4.26Afternoon 11.00 ± 3.24 23.33 ± 3.27

Table 3. Mean recreational boat fishing effort density standard error (boats hour km-2) in each management zone and outside Serra Gelada Marine Park during each time day and week period.

Weekday Weekend Morning Afternoon Morning Afternon

Special protection 0.73 ± 0.39 0.12 ± 0.07 0.55 ± 0.13 0.12 ± 0.15 Buffer area 0.44 ± 0.10 0.18 ± 0.07 0.65 ± 0.07 0.39 ± 0.02 No protection 0.03 ± 0.01 0.03 ± 0.00 0.08 ± 0.03 0.07 ± 0.03

Table 4. Three way ANOVA results for the recreational boat fishing effort density. Bold type indicates significant differences at the 99% level. d.f: degrees of freedom; M.S.: mean square; F: F ratio; P: significance level.

Source of variation Df M.S F P Weekend/Weekday (W) 1 0.027 0.75 0.395 Morning/Afternoon (T) 1 0.606 16.74 0.000 Protection level (P) 2 0.487 13.45 0.000 W x T 1 0.007 0.20 0.657 W x P 2 0.069 1.91 0.170 T x P 2 0.200 5.52 0.011 W x T x P 2 0.009 0.25 0.783 Residual 24 0.036

boats fishing effort was greater during weekend days and during the morning (Table 2). Effort density was significantly higher during the morning than in the afternoon (Table 3 and 4). Special Protection areas and Buffer area recorded a significant greater effort density than the Non-protected area on any week or diary period.

In the Buffer area there were three zones that concentrated intensively fishing effort (Figure5): around the two aquaculture installations and in the southern cape of the MPA (locally

know by the fishers as “Punta del Cavall”). In the Special Protection areas Benidorm and Mitjana Islands, light concentrations of fishing effort were also observed.

3.3. Recreational fishing competitions

The total catch recorded was 1162.5 kg (813 individuals) corresponding to 2500 hours of total fishing effort. The CPUE and BPUE estimated were 0.323 ± 0.054 individuals boat-1 hour-1 and 0.465 ± 0.082 kg boat-1 hour-1, respectively.

43


Figure 5. Spatial distribution of recreational boat fishing effort density in the Serra Gelada Marine Park

3.4. Estimations of total annual fishing effort and catch

Expanding the mean instantaneous recreational boats fishing effort (Table 2) to the four hours considered in each period of the day sampled, the total diary effort was estimated in 145.32 boats hour for weekdays and in 242.64 boats hour for weekend. As a result, 1309.20 boats hour were estimated as the total week effort. Following the results of the interviews, the 98.2% of the recreational fishers declared that are actively fishing in summer, consequently the effort of this period has been take as the maximum value and the total effort in the three summer moths has been estimated in 16191.72 boats hour. The 76.9% of the recreational fishers interviewed

declared that fish during the whole year, so taking into account this result, the effort for the non-summer period has been estimated on 37402.87 boats hour, therefore the total annual effort was 53594.60 boats hour. Considering this total annual effort and the CPUE and BPUE from the recreational fishing competitions, the total annual catch in the study area was estimated in 17418.2 individuals and 24.90 t.

Discussion

Serra Gelada Marine Park is used extensively as a recreational fishing location by the residents and the total annual effort estimated is very high, reflecting that is a very intense activity. Similar values on individual fishing diary effort were found in Majorca Islands and Cap de Creus MPA, where

44

____________________________________________________________________________Chapter 2

the mean fishing effort is about 4 and 4.1 hours day-1 (Morales-Nin et al., 2005; Lloret et al., 2008). However, the annual fishing effort on Serra Gelada Marine Park is higher since the recreational boat fishing is practised almost during the whole year.

As a consequence of this high fishing effort, the total catch extracted annually by recreational fishers is relevant and should not be trivialized by managers. The effects of recreational fishing on littoral species of Serra Gelada Marine Park could be substantial considering the large number of species targeted (32), which is coincident with the number of species caught by boat fishers in other similar areas in the north-western Mediterranean (e.g., 33 species in Cap de Creus; Lloret et al., 2008). Fifteen of the species frequently caught by recreational boat fishers in Serra Gelada Marine Park are considered vulnerably (Cheoung et al., 2007) and seven of these are included in international conventions for the protection as UICN, UNCLOS, OSPAR and BARCOM (Table 1), therefore the effects of this activity on threatened species is rather important.

As commercial and recreational fishing have similar demographic and ecological effects on fished population (Westera et al., 2003; Coleman et al., 2004; Cooke and Cowx, 2004, 2006; Lewin et al., 2006; Rius, 2007), both activities should be taken into account in the management of fishing resources. Furthermore in the Mediterranean, the interaction between commercial fishing and recreational boat fishing is really important because artisanal fisheries,

which exploit coastal areas, represent 80% of the EU Mediterranean fleet (COM, 2002). Some similarities have been found between the recreational boats fishing studied here and the artisanal fisheries described by Forcada et al. (2010) in a neighbour area. The fishing grounds of both fisheries have a common bathymetric range (mainly at depths <40 m) and the 56.25% of the most frequently catch species on this study are target species or species that represent an important proportion of the capture for the artisanal fisheries of the region (Forcada et al., 2010). Conflicts between both sectors is endemic to all development countries (Kearney 2002; Cooke and Cowx 2006) and these conflicts have increased in areas where professional fishing was forbidden but recreational fishing was allowed (Westera et al., 2005). In fact, some studies demonstrated that the recreational fishing pressure may be sufficient to deplete fished populations below that of adjacent protected areas (Westera et al., 2003). Denny and Babcock (2004) also indicated that the partial closures are ineffective as conservation tools and that the fishing pressure on closed commercial fishing areas where the recreational fishing is allowed is at least as high as other unprotected areas. However, on this study the major proportion of the recreational boat fishers interviewed affirmed that they have a good relationship with professional fishers. Nevertheless, artisanal fishers excluded from MPAs, in where recreational fishing is allowed, showed their dissatisfaction with this type of regulations, which may lead to the

45


rejection and failure of the MPA (Arlinghaus, 2005).

The inherent difficulty of monitoring the highly diverse and disperse angling activities may contributed to the underestimation of angling impacts (Arlinghaus et al., 2002). This lack of data of a real participation (“angling effort”) by the competent administrations and the high dispersion of the recreational fisheries hinder the obtainment of an accurate picture of the real impacts of angling which difficult the management actions. Although the control of license or the membership at fishing clubs may be a useful indicator of effort, as has been observed on this study, a high proportion of anglers did not have one or not be member of one (57%). The recreational competitions data seem be a good source of information about the species composition and rate of catches, as other authors had recognized previously (Coll et al., 2004; Morales-Nin et al., 2005; Gordoa 2009). On this study, obtaining of this type of data has been a hard work, because the recreational fishing clubs which provide the data were reluctant to provide the competition results. Probably this attitude reflects fear or distrust by the recreational fishers in front of the make decision managers in this case, which would indicated the lack of integration and cohesion of recreational fishing sector and management authorities.

Aerial surveys has turned out a good method for obtain spatial distribution of effort, which had been demonstrated previously (Malvestuto 1983; Pollock et al., 1994; Tisman and Withmore, 2006).

The spatial effort in the Serra Gelada Marine Park indicated the preference of the recreational boat fishers by concretely zones where probably the fishing yield is high or target some very appreciate species, which coincide with areas with a specific protection status. A study carried out in 2007 in the Serra Gelada Marine Park (Sanchez-Jerez, etal., 2007) by means of underwater visual census showed a higher abundance of fishes in the Mitjana, Benidorm Island and Punta del Cavall than in the other localities studied, which confirm the motivation of recreational boat fishers to choose these places. Concretely the Punta del Cavall zone is a preferred place by the recreational fishers because the high catches of Xyrichtys novacula, a more appreciate species in the zone which is caught during the afternoon, explaining why the pattern variation among protection levels differed over the dairy periods. A great effort also had been observed around the aquaculture installations. Machias et al. (2004 and 2006) reported an increase of the fishing production around the zones where the aquaculture installations had been located, which probably explain the cause of recreational fishers preference for these areas and the conflicts that recreational fishers declared to have with this use. This conflicts must be take into account for the effective management of coastal uses, mainly in MPA.

Total number of recreational fishing boats recorded during the time study in the two main harbours was 140, and 84 recreational boat fishers were interviewed, reflecting a high

46

____________________________________________________________________________Chapter 2

participation. The response rate during the interviews was high, reflecting the willingness of recreational boat fishers to cooperate, and a desire to understand the reasons for changes in fished resources availability and catch rate at different sites and over time. In general, recreational fishers have an inherent interest in the conservation and management of the fisheries resources on which their leisure activity depends (Arlighnaus, 2006). Allowing fishers to participate in developing regulations, could lead to increased management success (Sullivan, 2003). On a global study carried out by Granek et al., (2008) the potential for recreational fisheries to be an important conservation force was evidenced. The fishers engagement is more noticeable even thought the resources are small because fisher are more likely to feel responsible for its conservation (Arlinghaus 2006) and may perceive that there is a great probability that their actions will affect change (Ajzen 1991). This emphasize the need to involve the recreational fishers in the management of coastal areas.

This study through light upon the recreational fishermen profile, identifies the most affected fish species and quantifies the recreational fishing pressure in a representative

mediterranean coastal area. Combining three different sources of information (interviews, aerial survey and competition data) provided a more precise data set about the recreational fishing activity, resulting more useful for the management of the MPA. This is the first study in the Mediterranean that analysed the spatial and temporal recreational fishing effort distribution using aerial surveys and GIS methodology. The data obtained with these approaches led to reach coherent management strategies for the recreational fishing regulation inside MPAs, and could be applied in other MPAs of similar characteristics. The great potential impact of the recreational boat fishing on the resources had been demonstrated in Serra Gelada Marine Park. Therefore, based on the information and the data presented, some management strategies should be applied to regulate the recreational fishing activity in the study area in order to acchieve positive results from protection. An integrated management plan taking into account both professional and recreational activities is needed, because the declaration of a MPA without regulation does not guarantee the sustainability of the marine environment.

.

47

50

Publication

Luna-Pérez, B-, Valle, C., Sánchez-Lizaso, J.L. 2009. Benthic impacts of recreational

divers in a Mediterranean marine protected area. ICES Journal on Marine Science, 66:517-

523.

_________________________________________________________Chapter 3

Benthic impacts of recreational divers in a Mediterranean marine protected area

Abstract

The features of many MPAs have increased scuba diving tourism in these areas. Impacts caused by recreational scuba activity vary widely among different divers with differing underwater behaviour. We studied diver underwater behaviour, the effects on the natural environment, and the characteristics that may influence diver behaviour. In all, 181 recreational divers were followed, and contacts and the effects produced were recorded. Information on diver profile and dive features was recorded. Field sampling revealed that 175 of the divers observed (96.7%) made at least one contact with the seabed, with a mean contact of 41.20 ± 3.55 (mean ± s.e.) per diver per 10 min.

Flapping was the most frequent type of contact, and the main damage by this action was to raise of sediment. Contact with the seabed was greater for males than for females, inexperienced divers than for experienced divers, camera or lantern users than for non-users, and divers non-accompanied by a dive leader or who had not been briefed about avoiding seabed contact before undertaking a dive than for accompanied or briefed divers. A greater understanding of the causes of harmful behaviour may be useful for stricter management, reducing diving damage and assuring the sustainability of this activity in marine protected areas.

Keywords: diver damage, management of scuba diving, marine protected area, MPA, recreational impact.

1. Introduction

For decades, the creation of marine protected areas (MPAs) has been considered to be the only way to restore natural communities and to protect marine ecosystems (Milazzo et al., 2002). MPAs are now being established around the world rapidly (Ballantine, 1995), and “marine-based” tourism is growing even faster (Ribera, 1991; Boudouresque and Ribera, 1995; Davis and Tisdell, 1995). The aesthetic appeal of MPAs and the facilities they provide, together with the increased public awareness of nature, all contribute to creating massive tourism in MPAs (Ribera, 1991; Richez, 1991, 1992,

1993; Capellà et al., 1998; Badalamenti et al., 2000).

In the past 20 years, the number of visits to MPAs has increased globally (Dixon et al., 1993; Hawkins and Roberts, 1994; Kelleher et al., 1995), with an associated increase in the rates of participation in marine recreational activities, such as snorkelling, scuba diving, or boating (Tabata, 1989, 1992; Dignam, 1990; Marion and Rogers, 1994; Davis and Tisdell, 1995). New technology and the consequent safety improvements have greatly increased the number of recreational divers (Davis and Tisdell, 1995), as well as the extent of this activity around the world (Hawkins and Roberts, 1992).

51


BENIDORM

0º6’W0º10’W 0º2’W 0º2’E

38º32’N

38º36’N

1020

50

NORTH

ALTEA

Benidorm Island

Llosa reef

Mitjana Island

Mascarat

Portugal

France

MediterraneanSea

SpainPortugal

France

MediterraneanSea

Spain

Scuba diving may result in the deterioration of benthic communities, because divers can easily damage marine organisms through physical contact with their hands, body, equipment, and fins (Talge, 1992; Rouphael and Inglis, 1995, 1997; Tratalos and Austin, 2001; Zakai and Chadwick-Furman, 2002; Pulfrich etal., 2003; Uyarra and Côté, 2007). Although the damage produced by individuals is usually minor, there is some evidence that the cumulative effects of the disturbances can cause significant localized destruction of sensitive organisms (Garrabou et al., 1998; Hawkins et al., 1999; Plathong etal., 2000). There is a bigger problem when the diving activity focuses on MPAs. In some cases, the effects of a large number of divers in a few places in a marine reserve can be contrary to the main objectives of the creation of the MPA (Davis and Tisdell, 1995, 1996; Coma et al., 2004; Hawkins etal., 2005). However, some authors state that the impact of divers at a site may be influenced more by their experience and behaviour than by the number of people who frequent the site (Davis and Tisdell, 1995; Rouphael and Inglis, 2001; Barker and Roberts, 2004).

The relationship between diver behaviour and their impact on marine communities has been widely studied in coral reef areas (Caribbean, Red Sea, Australia), but more seldom in temperate systems. The main objectives of our study were to (i) characterize diver behaviour according to a diver profile, and (ii) evaluate and quantify their effects on the Mediterranean benthic community. We provide an

estimate of the damage rate of divers on the benthic communities and demonstrate the relationship between diver behaviour and profile, and the dive characteristics, and illustrate the most important damaging factors in order to propose educational tools to help reduce the negative effects.


2.1. Study area

The study was carried out at the Sierra Helada Marine Park (SHMP; 4920 ha), which is located between Altea and Benidorm (Alicante, Spain; Figure 1) and includes the Benidorm and Mediana Islands, the Galera and Olla reefs, the surrounding waters, and the adjacent mainland coast. The Posidoniaoceanica seabed in the Sierra Helada Marine Park seems to be in an optimum state of health, and Cystoseira seabed, rocky coralligenous, maërl, and cave habitats are also well represented.

Figure 1. Sampling area location.

52

_________________________________________________________Chapter 3

The physical and biological characteristics in and surrounding the SHMP are very important for marine recreation, and activities there have increased in recent years. The number of divers increased from 10 775 in 2005 to >24 000 in 2006 (Coselleria Territori i Habitatge), and there was strong seasonal and spatial concentration, with 13 336 divers in summer 2006 (55.4% of the annual total) visiting just three sites.

2.2. Sampling diver behaviour

In all, 181 scuba divers were followed at the most popular places in the park (Llosa Reef, Benidorm Island, and Mitjana Island) between July and September 2006. Divers were selected randomly from visiting groups and discreetly observed for 10 min each. The 10 min observation period was determined, from pilot studies, to be the most cost-effective for observing a large number of divers with the available resources (Rouphael and Inglis, 1997). The monitoring of divers started 10 min after their entry into the water, and after the divers adjusted their equipment and buoyancy.

On each dive, the number and the duration of contacts with the seabed were noted, as well as which part of the diver (hand, body, or knee) or equipment (tank, fins, camera, light, or octopus) was involved in the contact. The number of times a diver flapped fins, collected or handled organisms, fed organisms, or turned over rocks was also recorded during observation periods. For all actions, the consequences of this contact (raising

sediment, contact with fragile species, or removal of algae) were reported. Table 1. Factors influencing diver behaviour and assessment level.

Observers remained in the water behind their target divers, with visual contact (about 4–5 m away), so as not to influence diver behaviour. The observations were interrupted when the target diver modified his or her behaviour as a consequence of the

Diver factor Level Gender Male Female Age 1: <31 years old 2: 31–35 years 3: 36–40 years 4: 41–50 years 5: >50 years Diving qualification 1: Beginner 2: Moderately 3: Expert Years diving 1: <1 year 2: 1–2 years 3: 3–5 years 4: 6–10 years 5: >10 years Number of dives 1: <6 dives 2: 6–12 dives 3: 13–24 dives 4: 25–50 dives 5: 51–100 dives 6: 101–200 dives 7: >200 dives Perception of damage Yes No No answer Dive factors Level Depth 1: <12 m 2: 12–20 m 3: >20 m Leader Yes or no Briefing Yes or no Camera Yes or no Lantern Yes or no

53


presence of the observers, or when the target divers expressed curiosity about the observer’s work. We determined the influence of diver characteristics on their underwater behaviour. After each dive, target divers were asked about their diving experience, gender, age, and perception of the damage of scuba diving. Finally, we collected information on 11 factors that may influence diver behaviour underwater.

The factors that may influence diver behaviour were assessed separately (Table 1): (1) gender (male, female); (2) age (five age groups); (3) diving qualification (three categories); (4) total number of dives completed by diver (seven levels); (5) number of years diving (five levels); (6) whether or not a briefing was given; (7) the presence of the centre diving guide during a dive; (8) the diving depth (three levels of depth); (9) use or not of a camera; (10) use of a lantern for lighting; and (11) perception of the diving damage on the marine environment.

2.3. Statistical analyses

Differences in the mean frequency of contacts for each main factor were examined by one-way analysis of variance (ANOVA). For each factor, we evaluated three independent variables: the number of contacts with any part of the equipment and the body, and the total number of flapping events.

Prior to ANOVA, heterogeneity of variance was tested with Cochran’s C-test. Data were �x + 1 transformed if variances were heterogeneous, followed by ln(x + 1) transformation if necessary.

Where variance remained heterogeneous, non-transformed data were analysed, because ANOVA is robust to heterogeneity of variance, particularly for large, balanced experiments (Underwood, 1997). In those cases, we used a more strict significance level (p < 0.01).

3. Results

3.1. Divers profile

Of the 181 divers followed and interviewed immediately after diving, there were more men (72.9%) than women, and the ages ranged from 18 to 69 years. The mean age group for both genders was the same, 30–40 years. The percentages of men and women sampled within each age category were similar.

Most divers had beginner (44.7%) or medium-level dive qualification (42.0%); just 13.3% had expert level. Of the total number of divers observed, 20 (11%) were photographers and 44 (24.3%) used a lantern during their dive.

3.2. Effects of diver behaviour

Of the 181 divers observed, 175 (96.7%) made at least one contact with the seabed, with a mean of 41.20 ± 3.55 (mean ± s.e.) contacts per diver per 10 min. Flapping was the most frequent type of contact. On average, each diver made 32.56 ± 3.24 flapping contacts per 10 min (n = 5894 contacts), and the main observed consequences or damage were the raising of sediment (58.3%) and the removal of algae (38.1%).

54

_________________________________________________________Chapter 3

The mean number of contacts with any part of the equipment and any part of the diver were similar at 4.30 ± 0.35 and 4.34 ± 0.44, respectively. The type of contact varied with equipment part. Contact with fins accounted for the greatest proportion (78.5% of contacts; n = 778), with a mean per dive of 3.37 ± 0.30 and a duration of 8.90 ± 0.87 s per contact with fins. The most common consequence was the trampling of organisms (27.6% of 611 contacts), resuspension of sediment (25.2%), and the removal of algae (20.0%). The second element causing damage, with a total of 127 contacts, was alternative air sources (“octopus”) or manometer, with a mean per dive of 0.70 ± 0.12, and the most frequent effect of this activity was the resuspension of sediment (30.0%) and the removal of algae (26.8%). Contact with other elements, tank, camera, or light, were less common during the dives, but tank contact time was high (11.91 ± 2.55 s per contact).

For contacts with any part of the body (n = 786), the most frequent was contact with the hands (3.0 ± 0.32 per diver per 10 min, with a duration of 9.38 ± 0.85 s). The total number of contacts with hands was 539, and it usually involved trampling of organisms (43.2%), contact with fragile organisms (22.8%), and the raising of sediment (21.7%). The divers made 1.20 ± 0.19 contacts per 10 min with their knees (n = 218 in all), with a mean duration of 14.47 ± 2.10 s. The most common effects were trampling the organisms and removing the algae (51.0% and 5.6%, respectively). Although contact with the whole body when lying on the bottom was infrequent (0.16 ± 0.04 per dive of 10

min), the contact was longer (23.25 ± 6.87 s per contact). Trampling of organisms (34.5%) and the resuspension of sediment (31.2%) were the most common results of whole body contact (n = 29).

While divers are taking photos, they usually rest another part of their body or their equipment on the seabed, so a mean of 5.14 ± 0.90 contacts per 10 min was recorded, with a mean duration of 31.84 ± 25.04 s. This action made contact with fragile organisms (30.5%), and 19.4% caused trampling and removal of algae.

Divers who carried lanterns during their dive had similar behaviour to that observed for photographers. The divers with lanterns made 2.69 ± 0.66 contacts per 10 min, with an average time of 23.98 ± 11.20 s. Contact was made with fragile organisms in 54.3% of the total number of lantern contacts, and 17.1% caused trampling of organisms or raising of sediment, in the same proportion.

During the dives sampled, handling for organism recognition was observed 17 times, with an average duration of 30.72 ± 9.50 s. Octopuses were the organisms most frequently disturbed by divers, followed by snails, sea stars, and crabs. Collecting of organisms by divers was observed five times, 2.3% of the total number of divers. The divers collected small snails and one Myriaporatruncata colony. Stone-turning was only observed seven times, and was caused by inadvertent use of the fins.

55


Figure 2. Mean number of contacts by diver gender. Error bars indicate 1 s.e.

3.3. Diver behaviour according to profile

There was a significant gender difference (Figure 2) in the number of damaging contacts on the seabed. Female divers caused proportionally fewer contacts than males, but these differences were significant only for flapping (p < 0.001) and contacts with any part of the body (p < 0.05). The age of divers did not have a significant effect on behaviour, and no trend was observed between age and the number of contacts (Figure 3a). Diver qualification did not have a significant effect on the number of contacts (Figure 3b), but experience (number of completed dives; Figure 3c) and years diving (Figure 3d) were significant. The most experienced divers, who had more dives completed or more years of diving, made statistically fewer contacts. These differences in the two factors were significant for flapping (p< 0.001) and for contact with any part of the diving equipment (p < 0.001).

The divers who did not perceive any ecological damage by scuba diving (26%) or who did not answer the

question (9.4%) made more contacts with the substratum than divers who thought that scuba could cause ecological damage (Figure 3e). Flapping was the best variable for detecting these differences (p < 0.05).

3.4. Diver behaviour according to dive features

When the diving profile was analysed, the depth recorded did not show any significant differences in diver behaviour, although contacts by any part of the body were more prevalent in deeper dives, as were flapping and contact by equipment (Figure 3f).

The effect of a previous briefing by a senior member of the diving club significantly reduced (p < 0.01) the diver contact rates for flapping and contact with any part of the body (Figure 3g). Although no significant differences were found, contacts by any part of the diving equipment were also fewer. Similarly, when the dive was guided by a diving leader, flapping and the number of contacts with any part of the body were less (Figure 3h), although only the decrease in contacts with any part of the body was significant (p < 0.01).

Underwater photographers made contact with the seabed more frequently than non-camera users (Figure 3i), mainly through flapping (p < 0.05) and contact by parts of the equipment (p < 0.01). The divers carrying lanterns (Figure 3j), were normally closer to the seabed, looking for holes, fissures, etc., and made more contacts than non- lantern users (p < 0.001).

Type of contact

Flappings

Diver contacts

Equipment contacts

20

40

60

Male Female

56

_________________________________________________________Chapter 3

Figure 3. Mean number of contacts according to diver profile and dive features. Error bars indicate 1 s.e..

a) Age

<30 30-50 36-40 41-50 >50

10

20

30

b) Diving qualification

Beginner Moderate Expert

10

20

30

40

50

d) Diving years

1 2 3-5 6-10 >10

20

40

60

80

100

c) Number of dives

<6 6-1213-24

25-5051-100

100-200 >200

20

40

60

80

e) Perception of damage

Yes Not NA

20

40

60

80

Flappings Diver contacts Equipment contacts

g) Briefing

Yes Not

10

20

30

40

f) Depth

<12 12-20 >20

10

20

30

40

50

h) Leader

Yes Not

10

20

30

40

i) Camera

Yes Not

10

20

30

40

j) Lantern

Yes Not

5

10

15

20

25



57


4. Discussion

Identifying the different factors that describe diver behaviour and their environmental effects may help managers to develop more effective training procedures, pre-dive briefings, site regulations, etc., in order to prevent or reduce the incidence of damaging behaviours (Rouphael and Inglis, 2001). For this reason, we evaluated the relationships between factors that could influence underwater behaviour, and showed the existence of several intrinsic diver and dive factors that may influence the effects of scuba diving. Contact frequency with the seabed is strongly influenced by diver profile and immersion characteristics.

By following divers, we showed that 96.7% of them made some contact during the 10 min immersion, causing potentially serious damage to the environment. This rate is consistent with other works on scuba-diver impacts (Rouphael and Inglis, 2001; Uyarra and Côté, 2007), most divers (90%, 70%, and 75%, respectively) making one or more interactions with the substratum. We estimate that in a 45 min immersion, each recreational diver may make 100 interactions resulting in the raising of sediment, more than 60 removing algae, eight contacts with fragile organisms, and 14 contacts that result in the trampling of organisms. The other observed effects would be less frequent.

Most contacts were caused by flapping and contact with fins, confirming similar results from the Red Sea (Prior

et al., 1995; Zakai and Chadwick-Furman, 2002), Australia (Roberts and Harriot, 1994; Harriot et al., 1997; Rouphael and Inglis, 2001), and the Caribbean (Barker and Roberts, 2004; Uyarra and Côté, 2007) which attribute most diver damage to the the effect of fins. The hands were the part of body that made most impacts, as also observed for divers in the Red Sea (Zakai and Chadwick-Furman, 2002) and the Caribbean (Barker and Roberts, 2004; Uyarra and Côté, 2007). Most contacts appeared to be unintentional and caused by poor swimming technique, incorrect weighting, and lack of warning, factors that, in general, indicate a poor diving proficiency.

There were gender differences in the contact frequency for flapping and contact with any part of the body, women causing fewer impacts than men. In general, a male diver is less cautious in his underwater behaviour, tending to be more adventurous and more likely to take more risks than women, a relationship also shown by other studies of environmental attitudes and the behaviour of male and female divers (Hudgens and Fatkin, 1985; Vredenburgh and Cohen, 1993; Rouphael and Inglis, 2001). Another explanation for this result is that men are more likely to ignore pre-dive instructions on safety and environmental behaviour advice than women, having a more independent attitude, an observation made also be others (Vredenburgh and Cohen, 1993).

The effect of diver experience on the number of impacts has been documented in some areas (Roberts and

58

_________________________________________________________Chapter 3

Harriot, 1994), but not in others (Harriot et al., 1997; Rouphael and Inglis, 2001), but these differences could be because of differences in the definitions. In our study, diving experience was measured through three variables: level of the diving certificate, the total number of dives, and the number of years diving. Both total number of dives and the years of diving were associated with environmental impact, less damage being caused by the more experienced divers. The diving certificate level did not show this association, so it does not appear to reflect diving experience. A diver can obtain a higher certification by taking an advanced course, but this does not mean that the person becomes more proficient. Divers can even become diving instructors with fewer than 100 immersions. Dive-training certificates are lifetime qualifications and do not require periodic renewal or dive proficiency testing. Therefore, diver training level may not be a good indicator in determining current diving skills, and this factor is not sufficient to determine whether a diver is qualified for diving at a site. This topic needs to be considered when adopting management strategies in dive areas.

Briefings before the dive and underwater intervention by a dive leader were highly effective at reducing the average impact of flapping and contact by any part of the body, as found in other studies (Medio et al., 1997; Barker and Roberts, 2004; Uyarra and Côte, 2007). These differences in dive behaviour were more obvious for intentional contacts by any part of the body, confirming that deliberate contacts may be reduced by the

implementation of simple measures by the diving centres. Attending a briefing emphasizing the importance of buoyancy control and careful action (educational tools) increases the environmental awareness of recreational divers and might reduce diver damage at dive sites. Moreover, the use of dive leaders during dives is a clearly effective tool in minimizing scuba divers’ physical impact on their environment, because dive leaders can take measures when they see divers behaving inappropriately. For this reason, smaller dive groups tend to be better, dive leaders being able to supervise all members of the group adequately (Barker and Roberts, 2004). In any case, smaller groups are preferred by the divers themselves.

Control of instruments used at the most fragile sites by divers, such as cameras or lanterns, may be a good measure for controlling damage by scuba diving, because carrying any element causes divers to have a greater interaction with the environment than when divers carry nothing. This finding, in terms of photographers, has been observed in other studies (Medio et al., 1997; Rouphael and Inglis, 2001; Barker and Roberts, 2004; Uyarra and Côté, 2007), but the effect of carrying lanterns has not been evaluated before. The photographers being observed tended to adopt the most comfortable and best position to avoid movement and to obtain better images. They then cause damage by anchoring themselves at the bottom, using their knees, fins, elbows, etc. When carrying a lantern, divers exhibit a particular behaviour, looking for small holes, cracks, caves, or

59


animals to illuminate, and they often disregard their buoyancy or fail to keep their equipment or body off the seabed.

All these impacts have consequences for the benthic community at the most popular dive sites of the SHMP. The area hosts an advanced coralligenous community composed of many sessile, filter-feeding, long-lived organisms with fragile skeletons and slow rates of growth (Laborel, 1961; Ros et al., 1985). The risk of long-term degradation should be determined by the impact rate and the speed with which it is repaired (Rouphael and Inglis, 1997). The problem is that organisms living in Mediterranean coralligenous communities are not adapted to severe disturbance, and their recovery after moderate pressure might be difficult (Garrabou et al., 1998). The sustainability of diving activity at particular sites depends on both the number of divers accessing the sites and the capacity of the ecosystem to regenerate and recover from any damage incurred (Harriot et al., 1997). Monitoring programmes need to be established in order to detect environmental changes at dive sites before diving impact levels become critical and, perhaps, irreparable.

4.1. Management of recreational scuba diving

Divers who were interviewed and saw the negative effects of diving as ecologically critical behaved more carefully, suggesting that diving environmental awareness programmes could be considered. In this sense, managers need to make an effort to

teach divers the environmental value and the fragility of different species, as well as show the potential damage of diving activity and how to minimize the negative impact of scuba diving. Diver education programmes ought to include the environmental effects of diving. To improve diver skill, a management agency could require that diving operators attend update courses, pre-dive briefings, and employ a dive leader, to reduce diver impact. Of course, this should also apply to photographers and lantern-carriers.

Diving quotas may also be appropriate in sensitive areas, because the impact at a site is influenced by the number of visitors (Barker and Roberts, 2004). The idea that there is a limit – a “carrying capacity” in terms of human usage – needs to be embraced to ensure that natural resources are not destroyed (Salm et al., 2000). In recent decades, there have been attempts to estimate the carrying capacity of popular dive sites (Davis and Tisdell, 1995), but the results of the estimates vary extensively between different sites around the world, from 5000 to 50 000 divers per site per year (Hawkins and Roberts, 1992; Dixon et al., 1993; Chadwick-Furman, 1997; Harriot et al., 1997; Hawkins et al., 1999; Schleyer and Tomalin, 2000; Zakai and Chadwick-Furman, 2002). The disagreement between the various studies shows that carrying capacity can depend on a combination of many factors, including (i) biological characteristics of the dive site and the presence of vulnerable organisms (Riegl and Cook, 1995; Chadwick-Furman, 1997; Harriot et al., 1997; Rouphael and Inglis, 1997, 2001;

60

_________________________________________________________Chapter 3

Schleyer and Tomalin, 2000; Walters and Samways, 2001), (ii) the underwater activities pursued (Rouphael and Inglis, 1997, 2001; Uyarra and Côté, 2007), (iii) the level of environmental awareness of the divers and their experience or technical competence (Davis and Tisdell, 1995; Medio et al., 1997; Rouphael and Inglis, 1997), (iv) the physical conditions present during a dive, such as wave or current motion (Harriot et al., 1997; Rouphael and Inglis, 1997), (v) the presence of other anthropogenic stressors, such as boat-anchoring (Davis, 1977; Halas, 1985) or pollution (Hawkins and Roberts, 1997), and (vi) the frequency of large-scale natural accidents that may cause deleterious effects, such as events with massive

mortality (Nagelkerken et al. 1997; Cerrano et al., 2000; Pérez et al., 2000). Carrying capacity needs to be considered as an elastic factor that should be adapted to communities at each particular dive site and periodically reviewed in order to enhance its effect. Managers need to note that, although there may be too much emphasis on limiting visitor numbers, other management strategies (e.g. educational programmes) could be used to greater effect (Rouphael and Hanafy, 2007). Our results suggest that there is a need for diving activity managers at the most popular dive sites to maintain and conserve the aesthetic appeal and biological characteristics of their site, to achieve sustainable scuba diving attraction.

61

Publication

Luna-Pérez, B.; Valle, C.; Vega-Fernández, T.; Sánchez-Lizaso, J.L; Ramos-Esplá, A.A. 2010. Halocynthia papillosa (Linnaeus, 1767) as an indicator of SCUBA diving impact. Ecological Indicators, 10: 1017-1024.

64

_________________________________________________________Chapter 4

Halocynthia papillosa (Linnaeus, 1767) as an indicator of SCUBA diving impact

Abstract

The solitary ascidian Halocynthia papillosa (Linnaeus, 1767) is proposed as a good indicator of the deleterious effect of SCUBA diving on the Mediterranean coralligenous communities. A comparative survey of H. papillosa populations at frequented and unfrequented localities was carried out over a two year period (during 2006 and 2007), before and after a peak diving season in the Sierra Helada Marine Park (SW Western Mediterranean Sea). We observed bigger and more abundant individuals of H. papillosa at undived sites than at frequented dived sites during the period of study. Furthermore, individuals of H. papillosa in the most frequented localities occupied more cryptic positions than in the undived localities. H. papillosa was shown to be very sensitive to the adverse effects of SCUBA diving. This species could represent a reliable bioindicator of diving activity and as such constitute a useful tool for the quick and easy monitoring of impact on coralligenous communities before this damage becomes unmitigatable.

Keywords: SCUBA diving impact; bioindicators; tunicates; Halocynthia papillosa; coralligenous; repeated measures design.

1. Introduction

In the Mediterranean Sea, the deep littoral system is dominated by calcareous organisms, sessile filter-feeding animals and dead skeletons, generally referred to as “coralligenous” (Ballesteros, 2006). The coralligenous community is an example of a highly diversified and highly structured community of suspension feeders, which is the richest in terms of both species and organization in the Mediterranean benthos (Ros et al.,1985; Ballesteros and Ros, 1989). This high diversity of organisms and colourfulness of these habitats make them more attractive for divers (Lloret et al., 2006) and attracts ever increasing numbers of divers to these benthic habitats.

Recreational SCUBA diving is an important and growing component of the international tourism market (Tabata, 1992; Davis and Tisdell, 1995). Usually, this activity typically occurs in marine locations with impressive underwater scenery, high biological diversity (Tabata, 1989) or special value which in many cases concur with the status of protected areas (Davis and Tisdell, 1995). However, SCUBA diving may result in the deterioration of marine parks because divers can easily damage marine organisms directly (e.g. by physical contact with any part their body or equipment) and indirectly (e.g. by raising sediment) (Rouphael and Inglis, 1997; Tratalos and Austin, 2001; Zakai and Chadwick-Furman, 2002; Pulfrich et al., 2003; Luna et al., 2009).

65


Some studies of diving activity monitoring have shown that raising sediment was the most frequently effect produced by divers, followed by contact with fragile organisms such as gorgonians, corals and bryozoans (Garrabou et al., 1998; Zakai and Chadwick-Furnman et al., 2002; Barker and Roberts, 2004; Luna et al., 2009).

Among others, there are three major mechanisms of sediment raising that may directly affect rocky coastal organisms: burial or smothering, scour or abrasion by moving sediments and changes in the physical characteristics of the stability of hard substrate that can result in a loss of suitable settlement habitat (Airoldi, 1998). Silting up clearly influences several species of sessile invertebrates; among these the ascidian growth and survival of ascidian species are adversely affected by excessive sediment deposition which causes burial and clogging of the siphons and the branchial wall. Sediment also produces abrasion effects and hinders the settlement processes (Bakus, 1968; Gabriele et al., 1999).

On the other hand, divers can easily damage marine organisms through physical contact with their hands, body, equipment and fins (Rouphael and Inglis, 1997; Tratalos and Austin, 2001; Zakai and Chadwick-Furman, 2002; Pulfrich et al., 2003; Uyarra and Côté, 2007) causing breakage, fragmentation or abrasion effect on tissue (Hawkins and Roberts, 1992; Garrabou et al.,1998; Coma et al., 2004).

The combination of two effects, resuspension of sediments and physical

contact, is the main stressor agent which the benthic communities are subjected to by diving activity. However, the large number of species and the high level of species diversity in the coralligenous community hinder the assessment of diving impact on the community as a whole; therefore, the selection of a single indicator species would be advantageous (Sala et al., 1996).

Ascidians, as sessile suspension feeders, are nutritionally dependent on the surrounding water which provides them with resources. They are considered as non-selective filter feeders (Jørgensen, 1949, 1955; Stuart and Klumpp, 1984) able to capture particles up to 0.1 μm (Monniot, 1972), but normally from 0.5 to 100 μm (Fiala-Médioni, 1987) with high retention efficiency for particles larger than 0.6 μm (Fiala-Médioni, 1978; Stuart and Klumpp, 1984). In spite of its important structural role in benthic communities, as much in temperate (Pérès, 1967; Ramos-Esplá, 1991), tropical (Goodbody, 1995) and polar (Gili et al., 2001) waters, little effort has been made to acknowledge this role in ascidians. Moreover, very few species have been subject to conservation measures. In this sense, and in the Mediterranean Sea, only the Microcosmus spp. and H. papillosa have been proposed as vulnerable species (Pérès, 1991; Ramos-Esplá et al., 2008).

In this work, we hypothesized that an increase of resuspended sediment produced by SCUBA divers will have a negative effect on H. papillosa’s energy acquisition and development. In addition, the abrasive effect of the

66

_________________________________________________________Chapter 4

sediment on the tissue of ascidians as well as direct physical contact by divers may produce additive negative effects. This species could be a useful indicator for assessing and monitoring the diving effect on Mediterranean coralligenous community. The aim of this study was to assess the suitability of H. papillosa as a diving impact indicator and its potential role as an effective tool in the detection of disturbance to the ecological system. To achieve this main objective, the secondary objectives have been: (1) to quantify the abundance of H. papillosa between sites frequented and unfrequented by divers; (2) to study its size distribution between these sites; and (3) to study its degree of exposure on the substrata.


2.1. Study area

The present study was carried out during 2006 and 2007 in the Sierra Helada Natural Park, on the South-eastern Mediterranean Spanish coast (Fig. 1). This Natural Park occupies a total area of 5,665 ha, with 4,920 ha of marine zone, and comprises the Benidorm and Mitjana Islands, the surrounding waters and the adjacent mainland coast. The number of annual divers has increased from 10,775 in 2005 to more than 24,000 in 2006, mainly concentred from June to September. At present, SCUBA diving is allowed throughout the park and there is no control or restrictions on this activity.

Figure 1. Study area with indication of six sampled localities. The symbols indicate: � unfrequented localities and � frequented localities.

2.2. Studied species

H. papillosa is an endemic Mediterranean species (Pérès, 1964) which also occurs in the Eastern Atlantic within an area which extends from the Canary Islands (Lozano, 1972) to the Berlengas Islands in Portugal (Ramos-Esplá et al., 2008). It represents one of the most common solitary species on the rocky littoral of the South-western Mediterranean (Ramos-Esplá, 1991). The individuals show bottle form and measure up to 15 cm in height (Calvín-Calvo, 2000). H. papillosa is a hermaphroditic species that reproduces once per year in the late summer and achieves a peak of gonadal development from March to September for ovaries and from July to September for testes (Becerro and Turon, 1992). In

BENIDORM

0º6’W0º10’W 0º2’W 0º2’E

38º32’N

38º36’N

1020

50

NORTH

ALTEA

� Benidorm Island

� Llosa reef

� Mitjana Island

Mascarat

Portugal

France

MediterraneanSea

SpainPortugal

France

MediterraneanSea

Spain

Lighthouse �

� Cala Mina

� Punta Albir

67


Table 1. Physical and chemical variables which characterized the sampling localities.

a: Consellería Territori I Habitatge, 2005 b: MOPU, 1987 c: Own data recorded during sampling

contrast to other benthic filter-feeding species, H. papillosa does not exhibit aestivation (summer dormancy). In fact, it is in summer when its level of activity and growth rates are highest (Ribes et al., 1998).

2.3. Sampling procedures

The survey was conduced by SCUBA before and after of the peak diving activity (June and October) during a two year period (2006 and 2007). Six sampling localities were established inside the marine park limit: three impacted localities where diving activity is concentrated (Benidorm

Island, Llosa Reef and Mitjana Island), the three localities in unfrequented diving areas (Cala Mina, Punta Albir and Lighthouse) (Fig. 1). The sampling places were chosen in order to be as similar as possible among them, taking into account both environmental conditions and human impacts, except for the frequency of diving visitors. Two localities (Mitjana and Benidorm) are remote and uninhabited islands. Other localities (Llosa Reef, Cala Mina, Lighthouse, Punta Albir) are inaccessible because their emplacement, on cliffs or reefs locations. The localities had similar community, depth (19 to 25 meters), orientation, slope,

Variables Benidorm Island Lighthouse Llosa Reef Mitjana

Island Punta Albir Cala Mina

Most frequented wind direction a NE and E NE and E NE and E NE and E NE and E NE and E

Most frequented water current direction a SE and NW SE and NW SE and NW SE and NW SE and NW SE and NW

Water Maximum annual Temperature (ºC) a 26.1 26.0 26.0 26.0 26.0 26.1

Water Minimum annual of temperature (ºC) a 13.5 13.3 13.5 13.5 13.4 13.5

Salinity (‰) c 38.0 38.0 38.0 38.0 38.0 38.0 Transparency (m) a > 5 > 5 > 5 > 5 > 5 > 5 O2 dissoult (mg/L) a 8.4-9.4 8.4-9.4 8.4-9.4 8.4-9.4 8.4-9.4 8.4-9.4 Nitrites( �g/L) a 10 - 2 10 - 2 10 - 2 10 - 2 10 – 2 10 - 2 Nitrates (�g/L) a 80-40 79-36 80-40 80-40 72-40 80-40 Fosfates ( �g/L) a < 0.02 < 0.02 < 0.02 < 0.02 < 0.02 < 0.02 Seston (�g/L) a < 5 < 5 < 5 < 5 < 5 < 5 Survey stations deep (m) c 19.5-23 21-24 21-25 22-24 21-23.5 20-21 Slope (%) b 3 to 5 3 to 10 3 to 10 3 to 5 3 to 5 3 to 5 Mud of sediment (%) b < 20 < 20 < 20 < 20 < 20 < 20 Sands of sediment (%) b > 75 > 75 > 75 > 75 > 75 > 75 Graves of sediment (%) b � 1.1 � 1.1 � 1.1 � 1.1 � 1.1 � 1.1 Organic Matter of sediment (%) b 9 - 11 9 - 11 9 - 11 9 - 11 9 – 11 9 - 11

68

_________________________________________________________Chapter 4

and type of substrate, all determining factors in the structure of benthic communities (Ros et al., 1985) (Table 1).

We used permanent plots to remove the expected high local variability between randomly sampling units for effect of the mosaic distribution of H. papillosa (Ramos-Esplá, 1991). This high variation might obscure effects of our treatments thus complicating the detection of diving impacts. For this reason, we took a permanent biological unit to examine and monitor the ascidians populations and also to study the processes and mechanisms that determine the response of damaged populations (Quinn and Keough, 2002). For this reason, at each locality six metal screws marked the centre of six permanent circular plots of 4 meters radius (covering 50.26 m2).

During the survey, all relaxed H. papillosa individuals within the plots were counted and their height measured with a plastic rule to the nearest 0.1 cm. The height data was divided into five size classes, according to the maximum size recorded in bibliography (Calvín-Calvo, 2000): Size 1: 0.1 to 3 cm; Size 2: 3.1 to 6 cm; Size 3: 6.1 to 9; Size 4: 9.1 to 12 and Size 5: 12.1 to 15 cm. Moreover, for each individual, its position in the circular plot and the degree of exposure was recorded. This degree of exposure was assigned in five categories according to their position: (1) as epibionts, (2) on convex surfaces, (3) on flat surfaces, (4) in concavities, and (5) under overhang (Fig. 2). The periodical surveys of the permanent plots allowed the monitoring of 3

variables of the population at Sierra Helada Marine Park: degree of exposure, size class abundance and total abundance of individuals.

Figure 2. Different degree of exposure of individuals : (1) as epibionts, (2) on convex surfaces, (3) on flat surfaces, (4) in concavities and (5) under overhang.

2.4 Statistical design and analysis

The use of permanent plots assumes that the same sampling units are retested over time and that data are not independent. Simple analysis of variance can not be applied to this method of data collection since it does not obtain an independent measure of residual variation and does not fulfil the assumption of no interaction among treatments (Kingsford and Battershill, 1998). Repeated measures (RM) ANOVA analysis has been proposed to correct the non-independence or autocorrelation in the data (Winer, et al., 1991; Keough and Quinn, 1998; Kingsford and Battershill, 1998). We used this method for analyzing our data in order to identify changes in the community for diving effect over time.

Four experimental factors were considered, and assigned nomenclature according to common practice in RM

69


experiments. Diving intensity was the first between-subjects, fixed factor, with two levels (unfrequented vs. frequented); and locality was a second between-subjects, random factor, with three levels (locality 1, 2 and 3). Plot was the subject factor on which, random and nested within locality, with six levels (the replicates or experimental units 1 to 6). There were two blocking factors defining fixed periods of time in this experiment: year was the first

within-subject, fixed factor, with two levels (1 and 2); and Before vs. After the diving season was the second between-subjects, fixed factor, with two levels (before and after). We obtained a total of 144 replicates. Response variables were the total abundance of H.papillosa individuals, the percentage of size-classes, and the percentage of individuals for each degree of exposure.

For all analyses we examined primarily the assumption of normality, by examining residuals with the use of probability plots. Percentage data (degree of exposure and size class), were arc-sin square root transformed to achieve normality (Winer et al., 1991). When repeated-measures analyses were used, the more conservative Greenhouse-Geiser and Huynh-Feldt correct F test were also examined because this correction provides some protection against violations of assumptions of compound symmetry (Winer et al., 1991; Yandell, 1997). These tests did not produce results markedly different from the uncorrected ones, and evidence of strong violation of these assumptions were not detected since epsilon value was close to one, so only the standard F tests are presented here.

Table 2. Summary of results of repeated measures analyses of density and percentage of each size class. Tabuled values are F ratios of terms of interest. Significant values are indicated in boldface and level of significance as: *: p < 0.05; **: p < 0.01 and ***: p < 0.001. d.f.: Degree of freedom.

Source of variation D.f Density Size 1 Size 2 Size 3 Size 4 Size 5

A) Between plots effects Unfrequented vs. Frequented (D) 1 22.91*** 3.34 4.62* 9.38** 12.13** 7.15* Locality(D) 4 1.32 4.44** 3.26* 2.23 2.92* 3.70* B) Whithin plots Year 1 5.72* 2.02 9.08** 8.02** 0.87 1.57 Year x D 1 9.43** 1.51 0.95 0.67 3.78 3.88 Year x Locality(D) 4 3.65* 0.72 3.26* 1.61 1.07 0.53 Before vs. After (T) 1 3.39 1.38 0.38 1.01 0.05 0.03 T x D 1 16.07*** 0.19 3.59 1.19 0.62 2.20 T x Locality (D) 4 0.93 0.78 0.21 0.21 0.34 0.56 Year x T 1 0.45 2.56 0.54 0.26 0.76 0.36 Year x T x D 1 1.23 0.37 0.00 0.01 0.01 0.11 Year x T x Locality (D) 4 2.36 0.37 1.23 0.77 0.55 0.83 Transformation Square root Arc-sin Arc-sin Arc-sin Arc-sin Arc-sin

70

_________________________________________________________Chapter 4

3. Results

3.1. Effect of diving on the density

A total of 109 individuals of H. papillosa were found and measured at the beginning of the study at the permanents plots. Changes in density were observed in the sampling time since it was significantly lower (p < 0.001, Table 2) at frequented diving sites than at unfrequented sites throughout the whole study period (Fig. 3). Progressive increment in the density was observed at undived sites, ranging between 4.44 ± 0.85 (mean ± SE) individuals/replicate in first time sampling at first year and 7.05 ± 1.31 individuals/replicate after the peak diving season of the second year.

Repeated measure contrast, before and after peak diving season, was statistically significant for the frequented and unfrequented diving sites interaction (Table 2). The same pattern in density was observed at frequented diving sites, founding values significantly lower between before and

after the peak diving season (post-hoc Fisher LSD test; Before/After p < 0.01)

Density of ascidians was higher in the second sampling year than the first one, and when making a comparison between years, it increased at less frequented sites after the peak diving season (Fig. 3).

Figure 3. Mean density per replicate of individuals throughout study period, before and after of diving season during two years. Bars indicate the standard error and the symbols � and � frequented and unfrequented localities, respectively.

.

Table 3. Mean density per replicate of H.papillosa individuals at undived and dived localities, before and after diving season during two sampling years.

Control Impact

Before 1 After 1 Before 2 After 2 Before 1 After 1 Before 2 After 2 Size 1 0.17 ± 0.09 0.22 ± 0.10 0.89 ± 0.26 0.44 ± 0.19 0.06 ± 0.05 0.06 ± 0.05 0.28 ± 0.14 0.11 ± 0.08Size 2 1.67 ± 0.50 1.72 ± 0.49 1.17 ± 0.38 1.39 ± 0.46 0.67 ± 0.20 0.39 ± 0.17 0.78 ± 0.27 0.61 ± 0.21Size 3 1.83 ± 0.37 1.67 ± 0.40 1.83 ± 0.39 2.00 ± 0.39 0.72 ± 0.25 0.72 ± 0.25 0.61 ± 0.19 0.39 ± 0.17Size 4 0.61 ± 0.26 0.63 ± 0.24 1.56 ± 0.34 1.72 ± 0.41 0.11 ± 0.05 0.17 ± 0.09 0.22 ± 0.13 0.16 ± 0.12Size 5 0.17 ± 0.12 0.28 ± 0.20 1.22 ± 0.52 1.50 ± 0.57 0.06 ± 0.05 0.00 0.00 0.00

PeriodBefore 1 After 1 Before 2 After 2

Mea

n de

nsity

2

4

6

8

71


3.2. Changes in the structure of size class population

When comparing size distribution of ascidians population in the most frequented diving sites with less frequented sites (Fig.4), a significantly different pattern was observed for all sizes, except for the size class 1 (p < 0.01, Table 2). Population structure was characterized for individuals of size classes 2 and 3 in all sites but a lower proportion of the smallest size and an almost absence of individuals of the greatest was only recorded at the most frequented sites.

The detection of new individuals in permanent survey plots caused temporal change in the proportion of individuals of class 1 size (Table 3), which indicated newly settled individuals in the assemblage, thus showing lower recruitment at most frequented dive sites as compared with less frequented sites (Fig. 4).

Despite the observed changes in the structure of the population between before and after peak diving season, repeated measure analysis did not reveal any differences between the two times for any sites. High variability was also detected between years for size class 2 and 3 (Table 2), and in this case, repeated measure for year factor showed a significant change for size class 3 and the interaction year x locality for size class 2.

3.3 Changes in the degree of exposure

When analyzing the proportion of H. papillosa individuals at each degree of

exposure, differences between more and less dived sites over the study period were detected (Fig. 5). A higher proportion of individuals at more exposed levels (Levels 1, 2 or 3) was observed at unfrequented diving localities, although differences in this factor were only significant for level 2 individuals (Table 4). The presence of epibionts individuals (level 1) found at unfrequented localities was anecdotic because only one individual was found at one single locality, during the last sampling period. In contrast, more frequented diving localities showed a greater proportion of individuals of more cryptic locations (Level 5) (Table 4).

Repeated measure factor did not show any differences for levels of exposure over time, with the exception of the most cryptic location (Level 5), for which the proportion of individuals between before and after peak season were significantly different between times at the frequented diving sites (post hoc Fisher LSD test, p = 0.029).

Distribution of exposure levels of ascidians did not exhibit great fluctuations in unfrequented diving sites over sampling time. However, the repeated measure analyses for year (or some interaction of this factor), showed significant differences between the two years for all degrees of exposure (Table 4). A high spatial heterogeneity was observed in the proportion of individuals on degrees of exposure 1, 2 and 5, since analysis of variances detected significant differences between localities.

72

_________________________________________________________Chapter 4

Figure 4. Mean percentage of individuals for each size throughout study period, before and after of diving season during two years. Bars indicate the standard error and the symbols � and � frequented and unfrequented localities, respectively.

Figure 5. Mean percentage of individuals for each degree of exposure throughout study period, before and after of diving season during two years. Bars indicate the standard error and the symbols � and � frequented and unfrequented localities, respectively.

P e r io dB e fo re 1 A fte r 1 B e fo re 2 A f te r 2

Mea

n pe

rcen

tage

Siz

e 5

0

5

1 0

1 5

2 0


Mea

n pe

rcen

tage

Siz

e 4

5

1 0

1 5

2 0

2 5

3 0


Mea

n pe

rcen

tage

Siz

e 3

1 0

2 0

3 0

4 0

5 0


Mea

n pe

rcen

tage

Siz

e 2

1 0

2 0

3 0

4 0


Mea

n pe

rcen

tage

Siz

e 1

0

5

1 0

1 5

P e r io dB e fo re 1 A fte r 1 B e fo re 2 A fte r 2

Mea

n pe

rcen

tage

Lev

el 5

1 0

2 0

3 0

4 0

5 0


Mea

n pe

rcen

tage

Lev

el 4

1 0

2 0

3 0


Mea

n pe

rcen

tage

Lev

el 3

0

5

1 0

1 5

2 0


Mea

n pe

rcen

tage

Lev

el 2

0

2 0

4 0

6 0


Mea

n pe

rcen

tage

Lev

el 1

0

2

4

6

8

73


Table 4. Summary of results of repeated measures analyses of percentage of each degree of exposure. Tabuled values are F ratios of terms of interest. Significant values are indicated in boldface and level of significance expressed as: *: p < 0.05; **: p < 0.01 and ***: p < 0.001. d.f.: Degree of freedom.

4. Discussion

4.1 Indicator species for detect the SCUBA diving effect on coralligenous assemblages

Coralligenous assemblages are characterized by richness, biomass and high production, with values comparable to tropical reef assemblages, and as such, can be considered one of the most important and characteristic assemblages in the Mediterranean Sea (Ballesteros, 2006). These aesthetic characteristics make coralligenous systems one of the most attractive landscapes in the Mediterranean benthos and are often chosen by recreational divers. Despite

its importance, the coralligenous system is considered fragile, because its persistence is related to the maintenance of peculiar biotic and abiotic factors (Hong, 1983) and because it contains many sessile, long-living organism with fragile skeletons and slow growth rates (Ros et al., 1985).

Other studies have pointed to evidence of the impact of recreational SCUBA diving on this community of bryozoan and seafan species (Sala et al., 1996; Garrabou et al., 1998; Coma et al., 2004). Raising sediment is one of the most frequently effect produced by recreational SCUBA divers followed by contact with fragile organisms (Zakai and Chadwick-Furman, 2002; Barker

Source of variation D.f Exp 1 Exp 2 Exp 3 Exp 4 Exp 5

A) Between plots effects Unfrequented vs. Frequented (D) 1 2.00 38.55*** 3.44 1.72 0.14 Locality(D) 4 3.65* 4.20** 1.97 1.26 4.52** B) Whithin plots Year 1 6.84* 0.34 8.34** 8.24** 1.68 Year x D 1 0.96 7.87** 1.22 1.01 0.30 Year x Locality(D) 4 0.26 0.45 0.53 1.48 2.71* Before vs. After (T) 1 0.06 0.07 0.11 0.00 4.57* T x D 1 1.48 0.51 0.85 0.60 5.95* T x Locality (D) 4 0.42 0.85 0.64 0.50 0.65 Year x T 1 0.22 1.23 1.16 1.34 0.33 Year x T x D 1 0.09 0.29 0.05 0.04 0.18 Year x T x Locality (D) 4 0.78 0.40 0.60 0.55 0.18 Transformation Arc-sin Arc-sin Arc-sin Arc-sin Arc-sin

74

_________________________________________________________Chapter 4

and Roberts, 2004, Luna et al., 2009). Sedimentation and abrasion may affect benthic organisms, clogging filtering apparatus and inhibiting recruitment, growth and metabolic processes (Sala et al., 1996; Garrabou et al., 1998; Airoldi, 2003; Balata et al., 2005). These coralligenous species have low recruitment rates and long turnover-time (Mistri and Cecherelli, 1994), consequently damage produced by diver contact could be more severe and difficult to recover from with regard to these species. Rapport (1992) maintained that the appropriate choice of a set of indicators may provide early warning of problems and help monitor changes over large temporal and spatial scales. In this context, the find of an indicator coralligenous species which detect early the diving activity, before that this impact affect at these fragile species, would be a goal for the SCUBA diving management.

Indicators species are species whose parameters, such as density, distribution, population structure, presence or absence, are used as proxy measures of ecosystem condition (Vandermeulen, 1998). There are a number of studies published in the literature that outline criteria for selecting indicator taxa (Ott, 1987, Landres et al., 1988; Kremen, 1994). We maintain that in order to monitor the impact of diving, the most important criteria for reliable indicators are that they must be sensitive to change, easy to recognise, and be supported by reliable, readily measurable, as well as possessing wide temporal and spatial distribution. In addition, it is also considered to be important for there to

exist adequate baseline information and that their monitoring program is adequate in cost effectiveness. Sensitivity to change is an essential factor in an effective management tool to ensure that early impact detection is maximized while minimizing unpredictable fluctuations in population (Hilty and Merenlender, 2000).

4.2 Response of Halocynthia papilllosa to dive activity

Tunicates species are benthic active filter-feeders which are markedly influenced by hydrodynamic conditions and organic or inorganic suspended particles. While organic particles may be consumed as part of the organism’s diet, the simultaneous ingestion of inorganic particles produces abrasion and asphyxiating effects (Ramos-Esplá, 1991; Naranjo et al., 1996). Indeed, H. papillosa only inhabits natural and non-perturbed rocky areas and can therefore be categorized as a very stress-sensitive species, as well as a good conditions indicator (Naranjo et al., 1996).

We have observed differences in the H. papillosa population between frequented and unfrequented diving sites as well as changes after the main period of diving activity which has not been shown in the undived sites. Recently, a monitoring of recreational SCUBA diving behaviour was carried out at this same sampling zone (Luna et al., 2009): it was estimated that in a 45 minute immersion, each diver made over 100 interactions with the bottom, which resulted in the raising of sediment as well 22 organisms-contact. In the present study, we observed that

75


these unintentional impacts affect H.papillosa and produce three main consequences: (1) decreasing in density of individuals, and probably recruitment; (2) reduction of the proportion of large individuals and (3) displacing the individuals to more cryptic positions.

Density of individuals at control localities was higher than at impacted localities. Furthermore, at undived localities, this parameter increased over time while a decrease in the abundance after peak diving season was observed at the dive sites during both the two sampling years. The natural life cycle of H. papillosa previously described (Becerro and Turon, 1992; Ribes et al., 1998), is characterised by maximum growth rate values being reached in spring and summer and a reproductive season which spans the period between the second half of September and early November. Although we observed a high spatial and temporal heterogeneity in the sizes, the results obtained in our study for the control localities agree with this pattern in some of the two studied years. An increment of size 2 individuals and a dismissing of size 1 proportion were recorded from June to October in the second year. Whereas, size 5 individuals (the biggest size) presented a similar pattern; with a higher proportion of individuals of this size being found in October than in June. This tendency of somatic growth is more clearly observed at control localities than at impact sites. The effect of raising sediment and physical contact by divers may constitute the causes of this minor size increase, because an excessive sediment deposition causes

burial and clogging of the siphons and the branchial wall (Bakus, 1968). Additionally, a stress state caused by physical contacts may impede normal development and growth. Furthermore, sedimentation may inhibit settlement larvae, growth and metabolic processes (Airoldi et al., 2003; Balata et al., 2005). These synergic effects might explain the incidence of smaller individuals found at impact localities.

Another effect observed during the study is the cryptic position of the individuals in the most frequently dived localities. In the control localities the individuals of H. papillosa preferred more exposed locations than the individuals in the impacted sites. This suggests a preference by these suspension feeders for more conspicuous sites where water flow is more active, despite the degree of exposure and susceptibility to diver’s mechanical impact. On the contrary, at frequented dive sites, these more exposed sites might be considered less suitable for the development and settlement of the individuals and occupy more cryptic and less environmentally favourable positions. SCUBA diving activity might explain this particular spatial distribution because further evidence of similar effects produced by the diving has been shown by other filter feeding species (Sala et al., 1996; Garrabou et al., 1998).

4.3 H. papillosa as diving impact indicator

The sustainability of diving in particular areas depends on both the number and the level of training of individuals

76

_________________________________________________________Chapter 4

diving in those areas and on the capacity of the ecosystem to regenerate and recover from any damage incurred (Harriot et al., 1997). Environmental monitoring programmes are needed to detect environmental changes before the levels of damage caused by diving become critical because diving impact accumulates over time, a factor which could produce irreparable damage. Impacts are accumulated when the interval between successive disturbances exceeds the rate at which the affected environment is capable of recovering, sometimes referred to as “time-crowding” (Spaling and Smith, 1993). Our study shows that the impacts of SCUBA diving produced changes in H. papillosa population but there was no evidence of this damage accumulating over time for all the dived localities. However, in one of the most frequented localities of this study (Llosa Reef) the damage in the H. papillosa population had been shown to accumulate over time because the population could not recover after the peak diving season, before the start of the following peak diving season.

Benthic suspension feeders account for a significant fraction of the biomass and production in coastal marine ecosystems (Cloern, 1982). Recently, several studies have indicated that these benthic feeders play a very important role in plankton-benthos coupling (Gili and

Coma, 1998). Because of their efficient filtration ability (Vogel, 1994), they retain large quantities of plankton, and as such are considered to be responsible for a large share of the energy flow from pelagic to benthic systems (Gili and Coma, 1998). Furthermore, they capture large amounts of phytoplankton and may directly regulate primary production and indirectly regulate secondary production in littoral food chains (Kimmerer et al., 1994). For this reason, we believe that if the diving impact causes an accumulative effect on H. papillosa population, this effect may be manifested at other levels of the system. To detect this impact would be a goal for the monitoring diving activity which would allow to the early detection of the impact before this affect at other critical level of ecosystem

We maintain that the studied species meets the above criteria and we would emphasise its most notable characteristics as being: its sensitivity in the detection of impact, its ease of measurement and finally, its excellent cost effectiveness. For all the above reasons, we propose the use of H.papillosa as a highly suitable benthic coralligenous indicator species in the study of SCUBA diving impact in the Mediterranean.

77

Publication

Luna-Pérez, B-, Valle, C., Sánchez-Lizaso, J.L. 2010. Halocynthia papillosa as SCUBA

diving impact indicator: an in situ experiment. Journal of Experimental Marine Biology

and Ecology (Submitted)..

_________________________________________________________Chapter 5

Halocynthia papillosa as SCUBA diving impact indicator: an in situ experiment

Abstract

An in situ manipulative experiment was carried out to study the effect of SCUBA diving on the Halocynthia papillosa species and determine its utility as a bioindicator for this activity. Five experimental flapping impact intensities were considered: 0 (control), 3, 90, 270 and 330 flaps. Two experimental units, each with ten individuals, were assigned at random for each treatment. On each experimental unit, ten circular sediment traps were placed and then collected 24h after the experimental flapping event for subsequent analysis. The total abundance and survival of individuals of each size in each experimental unit were measured at eight time intervals: before the experiment and then 24 hours, 1 week, and 1, 3, 6, 9 and 12 months later. Repeated measures ANOVA was used to test the differences between treatments over time.

In the three most impacted sites, a reduction in the number of individuals was observed, with higher lethality rates in smaller individuals. Furthermore, in the two most impacted sites, the surviving individuals did not grow. A high negative correlation was observed between the abundance of individuals and the number of flaps (R2=0.96). The general regression model reflected a high correlation between the abundance of individuals and the sediment fractions considered. Our results indicate that the resuspension of sediments produced by flapping may explain the decrease of H. papillosa abundance that has been observed in areas of high diving intensity.

Keywords: manipulative experiment, tunicate, SCUBA diving impact, bioindicator, sediment resuspension.

1. Introduction

Coralligenous assemblages are characterized by richness, biomass and high production, with values comparable to tropical reef assemblages, and as such is one of the most important and characteristic assemblages in the Mediterranean Sea (Bianchi, 2001; Ballesteros, 2006). These aesthetic characteristics make coralligenous systems one of the most attractive landscapes in the Mediterranean benthos and are often chosen by recreational divers. Despite its importance, the coralligenous system

is considered fragile, because its persistence is related to the maintenance of peculiar biotic and abiotic factors (Hong, 1983), and because it contains many sessile, long-living organisms with fragile skeletons and slow growth rates (Ros et al., 1985).

SCUBA diving may result in the deterioration of this singular community because divers can easily damage marine organisms directly (e.g. by physical contact) and indirectly (e.g. by raised sediment) (Rogers, 1990; Rouphael and Inglis, 1995; Rouphael and Inglis, 1997; Tratalos and Austin, 2001; Zakai and Chadwick-Furman,

81


2002; Pulfrich et al., 2003; Luna et al., 2009).

Some studies on diving impact have shown that raising sediment was the most frequent effect produced by divers, followed by contact with fragile organisms such as gorgonians, corals and bryozoans, (Garrabou et al., 1998; Zackai and Chadwick-Furnman, 2002; Barker and Roberts, 2004; Luna et al., 2009). The combination of these two effects (the resuspension of sediments and physical contact) is the main stressor agent which the benthic communities are subjected to by diving activity. Little is known about the effect of sediment on rocky coralligenous assemblages, although it has been considered as a factor influencing their spatial and temporal variability (Morganti et al., 2001; Balata et al., 2005).

The large number of species and the high level of species diversity in the coralligenous community hinder assessment of the impact of diving on the community as a whole; selecting a single indicator species would therefore be advantageous (Sala et al., 1996). The most important criteria for reliable indicators among those considered were that they must be sensitive to change, easy to recognise and supported by reliable, readily measurable, as well as broad spatial and temporal distribution. Sensitivity to change is an essential factor of an effective management tool to maximize the detection of early impact, while minimizing unpredictable fluctuations in population (Hilty and Merenlender, 2000).

H. papillosa has been categorized as a stress-sensitive species, as well as a good indicator of natural conditions (Naranjo et al., 1996). A recent monitoring study showed that in sites with high SCUBA diving frequency, H. papillosa abundance is very low (Luna et al, in press). The study evaluated the sustainability of H. papillosa and proposed the species as an early indicator for assessing and monitoring the effect of diving on Mediterranean coralligenous communities. However, no cause-effect was established, and the immediate effect and mechanisms by which H. papillosa individuals are negatively affected by the SCUBA diving were not studied. The aim of the study was to determine this cause-effect by means of a manipulative field experiment, in order to shed light on the immediate and long-term responses of H. papillosa to disturbance caused by SCUBA diving. If the suitability of this species as a SCUBA diving indicator can be confirmed, it could be used in other Mediterranean regions.


2.1. Study area

The study was carried out in the Sierra Helada Natural Park, on the south-eastern Mediterranean Spanish coast, in an unfrequented and isolated diving site within a marine park known as Punta Albir (Fig.1), with a seabed occupied by a coralligenous community in an optimum state.

82

_________________________________________________________Chapter 5

Figure 1. Study area location.

2.2. Studied species

H. papillosa is an endemic Mediterranean species (Pérès, 1964) which also occurs in the Eastern Atlantic within an area which extends from the Canary Islands to the Berlengas Islands in Portugal (Ramos-Esplá et al., 2008). It is one of the most common solitary species on the rocky littoral zone of the south-western Mediterranean (Fiala-Médoni, 1973; Ramos-Esplá, 1991). It has a bottle-like form and can reach a height of 15 cm (Calvín-Calvo, 2000). Based on the maximum size recorded in bibliography, five size categories were used for this study: Size 1: 0.1-3 cm; Size 2: 3.1-6 cm; Size 3: 6.1-9 cm; Size 4: 9.1-12 cm; and Size 5: 12.1-15 cm. H. papillosa is a hermaphroditic species that reproduces once per year in the late summer and achieves a peak of gonadal development from March to September for ovaries and from July to September

for testes (Becerro and Turon, 1992). In contrast to other benthic filter-feeding species, H. papillosa does not exhibit aestivation (summer dormancy), and in fact it is in summer when activity levels and growth rates are highest (Ribes et al., 1998).

2.3. Experiment set-up

The experiment was run from August 2008 to August 2009. To test the effects of various short-duration flapping intensities on H. papillosa, five experimental impact intensities were considered: 0 (control), 3, 90, 270 and 330 flaps. These flapping levels were assigned on the basis of diver frequentation and number of flaps by a diver in a square meter in the area (Luna et al., 2009). Ten appropriately spaced experimental flapping units were chosen where ten medium-sized H. papillosa individuals were living. Two experimental units for each treatment were assigned at random, with ten sediment traps allocated in each one.

Prior to the experimental flapping, ten individuals of H. papillosa from each experimental unit were measured, and sediment traps were allocated. Experimental flapping consisted of performing a previously defined number of flaps (Impact 1 through 5) in order to raise sediment close to the H. papillosa individuals but without touching them directly. The 100 sediment traps were collected 24 hours after the experimental flapping event. The total abundance and survival of individuals of each size on each experimental unit were measured at eight sampling times:

83


before the experiment and 24 hours, 1 week, and 1, 3, 6, 9 and 12 months later.

2.4. Assessment of raised sediment

The circular sediment trap, used to collect the raised sediment during the 24 hours following the experimental flapping event, measured 86 mm in diameter and 96 mm in height. The sediment collected was separated by grain size using 2.0 and 0.0063 mm filters and the remaining solution with smaller particulates was used to obtain the mud fraction. The two sediment fractions obtained (gravel and sand) were labelled and dried for 48 hours at 110ºC to obtain the dry weight of each fraction (g/L). The ESS Method 340.2 (1993) was used on the remaining solution to obtain the organic and inorganic mud fractions (volatile suspended solids and total suspended solids). The sediment parameters used were total sediment, gravel, sand, mud and organic mud matter.

2.5. Statistical analysis

Two univariate techniques were used to analyse the response in abundance during the time of the study, and the survival of each size category at the end of study. By monitoring changes during the sampling time, the same sampling units were retested over time and data were not independent. Simple analysis of variance cannot be applied to this method of data collection, as it does not obtain an independent measure of residual variation, and does not fulfil the assumption of no interaction between treatments (Kingsford and Battershill, 1998). Repeated measures

(RM) ANOVA analyses correct the non-independence or autocorrelation in the data (Winer, 1971; Winer et al., 1991; Lively et al., 1993; Keough and Quinn, 1998; Kingsford and Battershill, 1998), and they were used to analyze our data to identify changes in abundance due to the effect of flapping over time.

Two experimental factors were considered, and the nomenclature was assigned according to common practice in RM experiments. Impact Level was the between-subject fixed factor, with five levels. Each experimental unit was the subject factor, which was randomized and nested within the Impact Level, with two levels (the replicates or experimental units). One blocking factor was used to define the time period in this experiment. Time was the within-subject fixed factor, with eight levels (one sampling before the experiment, and then 24 hours, 1 week, and 1, 3, 6, 9 and 12 moths later). A total of 80 replicates were obtained. The abundance of individuals in each experimental unit was the response variable analysed.

Before analysis, the assumption of normality was tested by examining residuals using probability plots (Winer et al., 1991). When repeated-measures analyses were used, the more conservative Greenhouse-Geiser and Huynh-Feldt correct F tests were also examined, because this correction provides some protection against violations of assumptions of compound symmetry (Winer et al., 1991; Yandell, 1997). These tests did not produce results markedly different from the

84

_________________________________________________________Chapter 5

uncorrected ones, and evidence of strong violation of these assumptions was not detected, as the epsilon value was close to one, so only the standard F tests are presented here. When the RM ANOVA F-test was significant, post-hoc analyses were conducted using the Bonferroni test for multiple comparisons (Winer et al., 1991).

Two-way ANOVAs were carried out to test differences in the sediment composition on each impact level, in absolute terms (Dry·g L-1) as well as percentages. The experimental design used contained two factors: Impact Level, a fixed factor with five levels; and Sites, a fixed factor with two levels. For each combination of factors, ten replicates were recorded (corresponding to each sediment trap), with a total of 100 replicates. The five variables obtained for the sediment analysis were tested separately with this design. Another two-way ANOVA was carried out to test differences in the survival percentage per size at the end of the experiment at each impact level (Underwood, 1981). The experimental design had two factors for this purpose: Impact Level, a fixed factor with five levels; and Size, a fixed factor with three levels (Sizes 2, 3 and 4). For each combination of factors, two replicates were recorded, with a total of 30 replicates. When the ANOVA F-test was significant, post-hoc analyses were conducted using Student-Newman-Keuls (SNK) multiple comparisons (Underwood, 1981). Prior to ANOVA, homogeneity variance was verified using Cochran’s test (Cochran, 1951), and arcsine transformations of the survival parameter were performed, as it

was taken as percentage (Underwood, 1997).

Two simple correlations were run between total sediments vs. number of flaps, and H. papillosa abundance vs. number of flaps, to quantify the possible relationship between them. A correlation matrix was calculated for sediment fractions in order to detect the existence of co-variables. A general regression model (GRM) was also calculated, with H. papillosa abundance as a dependent variable and the fractions of sediment as independent (predictor) variables. STATISTICA v.6.0 software was used for these correlations, variance analyses and the GRM.

Non-parametric multivariate techniques were used to analyse the response of H. papillosa to different impact levels and their relationship with the type of raised sediment at each impact level. Abundance data were not transformed before analysis, in order to detect minimum changes. Triangular similarity matrices were calculated using the Bray-Curtis similarity coefficient for abundance, and Euclidean distances for the raised sediment (Clarke and Warwick, 2001). A graphical representation of the response of H.papillosa to the impact level of the flapping was obtained using non-metric multidimensional scaling (nMDS). A bubble plot uses circles to show the relationship between abundance response and the total sediment raised at each level of flapping.

The Spearman’s correlation between abundance data and total sediment

85


fractions was determined using the RELATE procedure (Clarke, 1993). The BEST routine (BIO-ENV option) was performed to determine which sediment fraction best explained the response of H. papillosa to the impact level. All multivariate analyses were performed using the PRIMER-E statistical package (Clarke and Warwick, 2001).

3. Results

3.1. Analysis of the raised sediment

The percentages of sand and mud remained invariable, with no significant differences between proportions collected from each fraction at the different impact levels (F sand= 1.60, F mud= 1.44; p: non significant), but a difference was detected in the proportion of gravel between the two least impacted and the three most impacted treatments (F=56.1; p<0.01). ANOVA showed (Fig. 2) that the total sediment content was highest in the most impacted treatments (F= 3071.9; p<0.001). Post-hoc SNK analysis found that total sediment in less impacted places (Impact 1 and 2) was similar, and lower than in the more impacted places (SNK: 1=2<3<4<5). The number of instances of flapping was statistically correlated with the total sediment in the traps (R2=0.93, p<0.001).

Regarding the four sediment parameters obtained, two covariables were found: the mud fraction and the gravel fraction, the latter of which was removed for the following analysis. Differences between treatments were found for all sediment fraction variables tested with different fraction distributions for each impact level, as the post-hoc test showed (Table 1). Regression between flapping intensity and the concentration of each sediment fraction indicated a highly significant correlation for the four fractions, the highest being for the mud fraction (Table 1).

Figure 2. Total raised sediment (g.) of each fraction (mud, sans and gravel) at each impact level: 1,2,3,4 and 5.

Table 1. Summary of one-way ANOVA for each sediment fraction and correlation coefficients (R2), and correlation significances (p-value (R2)) of each sediment fraction with flapping intensity. D.f.: degree of freedom; M.S.: mean square; F: F ratio; Post hoc: results of SNK test; 1 to 5: impact levels.

Fraction df MS F p-level (F) Post hoc R2 p-level (R2) Mud 4 0.050 52.011 0.000 1=2=3<4<5 0.915 0.000 M.O. Mud 4 0.001 11.033 0.011 1=2=3=4<5 0.435 0.038 Sand 4 2.828 154.062 0.000 1=2<3<4<5 0.797 0.001

86

_________________________________________________________Chapter 5

Figure 3. Total mean and per size of H. papillosa abundance in each impact level along time study. Bars indicate the standard error.

87


Table 2. Summary of results of repeated measures abundance analysis. D.f.: degree of freedom; M.S.: Mean square; F: F ratio; p: p value.

Source of variation D.f M.S F p-value

Between effects Impact Level 4 191.500 547.143 0.000 Error 5 0.350 Within effects Time 5 3.737 53.381 0.000 Time x Impact Level 20 0.820 11.714 0.000 Error 25 0.070

3.2. Effect of flapping on Halocynthia papillosa

H. papillosa abundance decreased after the experimental flapping where the impact was highest (Impact 3, 4 and 5) (Fig. 3). In the experimental units subjected to the lowest flapping intensities (Impact 1 and 2), the abundances remained invariable throughout the study period. Consequently, repeated measures analysis showed significant differences between sampling times for some of the experimental flapping units (Table 2). Post-hoc testing (Bonferroni test) showed that the abundance in the most impacted experimental units decreased one week after the experimental impact (time 3) (Bonferroni: 1=2>3>4=5; p<0.001). The abundance at impact level 3 showed a reduction in the sampling at Time 3, but it remained steady until the end of the study (Bonferroni: Time 1=Time 2 > Time 3=Time 4= Time 5=Time 6= Time 7= Time 8; p< 0.001). Abundance at impact levels 4 and 5 decreased from sampling Time 3 until Time 5, when it stabilized until the end of study (Bonferroni: Time 1=Time 2 > Time 3 > Time 4 > Time 5=Time 6= Time 7= Time 8; p< 0.005) (Fig. 3).

Figure 4. Lineal regression between the H. papillosa abundance of each experimental unit one year later of experimental flapping event and the amount flaps.

One year later, the abundance on each experimental unit showed a statistically significant correlation with the flapping intensity (Fig 4). The survival of each size class was statistically different for each level of impact (F=12.08; p<0.001; Fig. 3). For Impacts 1 and 2, survival was equal for all size categories (100%). On the other hand, in experimental units of impact levels 3, 4 and 5, the survival of the smallest sizes (Sizes 2 and 3) were statically lower (SNK: p<0.05) than the survival for the biggest size (Size 4). At impact levels 4 and 5, all the individuals that survived belonged to the largest size class. During the sampling time, the growth of

88

_________________________________________________________Chapter 5

some individuals was recorded in the experimental units at impact levels 1, 2 and 3. In these cases, although the total abundance was constant, a change of proportion of individuals of each size category was observed.

3.3. Effect of raised sediment on Halocynthia papillosa

The H. papillosa MDS abundance plots showed a marked ordination of each experimental unit with regard to the impact level, and a clear relationship with the total sediment raised at each level (Fig. 5).

Fig. 5. MDS graph of H. papillosa abundance at each impact of the level of flapping, with bubble plot representing the total raised sediment.

RELATE testing demonstrated a highly significant correlation between abundance and total sediment variables (rho-Spearman=0.859; p=0.001). The BEST routine determined that mud and

a combination of this variable with the organic mater of the sediment were the variables that best explained the response of H. papillosa to the level of impact, with a partial Spearman’s correlation of 0.883 and 0.872, respectively. Analysis of the general regression model reflected a high correlation (R2=0.960; p<0.001) between H. papillosa abundance and each of the sediment fractions (Fig. 6), with beta parameters statistically significant for mud and OM content of sediments (Table 3). The mud and sand fractions were inversely correlated with abundance (negative values for Beta and b), while the OM content showed a positive correlation.

Fig 6. Predicted values vs. observed values of abundance of individuals according to calculated general regression model.

Table 3. Summary of the general regression model between the abundance of individuals and sediment fractions. Beta: the standardized regression; E.S. (Beta): standard error of beta; b: the raw regression coefficients; E.S.(b) standard error of b; p-level: statistical significance.

Fraction Beta E.S. (Beta) b E.S. (b) p-level

Mud -1.074 0.201 -29,423 5.518 0.002 M.O Mud 0.621 0.144 158.004 36.730 0.005 Sand -0.382 0.194 -1.4039 0.713 0.095

89


4. Discussion

4.1. Response of H. papillosa to SCUBA diving impact

Flapping produced sediment resuspension, which caused the death of H. papillosa individuals in the most impacted experimental units. The pulse of impact did not have an immediate lethal response on H. papillosa. Twenty-four hours after the flapping event, no mortality had been observed, although some of the most impacted individuals showed mucus covering their tunics. Although the flapping impact was of the ‘pulse’ type, the effect generated remained latent as a ‘pressure’, even when the sediment returned to the seabed. A drop in abundance was observed in the three most impacted experimental units, from three moths after the experiment to the end of the study. A higher lethality was observed in smaller individuals, as these were the first that died.

In natural conditions, H. papillosa concentrates its growth in spring and summer (Becerro and Turon, 1992), which was recorded during the study period (nine months after the flapping event), but only for the less impacted individuals (Impacts 1, 2, and 3). This may indicate that H. papillosa can adapt to episodic changes in sediment concentration around them up to a ‘critical concentration’, which is difficult to establish.

The physical characteristics of sediment (e.g. grain size, composition, density), and the quantity, degree, frequency and duration of raised sediment help to

determine the extent of disturbance by sediments, which can range from large-scale, sub-lethal chronicle stress, to an abrupt severe disturbance that locally disrupts the population (Airoldi et.al., 1998; Airoldi, 2003).

Tunicates species are benthic active filter-feeders which are markedly influenced by hydrodynamic conditions and by both organic and inorganic suspended particles. Ascidians are mainly characterised as non-selective suspension feeders (Jorgensen and Goldberg, 1953; Carlisle, 1966; Stuart and Klumpp, 1984) that continuously remove particles from passing water at a constant rate (Randlov and Riisgard, 1979; Klumpp, 1984). While organic particles may be consumed as part of the organism’s diet, the simultaneous ingestion of inorganic particles produces abrasion and asphyxiating effects (Ramos-Esplá, 1991; Naranjo et al., 1996). Resuspended sediment can change the quality of the diet and alter the animal’s ability to feed and gain energy (Arsworthy, 1993). A high inorganic fraction of the seston dilutes the nutritional content of the food and forces the suspension feeder either to use its energy reserves or to expend more energy obtaining sufficient organic matter (Windows et al., 1979; Vahl, 1980).

In general, ascidians are adapted to environments with low concentrations of particles and a small fraction of non-food particles in the water (Petersen, 2007). Although the filter feeding system of ascidians is non-selective with regard to particle size, the oral tentacles guarding the entrance of the

90

_________________________________________________________Chapter 5

brachial basket act as a coarse filter for large particles and sand grains (Godoy, 1974; Arsmorthy, 1993). H. papillosa is able to capture particles between 0.6 and 70 um (mud fraction) (Ribes et al., 1998), and the presence of coarser sediments has been observed as negative for these ascidians. Flapping produces the resuspension of both mud and larger particles, although only the smaller particles enter the brachial cavity. However, the synergic effect of higher concentrations inside the cavity and the stress generated by the constant rejection of larger particles may produce a deleterious effect on individuals. The entry of such particles induces siphon diameter constriction as well as ‘squirting’, and their subsequent ejection from the brachial siphon (MacGinite, 1939; Hoyle, 1953; Jorgensen and Goldberg, 1953; Armsworthy et al., 2001). Squirting is a movement of contraction of the mantle, which means that ascidians clear the brachial basket of waste material or in case of overloading (Petersen, 2007). During violent squirting, the mucus transports the feed particles across the brachial basket to the oesophagus (MacGinite, 1939; Goodbody, 1974) and may be expelled through the inhalant siphon (Werner and Werner, 1954). This response to high sediment concentration could explain the existence of mucus on some of the impacted individuals 24 hours after the flapping event.

Several studies have shown different responses by ascidians to an increase in suspended particles that affect the ‘clearance rate’ or ‘ingestion rate’, retention efficiency and/or absorption

efficiency, depending on the species, range of sediment concentration and particle size (Fiala-Médioni, 1979; Robbins, 1983; Petersen and Riisgard, 1992; Petersen et al., 1995; 1999; Armsorthy, 2001; Petersen, 2007). In the case of a similar species, H.pyriformis, it seems that it may compensate an episodic increase of sediment concentration by increasing the squirting frequency and reducing the siphon diameter. The absorption rate also increased, although a drop in absorption efficiency was observed (Armsworthy et al., 2001). Our results with H. papillosa indicated that this species has some physiological or behavioural mechanisms that may compensate the increase of sediments in the water, as no response was observed in the least impacted treatment. However these mechanisms are not sufficient to prevent damage to individuals when the impact exceeds a certain threshold limit. We remain unaware of all the physiological mechanisms used to compensate for this increased sediment. Other types of study would be needed to answer this question.

4.2. Use of H. papillosa as SCUBA diving impact indicator

The large number of species and the high level of species diversity in the coralligenous community hinder the assessment of the impact of diving on the community as a whole; selecting a single indicator species would therefore be advantageous (Sala et al., 1996). H. papillosa had previously been recognized as a good indicator of ‘natural conditions’ (Naranjo et al.,

91


1996). The impact of diving on H. papillosa has been shown (Luna et al., in press), but the cause and effect has not been recognized. In this study, the cause and effect between flapping, sediment resuspension and the response of H. papillosa is clearly shown.

The species studied is highly sensitive to the detection of impact, is easily measurable, and has an excellent cost-effectiveness, which are basic criteria

for being chosen as a good indicator. For all the above reasons, we propose the use of H. papillosa as a highly suitable benthic coralligenous indicator species in the study of the impact of SCUBA diving in the Mediterranean. H. papillosa would also be a good indicator for other types of impact that produce an amount of sediment in water, such as dumping or dredging.

92

_______________________________________________________General Discussion

GENERAL DISCUSSION

1. Increased of human-uses in MPA and associated conflicts

Tourism is now the largest single economic sector in the world, and its impact on coastal environments is considerable and extremely difficult to manage and limit (Davenport and Davenport, 2006). In recent years there has been a growth of tourism in coastal areas (Ribera 1991; Boudouresque and Ribera 1993; Davis and Tisdell, 1995), with around 70% of European holiday-goers choosing the Mediterranean coast for their vacations (European Commission, 2008). In 2008, of the 180 million tourists that visited Europe, about 40 million chose the Spanish Mediterranean coast as their destination, with leisure activities cited as being the main motivation (Spanish Ministry of Industry, Tourism and Trade, 2008).

For decades, the creation of marine protected areas (MPAs) has been considered to be the best way to restore natural communities and to protect marine ecosystems (Milazzo et al., 2002). MPAs are now being established around the world rapidly (Ballantine, 1995), and “marine-based” tourism is growing even faster (Ribera, 1991; Boudouresque and Ribera, 1995; Davis and Tisdell, 1995).

In the past 20 years, the number of visits to MPAs has increased globally (Dixon et al., 1993; Hawkins and Roberts, 1994; Kelleher et al., 1995), with an associated increase in the rates of participation in marine recreational activities, such as recreational fishing

(Francour et al., 2001; Lynch et al., 2004; Pawson et al., 2007), diving (Barker and Roberts, 2004; Hawkins et al., 2005; Asafu-Adjaye and Tapsuwan, 2008) or boat anchoring (Francour et al. 1999; Creed and Amado Filho, 1999; Francour et al., 2001; Milazzo et al., 2004; Montefalcone et al., 2006). Furthermore, the aesthetic appeal of MPAs and the facilities they provide, together with the increased public awareness of nature, all contribute to creating massive tourism in MPAs (Ribera, 1991; Richez, 1991, 1992, 1993; Capellà et al., 1998; Badalamenti et al., 2000; Milazzo et al., 2002). All these activities could damage the natural ecosystems, leading to changes in communities, and biodiversity (Sala et al., 1996; Zabala, 1996; Martínez et al., 1999; Brown and Taylor, 1999), in clear conflict with the conservation objectives of the MPAs. In the case of multiple use MPAs, all activities have to be managed in order to carry out the conservation and protection objectives, minimize conflicts and allow all users follow with their activities ecologically sustainable.

Technical advances in diving equipment, in addition to a rising interest in nature conservation and environmental recreation, have resulted in the increasing popularity of scuba diving (Zackai and Chadwick-Furman, 2002; Barker and Roberts, 2004). MPAs provide goods and services (i.e. attractive underwater flora and fauna, biodiversity and seascapes) that attract scuba divers to visit these areas (Hawkins et al., 2005; Dearden et al., 2006). This study (chapter 4) and many others papers (Hawkins and Roberts,

95


1997; Rouphael and Inglis, 1997, 2001; Garrabou et al., 1998; Hawkins et al., 2005; Luna et al., 2009; di Franco et al., 2009) have dealt with the impact of diving activity within MPAs. In the present study, as well as in others, it has been shown that scuba divers may affect organisms in several ways, both intentionally and unintentionally (Milazzo et al. 2002; Uyarra and Côtè, 2007). The follow of divers reflected that the damage was mainly due to physical contact with organisms (i.e. by divers’ fins, body and scuba gear) and increased sediment resuspension potentially affecting sessile organisms. Furthermore studies about the 'feeding' had found that the human presence in the water leading to changes in animal behaviour (Kulbicki, 1998; Milazzo et al., 2006). The MPA ‘diving attraction’ is a critical issue which has to be properly managed, mainly in Mediterranean where some habitat are highly vulnerably (i.e. coralligenous) and the diving activity is heavily seasonal and very concentrated in a few places, this fact has been clearly shown in Serra Gelada Marine Park during this work. A quotas system, the diversification of diving zones, creation of new diving zones (using i.e. artificial reefs or ships sunk) and proper zonation strategies are some alternatives which could help to control the diving attraction into MPA.

This ‘attraction phenomenon’ into MPA is valid too for recreational fishing. Paradoxically, fishing pressure may be even higher within the MPA than outside it as there may be a perception that, in the absence of commercial fishing, fish are larger and more

plentiful in MPA (Westera et al., 2003; Denny and Babcock, 2004). The recreational fishers know where are the fish, fact that was observed in the Sierra Gelada recreational fishing study (chapter 2). During the time of study, the recreational boat fishing effort was concentrated in zones with high abundance of target fishes, around to aquaculture installations and zones where the visual census had showed elevated abundances inside of Marine Park, according to other study carried out on the MPA (Sánchez-Jerez, et al., 2007).

Thus, MPA status and fishing gear restrictions, in fact, may result in exactly the opposite pattern to the intended, concentrating the recreational fishing effort in MPA which would lead to conflicts with the protection objectives. In France, Francour (1994) found that the density and biomass of fish in rocky reefs was lower in partially protected areas where the recreational fishing is allowed than unprotected areas. Similar results were found recently in Italy (di Franco et al., 2009) with no differences between partial protection zones inside of MPA and the locations outside the MPA. Due to such potential increase of recreational fishing pressure in MPA, there is a need to better asses the effectiveness of this partial fishing closed areas conservation tool taking into account the mobility of protected species as well as the size of total closed fishing areas. In light of the results, due to the large amount of biomass that recreational fishing is capable of extracting (chapter 1 and 2), an exhaustive control of this activity and their associated impact turns out

96


necessary. In those cases in which the effects of recreational fishing are confusing (for the dispersion of the activity, the nature of the species to be protected, etc.) or conflict with the objectives of protecting the MPA not allowing recreational fishing could be the more consistent management option.

The MPA of Columbretes or Tabarca are two clear examples of the ‘tourism attraction phenomenon’ of MPAs in the Mediterranean that necessarily led to change their management. The 30 nautical miles of distance between Columbretes Islands and land coast and their characteristics of small islands had been a natural protection for the marine environment. Since their established as MPA in 1990 (BOE 97/9432/1990) the number of visitors have increased notably, resulting in higher number of recreational fishers (78 recreational fishers in 1990 and 513 in 2000) and SCUBA divers (114 divers in 1990 and more than 1547 in 1994) (Jimenez, 1996; Lampérez, 2001). The increase in the amount of diving, induced the management authorities to prohibit the activity from December 1994 to February 1996 (Badalamenti et al., 2000). Since the reopening of the Reserve to diving in 1996, the number of divers has still increasing until 2856 divers in 1997 (Goñi, 1998) despite the more restrictive policy. Neither the isolation of the MPA or make more expensive recreational activities there deterred users to travel to Columbretes to enjoy their leisure activities. This situation of increased human use was adversely affecting marine environment and could jeopardize the objectives of protection for those was created it, with

the need to rethink the current management. In 2009 a new demarcation of the reserve was developed and expanded its size, with new geographical limits of special areas, and establishing a new management plan, using a more precise and exhaustive regulation of the activities and permitted uses (BOE 2/68/2009).

In Tabarca Marine Reserve, its relative proximity to the coast (only 3 nautical miles) further accentuated the situation of attracting visitors after its declaration as MPA in 1986 (BOE 112/1986 and DOGV 397/1986). In response to this growing recreational use and the marine environment impact associated new regulations have been implemented: in 2000 a quota management for diving activity was defined (BOE 184/2000 and DOGV 3868/2000), in 2006 finally any form of recreational fishing is banned (DOGV 5228/2006) and in 2007 the protected area was expanded 300 hectares (pending to publish). These two examples show the potential conflict that may arise between the objectives of protection and recreational use of MPAs, highlighting the importance of properly managing all permitted uses and implement adaptive management in line with the needs at all times.

The human element in marine protected areas must not be understated. The success of any protected area is closely related to how well user groups and stakeholders are identified and brought into the planning and management processes (Agardy, 2002). Despite this consideration, MPA research and the

97


resultant literature is generally lacking detailed accounts of the conflicts uses in MPAs, human responses and their final implications (Christie et al., 2003).

The conflicts between uses often stems from the marginalization of artisanal fisheries by other forms of resource utilization such as dive or recreational fishing. While this conflict (and its reporting) may be disconcerting to some environmentalists and scientists advocating MPAs, a careful consideration of the receptivity of local fishing communities to MPAs is fundamental for their long-term success (Agardy et al., 2003). Although enforcement is an important ingredient for the success of MPA management programs, it is important for long-term sustainability that wide stakeholder support exists (Peluso, 1992; Brechin et al., 2002; Lowe, 2003).

Some examples of the failure of MPAs success as consequence of an ineffective uses conflict resolution have been showed. In Twin Rocks MPA (Philipines) per example, the local artisanal fishers were gradually excluded of this zone and a great diving resort was established little by little (Christine et al., 2003; Oracion, 2003). Predictably, fishers, who initially voluntarily protected the no-take area (over a period of 7 years) as part of the community-based management regime, are either losing interest in the MPA or are plotting how to stop diving inside the reserve and reassert their influence. When asked why they are losing interest, informants expressed a general sense of mistrust of the dive industry

and concern that MPA management is no longer fair (Christie, 2004).

From a global perspective, such disregard for the community based regime represent a failure. This scenario of inter-stakeholder tensions, mainly between users interested in different recreational activities is common and has been shown in others MPAs such as Soufrière, St Lucia (Trist, 1999; Sandersen and Koester, 2000; Roberts et al., 2001), Bunaken National Park (Salm et al., 2000; Christie et al., 2003; Lowe, 2003) or Calamianes MPA (Fabinyi, 2008). In the Mediterranean context some similar examples have been also detected but with their own particularities.

If we look at the Mediterranean MPAs, there is great variation in the size of protected areas, the levels of protection, the activities carried out there, as well as between different wealth status, industrialization and economic development of zones (Badalamenti et al., 2000). The competition between recreational and artisanal fishermen for the natural resources in Mediterranenan MPAs is important, as the composition recreational catch has reflected in Tabarca, Benidorm and others Mediterranean regions (Lloret et al., 2008). Competition for the same resource and the exclusion of users can generate conflicts difficult to solve which may carry to the failure of MPAs in the future. In Tabarca MPA where only few artisanal boats can fish during a certain year period and with a particular gear, until the total prohibition of the recreational fishing, conflicts between those two types of

98


users were obvious. Nowadays, commercial fishing in Sierra Gelada Marine Park is not very substantial and has not any restriction. The conflicts between recreational and commercial fishing had not been identify still, but the clear overlapping both the targeted fished species and fishing places indicate that a potential conflict may be generated if the two activities are not properly manage in the future. On the other hand, recreational fishermen has the perception of to be in conflict with the aquaculture installation use of the zone, since they manifested that this activity contaminates the sea and their right to fish there has been usurped. Although the aerial survey data showed that the real situation is that the aquaculture installation concentrates a high recreational fishing effort around, probably due at the high catches there product of the attraction effect produced by the surplus feed of cages. By other hand, recreational fishermen have the perception that their impact on the fishing resources is negligible, nothing is further from reality as the estimated catches showed in Tabarca, Serra Gelada or in other Mediterranean regions (Morales-Nin et al., 2004; Lloret et al., 2008). In this sense, the recreational fishers perceive that commercial fishing or scuba diving are more harmful activities for the fishes communities and marine environment than their leisure practice. Giving know to users the real impact that recreational activities can produce, could help the understanding and acceptance of management measures adopted. This principle is also applicable to recreational scuba diving, a more

popular activity that is growing in the Mediterreanean MPA at the present.

Some conflicts related with scuba diving in MPA outside of Mediterreanean which have been cited previously could also found in MPAs here. Scuba diving conflict begin when the revenue derivate are not adequately shared with the local community. If local community is less directly involved in the economic life of MPAs, they may be less aware of the benefits brought by the creation of MPAs and are more likely to resent restrictions placed on their access. It seems highly desirable that local communities involve themselves in exploiting the economic potential of MPAs. Such potential are a lot and include involvement with diving centres, diving and snorkel guiding, tourist boat trips, hotels and hostels, outdoor equipment shops, local natural products, handicrafts, books, photography, films and restaurants offering local cuisine. If they are unwilling to do so, outsiders will move in and take over the task, causing ties between local communities and MPAs to be loosened (Badalamenti et al., 2000). In contrast to Sierra Gelada marine Park, diving in the Tabarca marine reserve was severely limited from the start to 50 divers per day (Ramos et al., 1992) and actually limited to 30 divers per day as a result of opposition from Tabarca fishers to divers. Diving is seasonal in both protected areas, with a large number of divers between April and November. In Tabarca the visitors to the island constantly increased from the establishment of the Reserve. In the same period, a doubling in the number

99


of restaurants and souvenir shops was observed (RANDOM, 1998). This is a clear and near example, where the local community took advantage of the opportunity created by the attraction of visitors to the AMP and economic benefits.

In order to solve the growing conflicts in the MPAs, an integrated approach to coastal and marine management is needed to consider both current and future interests. Many people have formulated and adapted various versions of Coastal Zoning Management Plans which can be applied in MPA for minimised uses conflicts of recreational activities (Wong, 1998; Kohn and Gowdy, 1999; Shi et al., 2001). A few common and indispensable strategies in the framework conflict resolution are described below (Lewis, 1996):

1. Getting Started/ Determining Roles: getting the conflict resolution process going.

2. Assessment: a structured attempt to assess the nature of the conflict, determines who is involved, and obtains other information that would be useful in designing an effective conflict resolution process.

3. Involving Affected Stakeholders: the communication/negotiation phase in which an attempt is made to find a resolution that addresses the interests of the stakeholders.

4. Implementation and Evaluation: solutions are implemented and evaluated.

Then, very often, the process repeats itself if more issues are identified, and more conflict occurs.

In the same way, to avoid or reduce the conflicts between economic exploitation for uses such as tourism, recreation and fishing and conservation, several strategies have been proposed, which are widely discussed below. In this sense, it may the useful of education programmes to increase environmental awareness and reduce the negative impact of visitors in popular sites. A rational zonation may also be an important management tool (Laffoley, 1995) which takes into consideration both traditional and new resource uses such as selective small-scale fishing and eco-tourism in delimited sub-areas. It is also important to define the social and biological carrying capacity or use limit of MPAs and to formulate suitable management responses to prevent deterioration and the consequent loss of their value (Davis and Tisdell, 1995).

2. Management strategies for recreational uses in MPA

2.1.Carrying capacity of recreational use and limit of acceptable change

The growing awareness that designation of protected areas does not ensure their preservation has stimulated an enormous level of discussion globally in recent years. While the biophysical characteristics of many protected areas remain the fundamental rationale for their initial designation, it has become quite clear that the values for which these areas were initially protected can be threatened by unmanaged or poorly

100


managed recreational use (Wong, 1998). The application of the carryingcapacity concept to recreational areas became popular with the development of a methodology for assessing this carrying capacity in 1973 (Urban Research and Development Corporation, 1980). This methodology was developed because of concerns about “resource overuse and user overcrowding” of recreation resources.

It was quickly recognized, that recreational carrying capacity was comprised of at least two components: (1) a biophysical component concerning impacts of visitors and their activities upon the resource, and (2) a social element dealing with the type and quality of experience users received during their activity (Stankey and McCool, 1984). This recognition had the net effect of complicating discussions about carrying capacities because little was known about visitor experiences and the recreation production process, their interactions with biophysical processes and conditions, and how establishing a carrying capacity would deal with such questions. The literature offers a several results of acceptable or suitable carrying capacities of recreational activities in marine protected areas around the world (among others, Hawkins and Roberts, 1992; Scura and van’t Hoff, 1993).

The major assumptions of the carrying capacity concept are that the amount of impact is related to the amount of use, that decreasing the amount of use will decrease the impacts, and that it is possible to calculate the number of users of each leisure activity, below

which impact will be acceptable (Mann, 2003). The problem is that the recreation impacts differ with the type of use (SCUBA diving, fishing, anchoring or trampling), timing of use, distribution of use, environmental setting, management actions, and with people’s expectations and norms (Mann, 2003).

Decreasing the number of users, may not necessarily lessen impacts. In fact, much research has shown that many impacts of recreational use are a function not so much of the number of people, but of their behaviour (McCool, 996). As has been observed on this study and others, the SCUBA diving impact are influenced for the underwater behaviour of divers, the type of dive and their intrinsic profile, experience (Roberts and Harriot, 1994; Medio et al., 1997; Rouphael and Inglis, 2001; Barker and Roberts, 2004; Uyarra and Côte, 2007; Luna et al., 2009). During the underwater monitoring surveys carried out in Serra Gelada MPA (Chapter 4) different behaviours with different effects on the habitat were observed between different type of users regarding their intrinsic and diving profile, experience and type of dive. Those results support the idea of the diving impact will depend both of number of people as their behaviour and the carrying capacity for a diving activity have to take into account this it.

The extent of anchoring impact is related with the type anchor used (Walker et al., 1989; Milazzo et al., 2004), the substrate characteristics (Francour et al., 1999; Cecherelli et al., 2007) and the health status of the

101


habitat impacted (Montefalcone et al., 2006).

The recreational fishing effect can be influenced by fishing equipment (gear, hook type, rood, bait) (Muoneke and Childress, 1994; Meyer, 2007; Lloret et al., 2008). The catch-and-release practice does not have the same impact on the community as the extractive recreational fishing. The effect of trampling is affected by the morphology and nature of marine organisms (Liddle, 1991), the intensity of damage (Milazzo et al., 2002) and the characteristics of the substrate (Wynberg and Branch, 1997; Eckrich and Holmquist, 2000).

This concept of carrying capacity thus tends to set specific limit to use, when the real goal is to limit adverse impacts in whatever way is most appropriate. By reformulating the question of “how much use is too much” to the related question of “how much impact is acceptable” it is possible to focus on the real issue and to set more flexible and realistic leisure activities levels.

This new viewpoint is known as “Limit of Acceptable Change” (LAC). The LAC model was development in 1985 and was first applied in North American terrestrial wilderness setting (Stankey et al., 1985). The Limits of Acceptable Change model better encompasses the critical interaction between human and natural systems at every level (Howard and Potter, 2002). The LAC model is considered by McCool and Cole (1997) to be the simplest available approach for effectively dealing with the complexity of the real world.

In order to define “acceptable change” from a scientific perspective, it is first necessary to define the scope of the system being managed, and to determine whether the system should be managed from an ecocentric or an anthropocentric viewpoint. Taking into account the MPA objectives an ecocentric viewpoint that is geared towards the protection and conservation of natural resources would be the most coherent option.

Applying a LAC management model at MPA multiple-uses consists for each recreational activity of (1) selecting key indicators of setting conditions, (2) specifying quantitative standards of quality for each selected indicator, (3) applying different LAC standards within different zones (4) implementing a range of management actions to maintain selected standards of environmental quality over time (McCool and Cole, 1998). The LAC framework is also an iterative, adaptive management process that includes ongoing monitoring and adjustments (Hammitt and Cole, 1998; McCool and Cole, 1998). The utility of the LAC concept in marine recreational activities planning and monitoring has been illustrated by previous research (e.g., Shafer and Inglis, 2000; Dinsdale and Harriott, 2004; Sorice et al., 2004).

A good example of the implementation of this management tool was investigated recently by Leujak and Ormond (2008) in a study carry out in Red Sea reef to trampling over the reef flat and could be applied for other recreational activities and other marine environments. The first step was

102


examined the relationship between coral cover and intensity of trampling. The limit of acceptable change was established, deciding what amount of coral damage the reef managers could permit before limiting the access. This same example could be applied to establish the limit of acceptable change for a Mediterranean diving zone. The relation between the abundance of Halocynthia papillosa and diving activity has been shown in this study (Chapter 5; Luna et al., 2010). A conservative acceptable limit of change could be the one who would ensure a constant population of H. papillosa. This would mean that starting from an initial level of abundance if, during the peak seasonal diving there were a decrease in population, during the period of "rest" of the diving activity was given a recovery in this population that would restore initial level. This recovery is related to the own resilience of the population which will depend on the intensity of the impact and the extent of damage. This threshold of acceptable change chosen, would have a determinate number of visitor (per example) in determinate situation, but if this situation changes, with a previous briefing (per example), the visitors could reduced their environment impact and at this same LAC now would correspond at higher number of visitors.

The LAC use as management tool is a complicated task since implies to know the relationships cause-effect of each recreational activity, the possible indirect effects derived from it as well as choose a proper and early indicator of each impact. In most case the direct or indirect effect produced by an

activity may affect to more than one species group and those effects have to be taken into account in the impact assessment and the LAC determination. Finding this suitable and early indicator for each recreational activity is not an easy task and should not been make lightly, as has been internationally recognized (Vandermeulen, 1998; Bustos and Fried, 2003; Rochet and Trenkel, 2003; Rogers and Greenaway, 2005; Cury and Christensen, 2005; Methratta and Link, 2006). Some studies have shown the deletereous effect of diving on certain species of bryozoans or seafans (Sala et al., 1996; Garrabou at al., 2004) pointing out their useful as diving impact indicators. In this work, we suggested H. papillosa as a possible indicator for monitoring the SCUBA diving impact which could be used in the LAC implementation for the diving activity monitoring. Due to the complexity of coralligenous community, maybe the best option is not to choose a single indicator if not the use of an indicator sets to establish this LAC. More studies focusing in this goal could be making in the future to elaborate a monitoring programme to manage suitably the SCUBA diving and achieve the sustainable use.

2.2. Zonation as management tool for recreational activities

Zoning is often a key component of MPA recreational activities management strategies. Zoning plans subdivide a managed area into two or more sub-areas, define classes of activities, and specify which activities are permitted and prohibited in each zone (Kenchington and Kelleher, 1995).

103


A key challenge is to integrate interdisciplinary information when formulating and implementing zoning plans. In many cases, data are available on just one aspect of the MPA, such as ecological values (e.g., Done, 1995; Riegl and Riegl, 1996; Edinger and Risk, 2000; Beger et al., 2003), or community perspectives (e.g., Inglis et al., 1999; Shafer and Inglis, 2000; Williams and Polunin, 2000; Pollnac et al., 2001). Relatively few published studies have integrated both natural and social science perspectives into marine tourism planning and management.

In the case of zoning to manage recreational activities as anchoring, trampling or SCUBA diving, a key ecological criterion to consider is the vulnerability of impacted habitats. The numerous research of trampling impact on the tropical coral reefs, probably the most studied recreational impact around the world, indicates that branching corals are more vulnerable than massive, plate-like, digitate, or submassive corals (Allison, 1996; Plathong et al., 2000; Schleyer and Tomalin, 2000; Walters and Samways, 2001; Zakai and Chadwick-Furman, 2002). Research also shows that soft corals and encrusting corals are least vulnerable to trampling (Riegl and Riegl, 1996; Plathong et al., 2000; Schleyer and Tomalin, 2000; Tratalos and Austin, 2001). These findings suggest that trampling impacts can be reduced by restricting access to reefs containing vulnerable hard and branching corals, while promoting access to less vulnerable reefs.

In the particular cases of Mediterranean habitats most frequented by diving, Lloret et al., (2006) evaluated the specific vulnerability of those communities. Cheoung et al., (2005) assigned a intrinsic vulnerability value to fishing based in the risk of extinction for some fish species. Those studies, together with others future similar studies, may constitute a useful tool for managing and zoning of scuba diving or recreational fishing in the Mediterranean MPAs. Identify which species or communities are most vulnerable to each type of impact and locate the places where they would be the first step in establishing some kind of special accord protection and restriction of use in these areas.

In the Mediterranean, the community more vulnerable to the diving impact is the coralligenous (Sala et al., 1996; Garrabou et al., 2004) and so these areas should be managed with special care and referred to in the zoning plans as areas with a high level of protection and controlled use.

In the case of the fish community, given their mobility, these special protection areas can be areas of concentrated high biomass for certain species or areas where key processes occur for the development of these species (feeding, nursery, reproduction, etc). In the study of recreational boat fishing in the Serra Gelada Marine Park of this work the areas around the aquaculture installations concentred high levels of effort, probably for the high fish abundances extracted there. The fish assemblage aggregated at these cages is composed mainly by pelagic species

104


The establishment of some special protection status around those zones may be a useful tool for protect a great amount of these pelagic species that feed near the fish farm cages. However this protection measure would not have any effect on coastal resident species. The mobility and home range of species objetive of the proection is a key issue to be considered when choosing the size and location of the area or areas to be protected.

MPA zoning plans should also consider biodiversity values and the most diverse habitats have to deserve higher degrees of protection within MPA zoning plans.

A range of social and economic factors related to recreational uses and tourism must also be incorporated in MPA zoning plans. Managers should ensure that (1) zoning and related access restrictions maintain or enhance the capacity of the MPA to provide satisfying visitor experiences, and (2) degradation of habitat amenity values does not occur over time. Meeting these objectives requires a good understanding of what visitors to MPA environments seek within their leisure experiences (Shafer and Inglis, 2000). Previous research has shown the critical importance of biophysical setting conditions, including water clarity, underwater rock formations, and the quality and diversity of the visited habitats as well as fish communities (Pendleton, 1994; Shafer and Inglis, 2000; Williams and Polunin, 2000; Bennett et al., 2003). Social setting conditions are also important, including the number of other people, the

behaviour of others, and the types of activities undertaken at a site (Inglis et al., 1999; Shafer and Inglis, 2000).

Judgements of what constitutes a ‘‘desirable’’ condition in natural areas vary widely among users (Manning, 2003). To accommodate a diversity of visitor preferences, park managers should provide a range of settings within different zones, a spatial visitor management strategy known in the literature as a ‘‘Recreation Opportunity Spectrum’’ (ROS) (Clarke and Stankey, 1979). A ‘‘Marine Recreation Opportunity Spectrum’’ (MROS) has also been developed as a more specific version of ROS for marine environmental applications (Orams 1999). Separate MROS classes are defined based on differences between sites in accessibility, degree of social interaction, presence of man-made physical structures, proximity to the shoreline, amount of anthropogenic impacts, and the amount and type of management regimentation (Clarke and Stankey 1979; Orams 1999). Through zoning and access restrictions, managers can aim to provide a sufficient range of recreational opportunities that meet the recreational aspirations of a diverse range of visitors without damage the biological integrity of MPA (Knopf 1990; Orams 1999).

2.3. Marine education and awareness programmes

Frost and McCool (1988) argue the case for education as an effective management tool, stating that "If the visitor understands the rationale for the regulation, there may be more

105

(Fernádez-Jover , et al 2010).


understanding of the regulation and consequently, more voluntary compliance with it... Thus perceptions of the adequacy of information seemed to be a factor in the acceptance of restrictions. With a good rationale and careful explanation of the rationale, visitor regulation, at least in some places, may enhance recreational experiences". In addition, good interpretation enhances visitor experience and gives greater satisfaction at their experience participating in recreational activities in the MPA.

Interpretation should not only comprise information on the marine environment and marine life found in the area, but should also communicate MPA policies and management objectives to visitors (Wong, 1996). There also needs to be an expansion and improvement of the interpretation materials used, employing a variety of methods. Brochures, posters and videos should be multi-lingual, taking into consideration the major nationalities that visit the MPA.

Staff of the MPA, as well as tour operators, owner of boat tours and dive operators should be adequately trained, to increase general knowledge and conservation awareness of the marine environment to the recreational users. In addition, training on nature interpretation should be provided for a qualified personal of the MPA. Training should endeavour to improve skills, develop understanding, raise motivation, and help ensure that limited resources for conservation and enjoyment of the marine environment are used more effectively (WWF, 2006).

In the cases of SCUBA diving, per example, briefings given by diving centres before the dive and underwater intervention by a dive leader were highly effective at reducing the diving impact (Medio et al., 1997; Barker and Roberts, 2004; Uyarra and Côte, 2007, Luna et al., 2009). These findings confirming that deliberate contacts may be reduced by the implementation of simple measures by the diving centres. Attending a briefing emphasizing the importance of buoyancy control and careful action (educational tools) increases the environmental awareness of recreational divers and might reduce diver damage at dive sites. Moreover, the presence of a leader of the diving centers during dives is a clearly effective tool in minimizing scuba divers’ physical impact on their environment, because dive leaders can take measures when they see divers behaving inappropriately.

The education of local residents on the long term benefits of the MPA is crucial to ensure acceptance of it, and to avoid any conflict or resentment with regards to any infringement of rights, especially pertaining to fishing. Fishing communities should be specifically targeted to ensure that they understand how the MPA helps to sustain their livelihood.

In addition, a code of good practice should be formulated for activities like swimming, snorkelling, anchoring, diving and recreational fishing, as the elaborated by the FAO recently (FAO, 2008). This must be effectively communicated to visitors to ensure that they are aware of the negative

106


consequences of their activities and to foster more responsibility towards the surrounding marine environment.

2.4.Long term monitoring programmes

Recreational activities and the tourism in temperate MPA has yet to be well-documented, monitored and evaluated, leaving managers with insufficient objective information for decision making and actions (Wong, 1996). In many cases, this has led to ad hoc and reactive responses rather than proactive and planned decision making. A suitable monitoring and evaluation of recreational impacts will enable the maintenance of a record of marine resource and possible conflicts between users over time, as well as will help managers assess the effectiveness of any management actions implemented.

Recreational users should continue to being monitored during the MPA life if their long time success is required. Social aspects such as users response to crowding and visitor satisfaction should also be monitored by means of a simple questionnaire, per example, a cheap and easy measure.

The monitoring impact of recreational activities undoubtedly has to carry out periodically. The specification of what level of effect should be detectable by any monitoring programme implicitly involves setting limits of acceptable change (LAC), which previously has been explained. When the monitoring results are obtained, any observed changes are compared against the chosen critical threshold and the need or not to apply some sort of management

action designed to rectify the situation will decide.

107

____________________________________________________________Conclusions

CONCLUSIONS

Chapter 1

1. Long-term evolution of recreational fishing effort at Tabarca Marine Reserve showed an increasing trend in recent years. Usually the highest angling effort was recorded during summer season.

2. Typical angler in the Tabarca Marine Reserve is a middle-age male, resident of near area who go fishing alone or in pairs. Many anglers fish all year but others only in the warmest season, mainly during weekend days. The mean expenditure per year per angler amounted to €902, on goods and services directly related to their fishing activities in the Tabarca Marine Reserve.

3. A high proportion of anglers did not have a fishing licence and some anglers were fishing illegally, as they had exceeded the number of lines allowed.

4. The greatest daily effort was recorded during summer weekends; and the lowest effort was recorded during autumn weekdays, according to in situ surveys.

5. The most abundant catch species were: Diplodus spp, Symphodus tinca, Sparus aurata and macrophagous as Dicentrachus labrax, Dentex dentex, Labrus merula.

6. The highest value of CPUE and BPUE was detected in summer and the lowest value in winter.

7. Angling activity in the Tabarca Marine Reserve would be capable of extracting 4.9 tons of fish per year.

8. The recreational fishing catch of Symphodus tinca and Serranus scriba proved to be greater than the artisanal fishing landing. For Dentex dentex, Muraena helena, Sciaena umbra, Diplodus spp and Sparus aurata, the angling catch was not greater than it but was of considerable proportions.

Chapter 2

9. The typical recreational boat fishermen is a middle age man, resident on the same town council, has an experience of more than 21 years and great proportion of they are not members at fishing club. They fish an average of 117.14 ± 12.5 days year-1 and 4.91 ± 0.16 hours day-1. Most of them knew the existence of the Marine Park protection figure although it does not have influence on their decision of fishing there.

10. Regarding to the relationship between recreational boat fisher and other users, according to their own viewpoint, the most conflictive sector is the aquaculture.

11. Recreational boat fisher identify that the species most frequently caught are: Seriola dumerili, Trachurus sp, Coryphaena hippurus, Pagellus erythrinus, Pagellus acarne, Scomber scombrus and Euthynnus alletteratus.

12. The recreational boat fishing effort showed is very high, mainly focused on the buffer zone and the special protection area. In the buffer zone, three

111


important areas were identified due to their highest effort density, two of them surrounding aquaculture installations.

13. The annual total catch of recreational boat fishing in the Serra Gelada Marine Park was estimated in 24.90 t.

Chapter 3

14. Most divers (96.7%) made at least one contact with the seabed, with a mean contact of 41.20 ± 3.55 (mean ± s.e.) per diver per 10 min.

15. Flapping was the most frequent type of contact, and the main damage by this action was to raise of sediment.

16. Contact with the seabed was greater for males than for females, inexperienced divers than for experienced divers, and camera or lantern users than for non-users.

17. Divers accompanied by a dive leader or who had been briefed about avoiding seabed contact before undertaking a dive produced fewer contacts than for non-accompanied or non-briefed divers.

Chapter 4

18. We observed bigger and more abundant individuals of Halocynthiapapillosa at undived sites than at frequented dived sites during the period of study.

19. Furthermore, individuals of H. papillosa in the most frequented localities occupied more cryptic positions than in the undived localities.

Chapter 5

20. A negative effect of experimental flapping on H. papillosa individuals in the three most impacted sites was observed. In those sites, a reduction in the number of individuals was detected, with higher lethality rates in smaller individuals. Furthermore, in the two most impacted sites, the surviving individuals did not grow.

21. A high negative correlation was observed between the abundance of individuals and the number of flaps.

22. Our results indicate that the resuspension of sediments produced by flapping may explain the decrease of H.papillosa abundance that has been observed in areas of high diving intensity.

23. H. papillosa is proposed as a highly suitable benthic coralligenous indicator species in the study of the impact of SCUBA diving in the Mediterranean, which would also be a good indicator for other types of impact that produce an amount of sediment in water, such as dumping or dredging.

112

______________________________________________________________________General Abstract

RESUM

1.Introducció General

La biodiversitat marina es extremadament valuosa per a l’home, ja que suposa aproximadament el 60% del valor econòmic de la biosfera (Constanza et al., 1997)i proporciona importants servicis bàsics per a la vida humana (Daily, 1997; Moberg i Folke, 1999). Tot i el seu immens valor, aquest no hi ha estat prou reconegut i els ecosistemes marins estan en l’actualitat deteriorant-se ràpidament degut principalment a les activitat humanes. Resulta obvi que els oceans es ressenteixen actualment del usos humans tals com la indústria del transport i el turisme, la sobrepesca, les descarregues de aigües residuals i la introducció d’espècies exòtiques (revisat per Norse, 1993; Peterson i Estes, 2001; Steneck i Carlton, 2001). Cal tindre en compte que aquestes amenaces a la biodiversitat marina son acumulatives i poden tindre efectes sinèrgics (Hughes, 1994). La capacitat dels oceans per a recuperar-se de les perturbacions globals i mantenir els ecosistemes en un estat saludable està debilitant-se (Wor, et al., 2006). Addicionalment a aquestos problemes derivats del ús humà incontrolat, la incertidumbre augmenta encara més degut al efecte del canvi climàtic. Conseqüentment, organismes internacionals han estat implicats en tasques de conservació i protecció dels oceans i han declarat la urgent necessitat d’aplicar noves i dràstiques mesures de gestió (IUCN, 1998).

Entre altres, les Àrees Marines Protegides (AMPs) han estat proposades com una de les mesures de gestió mes adequades per a fer front a la impredictible alteració a la que es veuen sotmesos els ecosistemes marins i mitigar els seus efectes (Lubchenco et al., 2003). Els efectes positius de les AMPs han estat provats al voltant del seus límits (per a una revisió consultar Halpern i Warner, 2002; Halpern, 2003; Gell i Roberts, 2003; PSICO, 2007; Claudet et al., 2008). Les AMPs proporcionen refugi per a espècies amenaçades, prevenen la degradació del habitats i possibiliten el desenvolupament natural de les comunitats marines. El establiment de les AMP podria afavorir el “spillover” de adults i juvenils que podrien recolonitzar les àrees adjacents, ajudar a recuperar els stocks de peixos o restaurar els ambients degradats.

Existeixen molt tipus d’AMPs amb diferents esquemes de protecció en funció dels seus objectius de creació. En general, els tipus de protecció aplicats en les AMPs de la Mediterrània son també molt diversos i posen en evidència les grans diferències politiques i culturals que existeixen entre els diferents països (Wood et al. 2008). La majoria han estat classificades com àrees marines de usos múltiples (Harmelin, 2000, Badalamenti et al., 2000, Francour et al., 2001). Aquestes AMP de usos múltiples intenten aconseguir el balanç entre la protecció de la biodiversitat i l’ús humà.

Actualment el turisme es un important sector econòmic en el mon i el seu impacte sobre els ambients costers es

115

Anthropic impacts in Mediterranean MPAs__________________________________________________

considerable i extremadament difícil de gestionar i limitar (Davenport i Davenport, 2006). En els últims anys ha hagut un important creixement del turisme cap a les àrees costeres (Boudouresque i Ribera, 1993; Davis i Tisdell, 1995; Ribera, 1991), especialment en el cas Mediterrani, el qual es el principal destí vacacional per al 70% del turistes que visiten Europa (Comissió Europea, 2008).

El suposat valor estètic de les AMPs ha portat a un turisme massiu a aquestes àrees (Ribera, 1991, 1992; Richez, 1991, 1992, 1993; Capellà et al., 1998; Badalamenti et al., 2000; Milazzo et al., 2002a), puguent arribar a convertir-les en verdaders “Parcs submarins”, com es el cas de la Reserva de les Illes Medes (Ribera-Siguan, 1992). Aquest turisme en massa implica el desenvolupament d’una sèrie d’activitats d’esplai, com son la pesca esportiva (Francour et al., 2001; Lynch et al., 2004; Pawson et al., 2007), el busseig (Asafu-Adjaye i Tapsuwan, 2008; Barker i Roberts, 2004; Hawkins et al., 2005) o l’ancoratge d’embarcacions (Creed i Amado Filho, 1999; Francour et al. 1999; Francour et al., 2001; Milazzo et al., 2004; Montefalcone et al., 2006). Totes aquestes activitats poden afectar als ecosistemes naturals, produint canvis en les comunitats i la biodiversitat entrant així en un clar conflicte amb els objectius de conservació de les AMPs (Sala et al., 1996; Zabala, 1996; Brown i Taylor, 1999; Martínez et al., 1999). El increment de la presència humana i les infraestructures accessòries poden superar la capacitat de carrega ambiental i produir conflictes difícils de

resoldre. En aquest sentit, les AMPs declarades sense una gestió adequada que no tinguen en compte al totes les parts implicades, poden arribar a ser víctimes del seu propi èxit degut al efecte de la “sobrefrequentació” (Ramos-Esplà, et al, 2004).

Els estudis dels impactes humans en AMPs han evidenciat l’important efecte que poden tindre aquestes activitats i la potencial amenaça que suposen per a la diversitat de les comunitats marines (Brown i Taylor, 1999). La major part d’aquestos estudis han estat fets en zones tropicals i temperades fora del Mediterrani, probablement degut a que l’establiment de AMPs com a mesura de gestió en el Mediterrani es una practica relativament novadora en comparació a altres zones del món (Milazzo et al., 2002a).

Al inici d’aquesta tesi, un total de 146 estudis publicats varen ser revisats en tal de abordar els coneixements actuals d’aquestos impactes i localitzar les carències existents en les AMPs de la Mediterrània. Aquestos estudis havien estat fets arreu del món i se centraven en el efecte produït per les activitats antròpiques tals com la pesca esportiva, el busseig, el ancoratge o el pisoteig sobre les comunitat marines. Basat en la falta de coneixement existent en la Mediterrània, revelat per aquesta revisió bibliogràfica, i degut a la seva importància socioeconòmica, la pesca esportiva i el busseig han estat les tasques escollides per a ser investigades al llarg d’aquesta tesi.

El estudi de la pesca esportiva s’ha abordat des de dues dimensions,

116


corresponent als dos primers capítols, primer el estudi de la pesca recreativa des de terra i segon el estudi de la pesca recreativa des de embarcació.

Per tal d’estudiar el impacte del busseig, el primer pas va ser conèixer el comportament in situ dels bussejadors, que correspon al capítol tercer. El segon pas, que correspon al capítol quart, va ser detectar l’existència d’espècies sensibles al impacte del busseig i que pogueren ser empleades com indicadors d’aquest impacte. Finalment, com a últim capítol, es va plantejar un estudi per tal de corroborar experimentalment la utilitat d’aquest possible indicador del impacte del busseig.

Cadascun del capítols han tingut els seus propis objectius específics així com la seva metodologia pròpia, que es detalla capítol per capítol a continuació junt amb una breu descripció del resultats més rellevants i les conclusions més importants obtingudes.

2. Impacte de la pesca recreativa en una Àrea Marina Protegida Mediterrània

La pesca recreativa és una de les activitats d'oci més populars i que practica un gran nombre de persones donant en conseqüència uns alts nivells d'esforç (Cowx, 2002; Pitcher i Hollingworth, 2002; al Westera et al., 2003; Cooke i Cowx, 2006), que en alguns casos poden arribar a igualar o superar l'esforç i la captura de la pesca comercial (Pollock, 1980; Dunn et al., 1989; Steffen et al., 1996; Beal et al., 1998; Murray-Jones i Steffen, 2000). La pesca comercial i esportiva tenen

similars efectes demogràfics i ecològics sobre les poblacions de peixos, i poden tenir conseqüències ecològiques i econòmiques igualment greus (Post et al., 2002; Coleman et al., 2004). Aquesta activitat d'oci pot tenir diverses conseqüències biològiques sota les poblacions objectiu, com ara el truncament de l’estructura de talles, una disminució de l'abundància i biomassa de peixos, així com canvis en la composició de les poblacions (Coll et al., 1999; McPhee et al. 2002; Westera et al., 2003; Coleman et al., 2004; Cooke i Cowx, 2004, 2006; Lewin et al., 2006, Rius, 2007). Per aquesta raó, per tal de garantir una ordenació pesquera eficaç, resulta necessari tindre en compte per part de la gestió, tant la pesca recreativa com la pesca comercial, especialment en les zones marines protegides. A pesar de la seva importància, l'impacte de la pesca recreativa sobre les poblacions de peixos no ha estat científicament tan discutit com el efecte de la pesca comercial (Cooke i Cowx, 2004, 2006).Concrectament, en la Mediterrànea existeixen molts pocs estudis referents a la pesca recreativa i els seus efectes sobre les comunitats (Coll et al., 2004; Morales-Nin et al., 2005; Lloret et al., 2008).

Aquest estudi no únicament analitza la activitat de pesca recreativa a llarg temps, sinó també es la primera investigació empírica que quantifica in situ la pressió de la pesca recreativa des de terra, el efecte directe sota les comunitats costeres de peixos, comparant aquesta amb la exercida per la pesca artesanal així com caracteritzar el perfil del pescador recreatiu. El

117


principal objectiu d’aquest capítol es per tant contribuir al coneixement sobre l’explotació dels recursos pesquers per part de la pesca recreativa i proporcionar dades essencials i necessàries per a la elaboració de estratègies de gestió integral de zones costeres, especialment en AMPs.

Per tal d’adreçar els objectius plantejats es va dur a terme en la Reserva Marina de Tabarca l’arreplegada de dades referents a la practica de la pesca recreativa, emprant una doble aproximació. Per una bia, per a tindre dades a llarg termini es varen emprar els part des de 1997 a 2004, facilitats per el servici de guarderia de la Reserva Marina. Així mateix, un intensiu mostreig del tipus “access point” es va dur a terme des de Setembre de 2003 fins a Maig de 2005.

L’ evolució a llarg termini de l'esforç de pesca recreativa en la Reserva Marina de Tabarca reflexa una tendència creixent en els darrers anys. En general, el major esforç de pesca es va registrar durant la temporada d'estiu en quasi tots els anys.

El pescador típic a la Reserva marina de Tabarca és un home de mitjana edat, resident de una àrea propera a la reserva que surt a pescar a soles o en parella. Molts pescadors desarrotllen la seva activitat de pesca durant tot l'any, però altres només en les estacions càlides, sobretot durant els dies de cap de setmana. La despesa mitjana per any i per pescador registrada durant el mostreig van ascendir a 902 €, principalment derivada de béns i serveis directament relacionats amb les seves

activitats pesqueres a la Reserva marina de Tabarca. Una alta proporció dels pescadors no tenen una llicència de pesca i alguns dels pescadors entrevistats estaven pescant il·legalment, perquè havien superat el nombre de línies permeses.

Segons els mostrejos in situ, el major esforç diari es va registrar durant els caps de setmana d'estiu, i el mínim esforç durant els dies d’entre setmana de la tardor, resultat en concordança amb les declaracions fetes pels pescadors durant les entrevistes. Un total de 1,376 peixos, corresponents a 28 especies es varen capturar durant el període de mostreig sent les espècies capturades més freqüentment: Diplodus spp, Symphodus tinca, Sparus aurata, Dicentrarchus labrax, Dentex dentex, iLabrus merula. El major valor de CPUE i BPUE es va detectar a l'estiu i el valor més baix a l'hivern. Es va estimar que l’activitat de pesca recreativa al AMP de Tabarca era capaç d’extreure a l’any 4,9 tones de peix. Les captures estimades d’espècies d’importància econòmica, com ara Dentex dentex, Muraena helena, Sciaena umbra, Diplodus spp i Sparus aurata, eren de proporcions considerables en comparació amb la captura per part de la pesca artesanal.

Tenint en compte les implicacions biològiques i socioeconòmiqes, resulta necessària l’aplicació d’una estratègia de gestió global que tinga en compte la pesca recreativa.

3. Evaluant l’impacte de la pesca recreativa al Parc Natural de la Serra Gelada (SW Mediterrani): un estudi de base per a la seva futura regulació

118


La pesca recreativa és una activitat d’oci molt important, que té efectes sobre les poblacions explotades molt similars als de la pesca comercial i que en alguns casos, pot arribar a generar esforç i captures iguals o superiors. Actualment, la pesca recreativa no està baix cap tipus de normativa al recent creat Parc Marí de la Serra Gelada i per tant, aquest estudi és una bona oportunitat per avaluar l'impacte associat a la pràctica de la pesca esportiva i fer el disseny de les futures estratègies de gestió a implantar. La pesca recreativa amb vaixell va ser estudiada utilitzant mitjans diferents de mostreig: reconeixements aeris, entrevistes directes als pescadors i fent ús de les dades disponibles de competicions de pesca esportiva. Les relacions dels pescadors esportius amb altres usuaris i altres aspectes sociològics es van estudiar també.

El pescador recreatiu des de embarcació típic del Parc marí de la Serra Gelada és un home de mitjana edat, resident local, que té una experiència de més de 21 anys i en la gran majoria no pertanyen a cap club de pesca. La pesca recreativa és una activitat d’oci intensament practicada durant tot l'any, ja que els pescadors pesquen una mitjana de 117,14 ± 12,5 dies any-1 i 4,91 ± 0,16 hores dia-1. La majoria d'ells coneixia l'existència de la figura de protecció del Parc Marí, encara que això no va tindre cap influència en la seva decisió de pescar-hi allà. Pel que fa a la relació entre els pescadors des d’embarcacions i altres usuaris, d'acord amb el seu propi punt de vista, el sector més conflictiu és l'aqüicultura. Un total de 33 espècies van ser declarades pels pescadors com

les més freqüentment capturades a l'àrea d'estudi, sent Seriola dumerili, Trachurus sp, Coryphaena hippurus, Pagellus erythrinus, Pagellus acarne, Scomber scombrus i Euthynnus alletteratus les pescades més a sovint.

El mostreig d’avioneta va mostrar que l’esforç pesquer des de embarcació és molt elevat en tot el parc i està centrat principalment en la zona d'amortiguació i la zona de especial protecció. En la zona d’amortiguació, tres àrees importants van ser identificades per la seva major densitat d'esforç, dos d'elles al voltant de les instal·lacions d’aqüicultura. Les dades de les competicions va mostrar una captura mitjana per unitat d'esforç de 0,465 ± 0,082 kg embarcacions-1 hora-1 i 0.323 ± 0.054 individus embarcacions-1 hora-1. La captura total anual per part de la pesca des d’embarcacions al Parc Natural de la Serra Gelada va ser estimada en 24,90 tones.

La importància de l'impacte de la pesca recreativa en els recursos pesquers i les seves implicacions sociològiques s’han posat de manifest en aquest estudi. La pesca recreativa des d’embarcació ha de ser tinguda en compte en l’elaboració de les estratègies de gestió, en aquesta i altres AMPs.

4. Impacte sobre el bentos del busseig recreatiu en una Àrea Marina Protegida Mediterrània

El suposat valor estètic de les AMPs i els avanços tècnics en el equipament de busseig han resultat en un increment del turisme de busseig cap a estes àrees marines protegides, en estos últims 20

119


anys (Hawkins i Roberts, 1992, Davis i Tisdell, 1995). El busseig pot resultar en el deteriorament de les comunitats bentòniques, perquè els bussos poden danyar fàcilment els organismes marins a través del contacte físic amb les seves mans, el cos, l'equip, o les aletes (Talgo, 1992; Rouphael i Inglis, 1995, 1997; Tracta'ls i Austin, 2001; Zakai i Chadwick- Furman, 2002; Pulfrich et al., 2003; Uyarra i Côté, 2007). Encara que els danys produïts per un únic bussejador a nivell individual son de poca importància, hi han evidències que els efectes acumulatius poden arribar a causar la destrucció d’organismes sensibles (Garrabou et al., 1998; Hawkins et al., 1999; Plathong et al., 2000 ). Hi ha un problema encara més greu quan l'activitat de busseig se centra en les zones marines protegides. Els efectes d'un gran nombre de bussejadors en uns pocs llocs d’una reserva marina poden ser contraris als objectius principals per als que es va crear l’AMP (Davis i Tisdell, 1995, 1996; Coma et al., 2004; Hawkins et al., 2005). No obstant això, alguns autors afirmen que l'impacte dels bussos en un lloc pot estar més influït per la seva experiència i pel seu comportament, més que pel nombre de persones que freqüenten el lloc (Davis i Tisdell, 1995; Rouphael i Inglis, 2001; Barker i Roberts, 2004 ).

Els objectius abordat en aquest capítol varen ser: estudiar el comportament del bussejadors sota l’aigua, evaluar els seus efectes sobre el medi marí i les característiques que poden influir en el comportament bussejador.

El present estudi es va dur a terme en el Parc Natural Marí de la Serra Gelada, el

qual ha vist incrementat el seu nombre anual de bussejadors des de la seva declaració com AMP. En total, 181 bussejadors recreatius van ser seguits sota l’aigua, i tots els contactes i els efectes produïts durant la immersió van ser registrats. Tota la informació sobre el perfil del bussejador i les característiques de la immersió es va registrar mitjançant una enquesta personal feta després del seguiment . Els resultats derivats del mostreig van revelar que 175 dels bussos observats (96,7%) van cometre almenys un contacte amb el fons marí, amb una mitjana de contactes de 41,20 ± 3,55 (mitjana ± es) per bussejador per 10 min.

L'aleteig és el tipus de contacte més freqüent i el dany principal derivat d'aquesta acció era la resuspensió de sediments. El nombre de contactes amb el fons va ser més elevat per a homes que per a les dones, per als bussos inexperts que per als bussejadors experimentats i per als usuaris de càmera de fotos o llanterna que per als no usuaris. Així mateix, els bussos no acompanyats per un líder de busseig o que no havien estat alertats d'evitar el contacte del fons marí mitjançant un “briefing” previ, ocasionaren més contactes amb el fons marí que els que anaven acompanyant o havien assistit prèviament a la immersió al “briefing”.

Un major coneixement de les causes que determinen una conducta perjudicials en el bussejadors, pot ser útil alhora de gestionar l’activitat de busseig, reduint el dany produït pels usuaris i assegurant la sostenibilitat d'aquesta activitat en les àrees marines

120


protegides.

En el Mar Mediterrani, el sistema litoral de profunditat en zones rocoses està dominat per organismes calcaris, animals sèssils filtradors i esquelets morts que es coneixen com a "coralígen" (Ballesteros, 2006). El coralígen és un exemple d'una comunitat molt diversificada i altament estructurada de suspensivors, una de les més riques en termes de nombre de especies i organització del bentos de la Mediterrània (Ros et al., 1985; Ballesteros i Ros, 1989). Aquesta gran diversitat d'organismes i vistositat d'aquests hàbitats els fan molt atractius per als bussejadors (Lloret et al., 2006).

El busseig recreatiu és un component important i creixent del mercat turístic internacional (Tabata, 1992; Davis i Tisdell, 1995). En general, aquesta activitat es produeix normalment en llocs marins amb un paisatge subaquàtic impressionant, alta diversitat biològica (Tabata, 1989) o de especial valor biològic, fets que en molts casos coincideixen en les àrees protegides (Davis i Tisdell, 1995). Tanmateix, el busseig pot resultar en la deterioració dels parcs marins, perquè els bussos poden fer malbé directament els organismes marins (per exemple, pel contacte físic amb alguna part del seu cos o equipament) o indirectament (per exemple, augmentant la resuspensió de sediments) (Rouphael i Inglis, 1997; Tratalols i Austin, 2001; Zakai i Chadwick Furman, 2002; Pulfrich et al.,

2003; Luna et al., 2009).

Alguns estudis de seguiment de l'activitat de busseig han demostrat que l’augment de sediments va ser el efecte més freqüentment produït pels bussos, seguit pel contacte amb els organismes fràgils, com les gorgònies, coralls i briozous (Garrabou et al., 1998; Zakai i Chadwick-Furnman, 2002; Barker i Roberts, 2004; Luna et al., 2009).

Entre altres, hi ha tres mecanismes principals d'elevar els sediments que puguin afectar directament organismes de costes rocoses: el soterrament o l'asfíxia, erosió o abrasió produïda pel movent dels sediments i els canvis en les característiques físiques de l'estabilitat del substrat dur que pot resultar en una pèrdua del hàbitat idoni (Airoldi, 1998). La resuspensió de sediment influeix clarament en diverses espècies d'invertebrats sèssils, afectant al creixement i la supervivència de espècies d’ ascídies, que es veuen afectades per la deposició excessiva de sediments que acaba provocant la sepultura i l'obstrucció dels sifons i l’aparell branquial. Els sediments també produeix efectes d'abrasió i dificulta la solució dels processos (Bakus, 1968; Gabriele et al., 1999).

D'altra banda, els bussejadors poden danyar fàcilment els organismes marins a través del contacte físic amb les seves mans, el cos, l'equip o les aletes (Rouphael i Inglis, 1997; Tractalos i Austin, 2001; Zakai i Chadwick Furman, 2002; Pulfrich et al., 2003 ; Uyarra i Côté, 2007) causant trencament, fragmentació o l'abrasió dels teixits (Hawkins i Roberts, 1992;

121

5. Halocynthia papillosa (Linnaeus, 1767) com indicador de l’impacte del busseig


Garrabou et al., 1998; Coma et al., 2004).

La combinació de dos efectes, la resuspensió dels sediments i el contacte físic, son els principals agents estressant als que les comunitats bentòniques es veuen sotmeses per part de l'activitat de busseig.

El gran nombre d'espècies i l'alt nivell de diversitat d'espècies en la comunitat de coral·ligen dificulten l'avaluació de l'impacte de busseig a la comunitat en conjunt, per tant, la selecció d'una espècie o grup d’espècies indicadores seria una avantatjosa ferramenta (Sala et al., 1996 ).

En aquesta part del treball, plantegem la hipòtesi de que un augment del sediment resuspès per els bussos tindrà un efecte negatiu sobre l'adquisició d'energia i el desenvolupament de l’ascídia solitaria Halocynthia papillosa. Aquesta espècie d’ascídia ja ha estat reconeguda com a indicadora de ambients poc perturbats (Naranjo et al., 1996). A més, l'efecte abrasiu del sediment en el teixit de ascídies i el contacte físic directe produït pels bussos poden produir efectes additius negatius. Aquesta espècie podria ser un indicador útil per avaluar i supervisar l'efecte de busseig a la comunitat mediterrània del coral·ligen. L'objectiu d'aquest estudi va ser avaluar la idoneïtat d’ H. papillosacom un indicador d'impacte de busseig i el seu paper potencial com una eina eficaç en la detecció de pertorbacions en el sistema ecològic. Per assolir aquest objectiu principal, els objectius secundaris han estat: (1) quantificar l'abundància d' H. papillosa entre els

llocs freqüentats pels bussejadors i aquells poc freqüentats, (2) estudiar la seva distribució de talles entre aquestos llocs, i (3) estudiar el seu grau d'exposició sobre el substrat.

Un estudi comparatiu de les poblacions d'H papillosa es va dur a terme durant un període de dos anys (durant 2006 i 2007), abans i després de la temporada alta de busseig al Parc Marí de Serra Gelada (SW Mediterrània occidental). El disseny emprat per a assolir els objectius plantejats va ser un disseny de Mesures Repetides amb quatre factors, dos “inter-grups” i dos més “intra.grups”. Les dades es varen arreplegar mitjançat un mostreig in situ no destructiu, en el que es varen registrar i mesurar tots el individus existents en cada parcel·la.

El resultats obtinguts assenyalen individus més gran i més abundant d’Hpapillosa en els llocs poc freqüentats per bussejadors que en els llocs més freqüentats durant el període d'estudi. D'altra banda, els individus d'Hpapillosa en aquestos llocs més freqüentats ocupen posicions més críptiques que en les localitats poc freqüentades.

H. papillosa ha demostrat ser molt sensible als efectes adversos de busseig. Aquesta espècie podria constituir un bioindicador fiable de l'activitat de busseig i com a tal podria ser una eina útil per al control ràpid i fàcil de l’impacte del busseig sobre les comunitats de coral·ligen.

122


H.papillosa ha estat classificada com una espècie sensible a l'estrès, així com una bona indicadora de condicions naturals en el medi marí (Naranjo et al., 1996). Un recent estudi de monitorització va demostrar que en llocs amb alta freqüència de busseig, l'abundància d'H papillosa és molt baixa i es troba restringida a ocupar posicions més críptiques que en zones on no es busseja (Lluna et al, 2010). Aquest estudi que correspon al capítol anterior va avaluar la sostenibilitat de H. papilosla i va proposar l'espècie com un indicador adequat per avaluar i supervisar l'efecte de busseig a la comunitats de coral·ligen Mediterràni. Tanmateix, la causa-efecte no hi es va corroborar mitjançat aquest estudi pasat, així com tampoc es va estudiar l'efecte immediat i els mecanismes pels quals els individus H. papillosa es veuen afectades negativament pel busseig. L'objectiu d’aquesta part de la tesi va ser determinar la causa-efecte mitjançant un experiment, per tal de llançar llum sobre les respostes immediates i a llarg termini de H. papillosa front la pertorbació causada pel busseig.

Per tal de complir amb aquest objectiu, un experiment de manipulació in situ es va dur a terme per estudiar durant un període d’un any. Es van considerar cinc intensitats d'impacte d’aleteig experimental: 0 (control), 3, 90, 270 i 330 aletejos. Dues unitats experimentals, cadascuna amb deu individus d’H.papillosa, van ser

assignats a l'atzar per a cada tractament. En cada unitat experimental, deu trampes de sediment circular es van col·locar. 24 h hores després del aleteig experimental les trampes es varen recol·lectar per al seu posterior anàlisi. L'abundància total i la supervivència dels individus de cada talla en cada unitat experimental es van mesurar durant vuit intervals de temps: abans de l'experiment i 24 hores, 1 setmana i 1, 3, 6, 9 i 12 mesos després. El disseny de l’experiment corresponia a un disseny de mesures repetides que va ser analitzat mitjançant un RM-ANOVA, per tal de provar les diferències entre els tractaments i al llarg del temps que va durar l’experiment.

En els tres llocs més afectats per el aleteig experimental es va observar una reducció en el nombre d'individus, amb taxes de letalitat més altes en aquells individus més petits. D'altra banda, en els dos llocs més afectats pel aleteig, els individus supervivents no van ser capaços de créixer. Es va observar una elevada correlació negativa entre l'abundància d'individus i el nombre d'aletejos (R2 = 0,96). El model de regressió general reflexa una alta correlació entre l'abundància dels individus i les fraccions dels sediments que es varen estudiar. Aquestos resultats indiquen que la resuspensió dels sediments produïts per aleteig podria explicar la disminució de l'abundància d’H. papillosa que s'ha observat en les zones on hi ha una alta intensitat de busseig.

A la llum dels resultats, H. papillosa es proposa com una espècie betònica del coral·ligen molt adequada en l'estudi de

123

6. Halocynthia papillosa com indicador de l’impacte del busseig: un experiment in situ


l'impacte de busseig a la Mediterrània, que també seria un bon indicador per a altres tipus d'efectes que produeixen una quantitat de sediment a l'aigua, com ara el dumping o el dragatge.

124

BYBLIOGRAPHY

Abdulla, A., Gomei, M., Maison, E. and Piante, C. 2008. Status of Marine Portected Areas in the Mediterranean Sea. UICN, Malaga and WWF, France. pp 152.

Agardy, T. 2003. Special Feature: Innovation and MPAs in the Mediterranean Sea. MPA News, 5(3), 2-4.

Agardy, T. 2005. Global marine conservation policy versus site-level implementation: the mismatch of scale and its implications. Marine Ecology Progress Series, 300, 242- 248.

Airoldi, L. 1998. Roles of disturbance, sediment stress, and substratum retention on spatial dominance in algal turf. Ecology, 79(8), 2759–2770.

Airoldi, L., 2003. The effects of sedimentation on rocky coast assemblages. Oceanography and Marine Biology: An Annual Review, 41, 161–236.

Ajzen, I. 1991. The theory of planned behavior. Organizational Behaviour and Human Decission Processes, 50, 179-211.

Allison, W.R. 1996. Snorkeler damage to reef corals in the Maldive Islands. Coral Reefs, 15, 15-18.

Anderson, M.J. 2001. Permutation tests for univariate or multivariate analysis of variance and regression. Canadian Journal of Fisheries and Aquatic Sciences, 58, 626-639.

Ardizzone, G. D. and Pelusi, P. 1984. Yield and damage evaluation of bottom trawling on Posidonia meadows. In: Boudouresque, C. F., Jeudy de Grissac, A., and Olivier, J (Eds), International Workshop on Posidonia oceanica beds (vol I), Porquerolles, France. pp. 255-259.

Arin, T. and Kramer, R.A. 2002. Divers'willingness to pay to visit marine sanctuaries an exploratory study. Ocean and Coastal Management, 45, 171-183.

Arlinghaus, R. 2005. A conceptual framework to identify and understand conflics in recreational fisheries systems, with implications for sustainable management. Aquatic Resources, Culture and Development, 1(2), 145-174.

Arlinghaus, R. 2006. On the Apparently Striking Disconnect between Motivation and Satisfaction in Recreational Fishing: the Case of Catch Orientation of German Anglers. North American Journal of Fisheries Management, 26(3), 592-605.

Arlinghaus, R. and Mehner, T. 2004. A management orientated comparative analysis of urban and rural anglers living in a metropolis (Berlin, Germany). Environmental Management, 33, 331-344.

Arlinghaus, R., Mehner, T. and Cowx, I.G. 2002. Reconcilling traditional inland fisheries management and sustainability in industrialized countries, with emphasis on Europe. Fish and Fisheries, 3, 261-316.

127

____________________________________________________________Bibliography

Armsworthy, S.L., 1996. Effects of suspended bottom sediment on the feeding activity of the ascidian Halocynthia pyriformis (Rathke). University of New Brunswick, Saint John Press, New Brunswick, Canada.

Armsworthy, S.L., MacDonald, B.A. and Ward, J.E. 2001. Feeding activity, absorption efficiency and suspension feeding processes in the ascidian, Halocynthia pyriformis (Stolidobranchia: Ascidiacea): responses to variations in diet quantity and quality. Journal of Experimental Marine Biology and Ecology, 260, 41-69.

Asafu-Adjaye, J. and Tapsuwan, S. 2008. A contingent valuation study of SCUBA diving benefits: Case study in Mu Ko Similan Marine National Park, Thailand. Tourism Management, 29(6), 1122-1130.

Attwood, C.G., Mann, B.Q., Beaumont, J. and Harris, J.M. 1997. Review of the state of marine protected areas in South Africa. South African Journal of Marine Sciences, 18, 341–367.

Ayvazian S. G., Wise B. S., Young G. C. Short-term hooking mortality of tailor (Pomatomus saltatrix) in Western Australia and the impact on yield per recruit. Fisheries Research (2002) 58:241–248.

Badalamenti, F., Ramos, A., Voultsiadou, E., Sanchez-Lizaso, J.L., D’Anna, G., Pipitone, G., Mas, J., Ruiz Fernandez, J.A., Whithmarsh, D. and Riggio, S. 2000. Cultural and socio-economic impacts of Mediterranean

marine protected areas. Environmental Conservation, 27(2), 1-16.

Baker, D.L. and Pierce, B.E. 1997. Does fisheries management reflect societal values? Contingent valuation evidence for the River Murray. Fisheries Management and Ecology, 4, 1-15.

Bakus, G.J. 1968. Sedimentation and benthic invertebrates of Fanning Island, Central Pacific. Marine Geology, 6, 45–51.

Balata, D., Piazzi, L., Cecchi, E. and Cinelli, F. 2005. Variability of Mediterranean coralligenous assemblages subject to local variation in sediment deposits. Marine Environmental Research, 60, 403-421.

Ballantine, W.J. 1995. Networks of “no take” marine reserves are practical and necessary. In: N. L. Shackell, and J. H. Martin Willison (Eds.), Marine Protected Areas and Suitable Fisheries. Science and Management of Protected Areas Association, Wolfville, NS, Canada. pp. 13-20.

Ballesteros, E. 2006. Mediterranean coralligenous assemblages: a synthesis of present knowledge. Oceanography and Marine Biology: An Annual Review, 44, 123-195.

Ballesteros, E., Ros, J.D. 1989. Els ecosistemes bentònics, In: Terradas, J., Prat, N., Escarré, A. and Margalef, R. (Eds.), Historia Natural dels països catalans. Enciclopédia Catalana. Institute d’Estudis Catalans, Barcelona, Vol 14, pp. 119-176.

128

Antrhopic impacts in Mediterranean MPAs___________________________________

Barker, N.H.L. and Roberts, C.M. 2004. Scuba diver behaviour and the management of diving impacts on coral reefs. Biological Conservation, 120, 481-489.

Bates, B.C., Kundzewicz, Z.W., Wu, S. and Palutikof, J.P. 2008. Climate Change and Water. Technical Paper of the Intergovernmental Panel on Climate Change, IPCC Secretariat, Geneva.

Beal, R.E., Desfosse, J.C., Field, J.D.and Schick, A.M. 1998. Review of Interstate Fishery Management Plans, Atlantic States Marine Fisheries Council, Washington, D.C.

Becerro, M.A. and Turon, X. 1992. Reproductive cycles of the ascidians Microcosmus sabatieri and Halocynthiapapillosa in the Nortwestern Mediterranean. PSZN I: Marine Ecology, 13, 363-373.

Beger, M., Jones, G.P. and Munday, P.L. 2003. Conservation of coral reef biodiversity: a comparison of reserve selection procedures for corals and fishes. Biological Conservation, 111, 53–62.

Bell, J.D. and Harmelin, M.L. 1983. Fish fauna of French Mediterranean Posidonia oceanica meadows. II: feeding habits. Tethys, 11(1): 1-14.

Bennett, B.A. and Atwood, C.G. 1993. Shore-angling catches in the De Hoop Nature Reserve, South Africa, and further evidence for the protective value of marine reserves. South African Journal of Marine Science, 13, 213–222.

Bennett, B.A., and Attwood, C.G. 1991. Evidence for recovery of a surf zone fish assemblage following the establishment of a marine reserve on the southern coast of South Africa. Marine Ecology Progress Series, 75(2–3), 173–181.

Bennett, E.M., Carpenter, S.R., Peterson, G.D., Cumming, G.S., Zurek, M. and Pingali, P. 2003. Why global scenarios need ecology. Frontiers in Ecology and the Environment: 1(6), 322-329.

Béthoux J.P., Gentili B. and Tailliez D. 1998. Warming and freshwater budget change in the Mediterranean since the 1940s, their possible relation to the greenhouse effect. Geophysical Research Letters, 25(7) 1023-1026.

Bianchi C.N. 2007. Biodiversity issues for the forthcoming tropical Mediterranean Sea.

Bianchi, C.N. 2001. La biocostruzione negli ecosistemi marini e la biologia marina italiana. Biologie Marine Mediterranea. 8, 112–130.

Blanc, J.J. 1974. La sédimentation sur le précontinent de Provence et ses modalités. Comparaisons avec d'autres secteurs de la Méditerranée. Bulletin Bureau il Rechourses Géologiques e minere, 2 (4-3), 133-164.

Blue Plan. 2005. A Sustainable Future for the Mediterranean. The Blue Plan’s Environment and Development Outlook. UNEP-MAP Regional Ativity Centre, Sophia Antipolis, Frnace. 33 pp.

Bohn, J. and Roth. E. 1997. Survey on angling in Denmark 1997 – Results and

129

____________________________________________________________Bibliography

Comments. In: Toivonen, A.L. Tuumaimem, P. (Eds.). Socio-Economics of Recreational Fishery.Nordic Council of Ministers, Temanord, Copenhagen. Vol. 604: 79-88

Boudouresque C. F. and Ribera, M. A. 1993. Les espèces et les espaces protégés marins en Méditerranée. Situation actuelle, problèmes et priorités. Actes du coldogue international «Les zones protégëes en Méditerranéen, Tunis, 25-27 nov. 1993. Centre d’Etude, de Recherches et de Publications (C.E.R.P.) et Comité pour Les Etudes Mediterranéene (C.E.M.), Tunis.

Boudouresque C.F. and Verlaque M. 2002. Biological pollution in the Mediterranean Sea: invasive versus introduced macrophytes. Marine Pollution Bulletin, 44, 32–38.

Boudouresque, C. F. and Jeudy de Grissac, A. 1983. L'herbier à Posidoniaoceanica en Méditerranée: les interactions entre le plant et le sédiment. Journal de Recherche Océanographique, 8 (2–3), 99–122.

Boudouresque, C.F. and Ribera, M.A. 1995. Les espèces et les espaces protégés marins en Mediterranée, situation actuelle, problèmes et prioritès. In: Actes Colloque Les Zones Protegées en Méditerranée : Espaces, Espèces et Instruments d’Application des Conventions et Protocoles de la Méditerranée, Tunis, pp. 93–141. Centre d’Etude, de Recherches et de Publications (C.E.R.P.) et Comité pour Les Etudes Mediterranéene (C.E.M.), Tunis.

Brenchin, S.R., Wilshusen, P.R., Fortwangler, C.L., and West, P.C. 2002. Beyond the square wheel: towards a more comprehensive understanding of biodiversity conservation as social and political process. Society and Natural Resources, 15, 41-64.

Briand, F., Giuliano, L. 2007. CIESM Contribution to the Green Paper on EU Maritime Policy Priorities for Marine Research and Policy in the Mediterranean Sea—A Multilateral View. Web: http://www.ciesm.org/news/policy

Broadhurst, M. K., Larsen, R. B., Kennelly, S. J. and McShane, P. 1999. Use and success of composite square-mesh codends in reducing bycatch and in improving size-selectivity of prawns in Gulf St. Vincent, South Australia. Fisheries Bulletin, 97, 434–48.

Broadhurst, M.K., Gray, C.A., Reid, D.D., Wooden, M.E.L., Young, D.J., Haddy, J.A. and Damiano, C. 2005. Mortality of key fish species released by recreational anglers in an Australian estuary. Journal of Experimental Marine Biology and Ecology, 321, 171–179.

Broderick, A.C., Glen, F., Godley, B.J. and Hays, G.C. 2002. Estimating the number of green and loggerhead turtles nesting annually in the Mediterranean. Oryx 36(3), 227-235.

Brosnan, D.B. and Crumrine, L.L. 1994. Effects of human trampling on marine rocky shore communities. Journalof ExperimentalMarine Biologyand Ecology, 177 (1), 79-97.

130


Brown, P.J. and Taylor, R.B. 1999. Effects of trampling by humans on animals inhabiting coralline algal turf in the rocky intertidal. Journal of Experimental Marine Biology and Ecology, 235(1), 45-53.

Bustos, S. and Frid, C. 2003. Using indicator species to assess the state of macrobenthic communities. Hydrobiologia, 496, 299-309.

Calvín-Calvo, J.C. 2000. El ecosistema Marino Mediterráneo: guía de flora y fauna, third ed. Edit. Calvín Calvo, Madrid.

Calvisi, G., Moretto, S., Pessani, D. and Odorico, R. 2003. Tipologia del turista subacqueo nella Riserva di Miramare come contributo per la corretta gestione. Sommario ed XXXIV Congresso della Società Italiana di Biologia Marina, Port El Kantaoui, 31 maggio-7 giugno 2003, Tunisa.

Capellà, J., Donaire, J.A., Muñoz, J.C. and Ullastres, H. 1998. Turisme Sostenible a la Mediterània: Guia per a la Gestió Local. Brau Edicions, Barcelona, 156 pp.

Carlisle, D.B. 1966. The ciliary current of Phallusiaw Ascidiaceaxand the squirting of sea squirts. Journal of the Marine Biological Association of the United Kingdom, 46, 125–127.

Casu, D., Ceccherelli, G., Curini-Galletti, M. and Castelli, A. 2006. Human exclusion from rocky shores in the Mediterranean marine protected area (MPA): An opportunity to investigate the effects of trampling. Marine Environmental Research, 62, 15-32.

Ceccherelli, G., Campo, D. and Milazzo, M. 2007. Short-term response of the slow growing seagrass Posidonia oceanica to simulated anchor impact. Marine Environmental Research, 63(4),341-9.

Cerrano, C., Bavestrello, G., Bianchi, C.N., Cattaneo-Vietti, R., Bava, S., Morganti, C., Morri, C., et al. 2000. A catastrophic mass-mortality episode of gorgonians and other organisms in the Ligurian Sea (North-western Mediterranean), summer 1999. Ecology Letters, 3, 284–293.

Chadwick-Furman, N.E. 1997. Effects of SCUBA diving on coral reef invertebrates in the US Virgin Islands: implications for the management of diving tourism. In: den Hartog, J. C., van Ofwegen, L. P. and van der Spoel, S (Eds.). Proceedings of the 6th International Conference on Coelenterate Biology, National Naturhistorisch Museum, Leiden, the Netherlands. pp. 91–100.

Cheung, W.W.L., Pitcher, T.J. and Pauly, D. 2005. A fuzzy logic expert system to estimate intrinsic extinction vulnerabilities of marine fishes to fishing. Biological Conservation, 124, 97–111.

Cheung, W.W.L., Watson, R., Morato, T., Pitcher, T.J. nad Pauly, D. 2007. Intrinsic vulnerability in the global fish catch. Marine Ecology Progress Series, 333, 1–12.

Chipman, B. D. and Helfrich, L. A. 1988. Recreational Specializations and Motivations of Virginia River Anglers.

131

____________________________________________________________Bibliography

North American Journal of Fisheries Management, 8, 390-398.

Christie, P. 2004. Marine protected areas as biological successes and social failures in Southeast Asia. American Fisheries Society Symposium, 42,155–164.

Christie, P., White, A. T., Stockwell, B. and Jadloc, C. R. 2003. Factors influencing integrated coastal management sustainability: focus on environmental conditions in two locations in t the Philippines. Silliman Journal, 44(1),286–323.

CIESM. 2002. Alien marine organisms introduced by ships in the Mediterranean and Black seas. In: CIESM Workshop Monographs. N. 20.

CIESM. 2004. Atlas of Exotic Species in the Mediterranean - Vol.3. Molluscs by Zenetos, A., Gofas, S., Russo, G. and Templado, J.

Clarke, K.R. and Warwick, R.M. 2001. Changes in marine communities: an approach to statistical analysis and interpretation, 2nd ed. Plymouth: PRIMER-E, 172 pp.

Clarke, K.R., 1993. Non-parametric multivariate analyses of changes in community structure. Aust. J. Ecol. 18, 117-143.

Clarke, N. and Stankey, G.H. 1979. Determining the acceptability of recreation impacts: An application of the Outdoor Recreation Opportunity Spectrum. Proceedings of the Wildland Recreation Impacts Conference,

October 27-29, 1978, Seattle, Washington DC.

Claudet, J., Osenberg, C.W., Benedetti-Cecchi, L., Domenici, P., García-Charton, J.A., Pérez-Ruzafa, Á., Badalamenti, F., Bayle-Sempere, J., Brito, A., Bulleri, F., Culioli, J.M., Dimech, M., Falcón, J.M., Guala, I., Milazzo, M., Sánchez-Meca, J., Somerfield, P.J., Stobart, B., Vandeperre, F., Valle, C., Planes, S. 2008. Marine reserves: Size and age do matter. Ecology Letters, 11(5), 481-489.

Cloern, J.E. 1982. Does the benthos control phytoplankton biomass in South San Francisco Bay? Marine Ecology Progress Series, 9, 191-202.

Cochran, W.G. 1951. Testing a linear relation among variances. Biometrics, 7, 17-32.

Coleman, F., Figueira, W.F., Ueland, J.S. and Crowder, L.B. 2004. The impact of United States recreational fisheries on marine fish populations. Science, 305, 1958-1989.

Coll, J., Garcia-Rubies, A., Moranta, J., Stefanni, S. and Morales-Nin, B. 1999. Efectes de la prohibició de la pesca esportiva de l’anfós a Cabrera. Bolletí de la Societat d’Història Natural de les Illes Balears, 42, 125-138.

Coll, J., Linde, M., García-Rubies, J., Riera, F. and Grau, A.M. 2004. Spearfishing in the Balearic Islands (west central Mediterranean): species affected and catch evolution during the period 1975–2001. Fisheries Research, 70, 97–111.

132


COM, 2002. 535 final. Communication from the Commission to the Council and the European Parliament laying down a Community Action Plan for the conservation and sustainable exploitation of fisheries resources in the Mediterranean Sea under the Common Fisheries. Commission of the European Communities, Brussels. 37 pp.

COM, 2002. 535 final. Communication from the Commission to the Council and the European Parliament laying down a Community Action Plan for the conservation and sustainable exploitation of fisheries resources in the Mediterranean Sea under the Common Fisheries. Commission of the European Communities, Brussels. 37 pp.

Coma, R., Pola, E., Ribes, M. and Zabala, M. 2004. Long-term assessment of temperate octocoral mortality patterns, protected vs. unprotected areas. Ecological Applications, 14, 1466–1478.

community of the Lebanese rocky coast (eastern Mediterranean Sea) with emphasis on Red Sea immigrants. Biological Invasions, 7(4), 625-637.

Connelly, N.A., Brown, T.L. 2000. Options for maintaining high fishing satisfaction in situations of declining catch rates. Human Dimensions of Wildlife, 5, 18-31.

Consellería de Territori i Habitatge, 2005. Plan de Ordenación de los Recursos Naturales de la Serra Gelada y su zona litoral. Memoria Descriptiva y Justificativa. Generalitat Valenciana, Valencia, pp 25-47.

Constanza, R., Arge, R., de Groot, R., Farber, S., Grasso, M., Hannon, B., Limburg, K., Naeem, S., O’Neill, R. V., Paruelo, J., Raskin, R. G., Sutton, P. and van den Belt, M. 1997. The value of the world’s ecosystem services and natural capital. Nature, 387, 253-260.

Cooke, S.J. and Cowx, I.G. 2004. The role of recreational fishing in global fish crises, Bioscience 54, 857–859

Cooke, S.J. and Cowx, I.G. 2006. Contrasting recreational and commercial fishing: searching for common issues to promote unified conservation of fisheries resources and aquatic environments, Biological Conservation, 128, 93–108.

Cooke, S.J., Danylchuck, A.J., Danylchuck, S.E., Suski, C.D. and Godberg, T.L. 2006. Is catch and release recreational angling compatible with no-take marine protected areas? Ocean Coastal Management, 49(5/6), 342-354.

Cooke, S.J., Suski, C.D., Barthel, B.L., Ostrand, K.G., Tufts, B.L. and Philipp, D.P. 2003. Injury and mortality induced by four hook types on bluegill and pumpkinseed. North American Journal of Fisheries Management, 23, 883-893.

Cowx, I.G. 2002. Analysis of threats to freshwater fish conservation: past and present challenges. In: Collares-Pereira, M.J., Cowx, I.G. and Coelho M.M. (Eds.). Conservation of freshwater fishes: options for the future. UK: Blackwell Scientific Press, pp 201–220.

Cowx, I.G. 2002. Recreational fishing. In: Hart, P. and Reynolds, J.D. (Eds.).

133

____________________________________________________________Bibliography

Handbook of Fish Biology and Fisheries vol. II, Blackwell Science, Oxford. pp. 367–390.

Cowx, I.G., Van Knaap, M., Van Der Muhoozi, L. and Othina, A. 2003. Improving fishery catch statstics for Lake Victoria. Aquatic Ecosystem Health Management, 6, 299–310.

Cox, S. and Walters, C. 2002. Maintaning quality in recreational fisheries: How success breeds failure in management of open-access sport fisheries. In: Pitcher TJ, and Hollingrowth, C.E. (Eds.). Recreational fisheries: ecological, economic and social evaluation, pp 107-119.

Creed, J.C. and Amado-Filho, G.M. 1999. Disturbance and recovery of the microflora of a seagrass (Halodule wrightii Acherson) meadow in the Abrolhos Marine National Park, Brazil: an experimental evaluation of anchor damage. Journal Experimental Marine Biology and Ecology, 235, 285-306.

Cury, P., Mullon, C., Garcia, S.M. and Shannon, L.J. 2005. Viability theory for an ecosystem approach to fisheries. ICES Journal Marine Science, 62, 577-584.

Cury, P.M. and Christensen, V. 2005. Quantitative ecosystem indicators for fisheries management: an introduction. ICES Journal of Marine Science, 62, 307-310.

Daily, G. C., Alexander, S., Ehrlich, P. R., Gouler, L., Lubchenco, J., Matson, P. A., Mooney, H. A., Postel, S., Schneider, S. H., Tilman, D. and Woodwell, G. M. 1997. Ecosystem

services: benefits supplied to human societies by natural ecosystems. Issues in Ecology, 2,1–18.

Davenport, J. and Davenport, J.L. 2006. The impact of tourism and personal leisure transport on coastal environments: A review. Estuarine, Coastal and Shelf Science, 67(1), 280-292.

Davis, D. and Tisdell, C. 1995. Recreational scuba-diving and carrying capacity in marine protected areas. Oceans and Coastal Management, 26, 19-40.

Davis, D. and Tisdell, C. 1996. Economic management of recreational scuba diving and the environment. Journal of Environmental Management, 48, 229–248.

Davis, G.E. 1977. Anchor damage to coral reef on the coast of Florida. Biological Conservation, 11, 29–34.

Davis, M. W., Olla, B. L. and Schreck, C. B. 2001. Stress induced by hooking, net towing, elevated sea water temperature and air in sablefish: lack of concordance between mortality and physiological measures of stress. Journal of Fish Biology, 58,1–15.

De Roos, A.M. and Person, L. 2002. Size-dependent life-history traits promote catastrophic collapses of top predators. Proceedings of the national Academy of Sciences of the United States of America, 99 (Suppl. 20), 12907-12912.

Dearden, P., Bennett, M. and Rollins, R. 2006. Implications for coral reef

134


conservation of diver specialization. Environmental Conservation, 33 (4), 353–363.

Delbaere, B. 1998. Facts and Figures on Europe’s biodiversity - state and trends 1998- 1999, Technical Report Series. European Centre for Nature Conservation, Tilburg.

den Hartog, C. 1970. The seagrasses of the world. Verhandelingen der Koninklijke Nederlandse Akademie van Wetenschappen Afdeling Natuurjunde 59: 1-275.

Denny, C.M. and Babcock, R.C. 2004 Do partial marine reserves protect reef fish assemblages? Biological Conservation, 116(1), 119-129.

Depondt, F., and Green, E. 2006. Diving user fees and the financial sustainability of marine protected areas: Opportunities and impediments. Ocean and Coastal Management, 49, 188-202.

di Franco, F., Bussotti, S., Navone, A., Panzalis, P., Guidetti, P. 2009. Evaluating effects of total and partial restrictions to fishing on Mediterranean rocky-reef fish assemblages. Marine Ecology Progress Series, 387, 275-285.

Diaz-Almela, E., Marbà, N. and Duarte, C.M. 2007. Consequences of Mediterranean warming events in seagrass (Posidonia oceanica) flowering records. Global Change Biology, 13, 224–235

Diggles, B.K. and Ernst, I. 1997. Hooking mortality of two species of shallow-water reef fish caught by

recreational angling methods. Marine and Freshwater Research, 48, 479–483.

Dignam, D. 1990. Scuba gaining among mainstream travellers. Tourism and Travel News, 26, 44–45.

Dinsdale, E.A. and Harriott, V.J. 2004. Assessing anchor damage on coral reefs: A case study in selection of environmental indicators. Environmental Management, 33 (1), 126-139.

Dixon, J.A., Scura, L.F. and Van’t Hof, T. 1993. Meeting ecological and economic goals: marine parks in the Caribbean. Ambio, 22, 117-125.

Done, T.J. 1995. Ecological criteria for evaluating coral reefs and their implications for managers and researchers. Coral Reefs, 14, 183-92.

Duffy, J.E. 2006. Biodiversity and the functioning of seagrass ecosystems. Marine Ecology Progress Series, 311, 233-250.

Dunn, M., Potten, S., Radford, A. and Whitmarsh D. 1989. An economic appraisal of the fishery for bass Dicentrarchus labrax in England and Wales. Report of the centre for economics and management of Aquatic Resources (CEMARE), University of Portsmouth, UK, No. 14, 720pp.

Eckrich, C.E. and Holmquist, J.G. 2000. Trampling in a seagrass assemblage: direct eVects, response of associated fauna, and the role of substrate characteristics. Marine Ecology Progress Series, 201, 199-209.

135

____________________________________________________________Bibliography

Edinger, E.N. and Risk, M.J. 2000. Reef classification by coral morphology predicts coral reef conservation value. Biological Conservation, 92, 1-13.

EU, 2004. Mediterranean: Guaranteeing Sustainable Fisheries. Fishing in Europe, vol. 21, 12 pp.

European Comission. 2008. Eurostat regional yearbook 2008. Office for Publications of the European Communities. 187 pp.

European Parliament, 2009. European Parliament legislative resolution of 22 April 2009 on the proposal for a Council regulation establishing a Community control system for ensuring compliance with the rules of the Common Fisheries Policy (T6-0255/2009). Available from: http://www.europarl.europa.eu.

Fabinyi, M. 2008. Dive tourism, fishing and marine protected areas in the Calamianes Islands, Philippines. Marine Policy, 32(6), 898-904.

FAO. 2008. web: http://www.fao.org/fishery/ccrf/en

FAO. 2009. web: http://www.fao.org/news/story/en/item/10270/icode/

Farrugio, H., Olivier, P., and Biagi, F. 1993. An overview of the history, knowledge, recent and future research trends in Mediterranean fisheries. Scientia Marina, 57, 105–119.

Feathers, M.G. and Enable, A.E. 1983. Effects of depressurization upon

largemouth bass. North American Journal Fisheries Management 3, 86–90

Fedler, AJ. and Ditton, R.B. 1994. Understanding angler motivations in fisheries Management. Fisheries, 19(4), 6–13.

Fernandez-Jover, D., Sanchez-Jerez, P., Bayle-Sempere, J., Valle, C. and Dempster, T. 2008. Seasonal patterns and diets of wild fish assemblages associated with Mediterranean coastal fish farms. ICES Journal of Marine Science, 65, 1153-1160.

Ferrer-Montaño, J.O.¸ Dibble, E.D., Jackson, D.C. and Rundle, K.R. 2005. Angling assessment of the fisheries of Humacao Natural Reserve lagoon system, Puerto Rico. Fisheries Research, 76 (1), 81-90.

Fiala-Medioni A., 1987. Lower chordates. In: Pandian, T.J., Vernberg, F.J. (Eds.), Animal energetics, Vol 2. Academic Press, San Diego, pp. 323-356.

Fiala-Medioni, A. 1978. Filter-feeding ethology of benthic invertebrates (ascidians). IV. Pumping rate, filtration rate, filtration efficiency. Marine Biology, 48, 243-249.

Fiala-Médoni, A. 1973. Ethologie alimentaire d’invertébrés benthiques filtreurs (ascidies). I. Dispositif expérimental. Taux de filtration et de digestion chez Phallusia mammillata. Marine Biology, 23, 137-145.

Fisher,M.R. and Ditton, R.B. 1993. A social and economic characterization of the U.S. Gulf of Mexico recreational

136


shark fishery. Marine Fisheries Review, 55, 21-27.

Flagella, M.M. and Abdulla, A. 2005. Ship ballast water as a main vector of marine introductions in the Mediterranean. WMU Journal of Maritime Affairs, 4(1), 97-106.

Forcada, A., Valle, C., Sánchez-Lizaso, J.L., Bayle-Sempere, J.T. and Corsi, F. 2010. Structure and spatio-temporal dynamics of artisanal fisheries around a Mediterranean marine protected area. ICES Journal and Marine Sciences, 67, 191-203.

Forcada, A., Valle, C., Sánchez-Lizaso, J.L., Bayle-Sempere, J.T. and Corsi, F. 2010. Structure and spatio-temporal dynamics of artisanal fisheries around a Mediterranean marine protected area. ICES Journal and Marine Sciences, 67: 191-203.

Francour, P. 1994. Pluriannual analysis of the reserve effect on ichthyofauna in the Scandola natural reserve (Corsica, Northwestern Mediterranean). Oceanologica Acta, 17 (3), 309-317.

Francour, P., Harmelin, J.G., Pollard, D. and Sartoretto, E. 2001. A review of marine protected areas in the northwestern Mediterranean region: siting, usage, zonation and management. Aquatic Conservation: Marine and Freshwater Ecosystems, 11, 155-188.

Francour, P.; Ganteaume, A. and Poulain, M. 1999. Effects of boat anchoring in Posidonia oceanica seagrass beds in the Port-Cros National

Park (north-western Mediterranean Sea). Aquatic Conservation, 9, 391-400.

Fraschetti, S., Terlizzi, A., Bussotti, S., Guarnirei, G., D’Ambrosio, P. and Boero, F. 2005. Conservation of Mediterranean seascapes: analyses of existing protection schemes. Marine Environmental Research, 59(4), 309-332.

Fraschetti, S., Terlizzi, A., Micheli, F., Benedetti-Cecchi, L., and Boero, F. 2002. Marine Protected Areas in the Mediterranean Sea: objectives, effectiveness and monitoring. Marine Ecology, 23(1), 190-200.

Froese R, Pauly D. 2004. Fishbase. World Wide Web electronic publication. www.fishbase.org

Frost, J.E. and McCool, S.F. 1988. Can visitor regulations enhance recreational experiences. Environmental Management, 12(1), 5-9.

Gabriele, M., Bellot, A., Gallotti, D. and Brunetti, R. 1999. Sublitoral hard substrate communities of the northern Adriatic Sea. Cahiers de Biologie Marine, 40, 65–76.

Galil, S.B. 2000. A sea under siege – alien species in the Mediterranean. Biological Invasions, 2, 177–186.

Galil, S.B. 2007. Loss or gain? Invasive aliens and biodiversity in the Mediterranean Sea. Marine Pollution Bulletin, 55, 314–322.

Gambi M. C., Borg J. A., Buia M. C., Di Carlo G., Pergent-Martini C., Pergent G. and Procaccini G. 2006.

137

____________________________________________________________Bibliography

Proceedings of the Mediterranean Seagrass Workshop 2006, 29 May – 4 June 2006, Marsascala, Malta; Biologia Marina Mediterranea, 13(4), 293pp.

Ganteaume, A., Bonhomme, P., Bernard, G. and Poulain, M. 2004. Suivi de l’impact des mouillages forains sur l’herbier à Posidonia oceanica dans le Parc National de Port-Cros (Méditerranée nord-occidentale). Contrat Parc National de Port-Cros and GIS Posidonie. GIS Posidonie publications, France. 40 pp.

García-Charton, J.A., Bayle, J.T., Sánchez-Lizaso, J.L., Chiesa, P., Llaurado, F., Pérez, C. and Djian, H. 1993. Respuestas de la pradera de Posidonia oceanica y su ictiofauna asociada al anclaje de embarcaciones en el parque Nacional de Port-Cros (Francia). In: Pérez-Ruzafa,A. and Marcos-Diego, C (Eds,). Estudios del bentos marino. Instituto español de Oceanografía, 11,423-430.

García-Rubies, A. And Macpherson, E. 1995 Substrate use and temporal pattern of recruitment in juvenile fishes of the Mediterranean littoral. Marine Biology, 124(1), 35-42.

Garrabou, J., Perez, T., Sartoretto, S. and Harmelin, J.G. 2001. Mass mortality event in red coral Corallium rubrum populations in the Provence region (France, NW Mediterranean). Marine Ecology Progress Series, 217,263-272.

Garrabou, J., Sala, E., Arcas, A. and Zabala, M. 1998. The impact of diving on rocky sublittoral communities: a case

study of a bryozoan population. Conservation Biology, 12(2) 302–312.

Gell, F.R. and Roberts, C.M. 2003. Benefits beyond boundaries: the fishery effects of marine reserves. Trends in Ecology and Evolution, 18(9), 448-455.

Generalitat Valenciana, 1998. Ley 9/1998, del 15 de Diciembre, de pesca Marítima de la Comunidad Valenciana [1998/q11013](DOGV nº. 3395, de 17.12.98).

Generalitat Valenciana, 2005. Pla d´Ordenació dels Recursos Naturals (P.O.R.N.) de la Serra Gelada i la seua Zona Litoral. Conselleria de Territori i Habitatge. DOGV nº 4967, de 16.03.05.

Generalitat Valenciana, 2005. Pla d´Ordenació dels Recursos Naturals (P.O.R.N.) de la Serra Gelada i la seua Zona Litoral. Conselleria de Territori i Habitatge. DOGV nº 4967, de 16.03.05.

GESAMP. 1993. Impact of oil and related chemicals and wastes on the marine environment. Reports and Studies GESAMP 50. International Maritime Organization, London, U.K. 180p.

Gili, J. M. and Coma R. 1998. Benthic suspension feeders: Their paramount role in littoral marine food webs. Trends in Ecology and Evolution, 13, 316–321.

Gili, J.M., Coma, R., Orejas, C., López-González, P. and Zabala, M., 2001. Are Antarctic suspension-feeding communities different from those elsewhere in the world? Polar Biology, 24, 473-485.

138


Glover, J.H. 1980. Allocation of Fisheries Resources. Proceedings of the Technical Consultation on Allocation of Fishery Resource held in Vichy France, 20-23 April. Rome, Italy: FAO. 623pp.

Goñi, R., Polunin, N.V.C. and Planes, S. 2000. The Mediterranean: marine protected areas and the recovery of a large marine ecosystem. Environmental Conservation, 27, 95-97.

Gonzalez-Correa, J.M., Bayle, J.T., Sánchez-Lizaso, J.L., Valle, C., Sanchez-Jerez, P. and Ruiz, J.M. 2005. Recovery of deep Posidonia oceanica meadows degraded by trawling. Journal Experimental Marine Biology and Ecology, 320, 65-76.

Goodbody, I. 1974. The physiology of Ascidians. In: Russel, F.S. and Yonge, M. (eds.), Advances in Marine Biology. Academic Press, New York, pp. 2-149.

Goodbody, I. 1995. Ascidians communities in Southern Belize – a problem in diversity and conservation. Aquatic Conservation: Marine and Freshwater. Ecosystems, 5, 355-358.

Gordoa, A. 2009. Characterization of the infralitoral system along the north-east Spanish coast based on sport shore-based fishing tournament catches. Estuarine, Coastal and Shelf Science, 82, 41-49.

Granek, E.F., Madin, E.M., Brown, M.A., Figueira, W., Cameron., D.S., Hogan, Z., Kristianson,G., de Villiers, P., Williams, J.E., Post, J., Zahin, S. and Arlinghaus, R., 2008. Enganging Recreational Fishers in Management and Conservvation. Global Case

Studies. Conservation Biology, 22(5), 1125-1134.

Green, E. and Donnelly, R. 2003. Recreational Scuba Diving In Caribbean Marine Protected Areas: Do the users Pay?. Ambio, 32 (2), 140-144.

Grime, J.P. 1977. Evidence for the existence of three primary strategies in plants and its relevance to ecological and evolutionary theory. American Naturalist, 111, 1169–1194.

Halas, J.C. 1985. A unique mooring system for reef management in the Key Largo National Marine Sanctuary. In: Gabrie, C. and Salvat, B. (eds). Proceedings of the 5th International Reef Congress. Antenne Museum-EPHE, Moorea, French Polynesia, pp. 237–242.

Hall-Spencer, J.M., Rodolfo-Metalpa, R., Martin, S., Ransome, E., Fine, M., Turner, S.M., Rowley, S.J., Tedesco, D. and Buia, M.C. 2008. Volcanic carbon dioxide vents show ecosystem effects of ocean acidification. Nature, 454, 46-7.

Halpern, B.S. 2003. The impact of marine reserves: do reserves work and does reserve size matter? Ecological Application, 13,117–137.

Halpern, B.S.and Warner, R.R. 2002. Marine reserves have rapid and lasting effects. Ecological Letters, 5,361–366.

Hammitt, W.E. and Cole, D.N., 1998, Wildland Recreation, Ecology and Management, 2nd ed., John Wiley & Sons, Inc, New York, EEUU. 361 pp.

139

____________________________________________________________Bibliography

Harlin, M.M. 1975. Epiphyte-host relations in seagrass communities. Aquatic Botany, 1, 125-131.

Harmelin, J.G. 2000. Mediterranean marine protected areas: some prominent traits and promising trends. Environmental Conservation, 27, 104-105.

Harmelin-Vivien, M., Bitar, G., Harmelin, J.G. and Monestiez, P. 2005. The littoral fish

Harriot, V. J., Davis, D. and Banks, S. A. 1997. Recreational diving and its impact in marine protected areas in eastern Australia. Ambio, 26, 173–179.

Harriott VJ. 2002. Marine tourism impacts and their management on the Great Barrier Reef. CRC Reef Research Centre Technical Report No. 46. 41pp.

Hawkins, J.P. and C.M. Roberts. 1997. Estimating the carrying capacity of coral reefs for recreational scuba diving. Procedings of 8th International Coral reef Symposium, Panama, vol 2, 1923-1926.

Hawkins, J.P. and Roberts, C.M. 1992. Effects of recreational SCUBA diving on fore-reef slope communities of coral reefs. Biological Conservation, 62, 171–179.

Hawkins, J.P. and Roberts, C.M. 1994. The growth of coastal tourism in the Red Sea: present and future effects on coral reefs. Ambio, 23, 503–508.

Hawkins, J.P., Roberts, C.M., Kooistra, D., Buchan, K. and White, S. 2005. Sustainability of scuba diving tourism

on coral reefs of Saba. Coastal Management, 33, 373-387

Hawkins, J.P., Roberts, C.M., van’t Hof, T., de Meyer, K., Tratalos, J. and Aldam, C. 1999. Effects of recreational scuba diving on Caribbean coral and fish communities. Conservation Biology, 13, 888–897.

Hayne, D.W. 1991. The access point survey: procedures and comparison with the roving-clerk creel survey. American Fisheries Society Symposium, 12, 123–138.

Hilty, J.A. and Merenlender, A. 2000. Faunal indicator taxa selection for monitoring ecosystem health. Biological Conservation, 92(2), 185-197.

Hong, J.S. 1983. Impact de la pollution on the benthic community. Environmental impact of the pollution on the benthic coralligenous community in the Gulf of Fos, northwestern Mediterranean. Bulletin of the Korean Fisheries Society, 16, 273–290.

Howard, M. and Potter, B. 2002. Wise Coastal Practices for Sustainable Human Development Forum. “Small islands: limits of acceptable change.” In Response to:“Methodologies for carrying capacity in small island states, Indian Ocean.” Posted on 10 May 2002. web: http://www.csiwisepractices. org/?read=420

Hoyle, G., 1953. Spontaneous squirting of an ascidian, Phallusia mammillata Cuvier. Journal of the Marine Biological Association of the United Kingdom, 31, 541-562.

140


Hudgens, G. and Fatkin, L. 1985. Sex difference in risk taking: repeated sessions on a computer simulated task. Journal of Psychology, 119, 197–206.

Hudgins, M.D. 1984. Structure of the angling experience. American Fisheries Society, 113, 750-759.

Hughes, T.P., 1994: Catastrophes, phase shifts, and large-scale degradation of a coral reef. Science. 265. 1547-1551.

Hydrobiologia, 580(1), 7-21

Imber, S., Wasil, E., Golden, B. and Stagg, C. 1991. Selecting a survey design to monitor recreational angling for striped bass in the Chesapeake Bay. Socio-Economic Planning Science, 25(2), 113-121.

IUCN. 2007. IUCN Red List of Threatened Species. Web: http://www.iucnredlist.org.

IUCN. 2008. Maritime traffic effects on biodiversity in the Mediterranean Sea: Review of impacts, priority areas and identification of biodiversity offsets. IUCN Technical Paper.

Jørgensen, C.B. and Goldberg, E.D., 1953. Particle filtration in some ascidians and lamellibranchs. Biological Bulletin, 105, 477–489.

Jørgensen, C.B., 1949. Feeding rates of sponges, lamellibranchs and ascidians. Nature, 163,192.

Jøsrgensen, C.B., 1955. Quantitative aspects of filter-feeding in invertebrates. Biological Reviews, 30, 391-454.

Kearney, R.E. 2002. Co-management: the resolution of conflict between commercial and recreational fishers in Vistoria, Australia. Ocean and Coastal Management, 45, 201-214.

Keith, J.E., Fawson,C. and Van Johnson,V. 1996. Preservation or use: a contingent valuation study of wilderness designation in Utah. Ecological Economics, 18, 207-214.

Kelleher, G., Bleakey, C. and Wells, S. 1995. Introduction. In: Kelleher, G., Bleakey, C. and Wells, S. (Eds.), A Global Representative System of Marine Protected Areas, pp. 1–44. The Great Barrier Reef Marine Park Authority, The World Bank and the World Conservation Union (IUCN), Washington, DC. 147 pp.

Kenchington, R. and Kelleher, G. 1995. Making a management plan. In Gubbay, S. (Ed.): Marine protected areas: principles and techniques for management. Londres, Chapman & Hall, pp. 85-118.

Keough, M.J. and Quinn, G. P. 1998. Effects of periodic disturbances from trampling on rocky intertidal algal beds. Ecological Applications, 8, 141-161.

Kimmerer, J., Gartsidea, E. And Orsi, J., 1994. Predation by an introduced clam as the probable cause of substantial declines in zooplankton in San Francisco Bay. Marine Ecology Progress Series, 113, 81-93.

Kingsford, M. And Battershill, C. 1998. Studying Marine Temperate environments: A handbook for

141

____________________________________________________________Bibliography

ecologists. Canterbury Univ. Press. Christchurch, New Zealand.

Klumpp, D.W. 1984. Nutritional ecology of the ascidian Pyura stolonifera: influence of body size, food quantity and quality on filter-feeding, respiration, assimilation efficiency and energy balance. Marine Ecology Progress Series, 19, 269-284.

Knopf, R. 1990. People and protected natural environments: Emerging research concerns. Towards Serving Visitors and Managing Our Resources: Proceedings of a North American Workshop on Visitor Management in Parks and Protected Areas. University of Waterloo, Waterloo, Ontario, Canada. pp 107–117.

Kohn, J. and Gowdy, J. 1999. Coping with complex and dynamic systems. An approach to a transdisciplinary understanding of coastal zone developments. Journal of Coastal Conservation, 5 (2), 163.

Kremen, C. 1994. Biological inventory using target taxa: a case study of the butterflies of Madagascar. Ecological Applications. 4, 407-422.

Kulbicki, M. 1998. How the acquired behaviour of commercial reef fishes may influence the results obtained from visual censuses. Journal of Experimental Marine Biology and Ecology, 222, 11–30.

Kwak, T.J. and Henry, M.G. 1995. Largemouth bass mortality and related causal factors during live-release fishing tournaments on a large Minnesota lake.

North American Journal of Fisheries Management, 15, 621-630.

Laborel, J. 1961. Le concrétionnement algal “coralligène“ et son importance géomorphologique en Méditerranée. Recueil des Travaux de la Station Marine d’Endoume, 37, 37–60.

Laffoley, D. 1995. Techniques for managing marine protected areas: zoning. In: Gubbay, S. (Ed.), Marine Protected Areas. Principles and techniques for management. Chapman and Hall, London. pp 61-84.

Lampérez, T. 2001. Gestion de la reserva Marina de las Islas Columbretes. Actas de Las I Jornadas Sobre Reservas Marinas y I Reunión de La Red Iberoamericana de Reservas Marinas (Rirm). Cabo de Gata, Almería,17-23 de septiembre de 2001. MAPA.

Landres, P.B., Verner, J. and Thomas, J.W. 1988. Critique of vertebrate indicator species. Conservation Biologist. 2, 316-328.

Lelieveld, J., Berresheim, H., Borrmann, S., Crutzen P.J., Dentener, F.J., Fischer, H., Feichter, J., Flatau, P.J., Heland, J., Holzinger, R., Korrmann, R., Lawrence, M.G., Levin, Z., Markowicz, K.M., Mihalopoulos, N., Minikin, A., Ramanathan, V., De Reus, M., Roelofs, G.J., Scheeren, H.A., Sciare, J., Schlager, H., Schultz, M., Siegmund, P., Steil, B., Stephanou, E.G., Stier, P., Traub, M., Warneke, C., Williams, J. and Ziereis, H. 2002. Global air pollution crossroads over the Mediterranean. Science, 298(5594), 794-9.

142


Leriche, A., Pasqualini, V., Boudouresque, C.F., Bernard, G., Bonhomme, P., Clabaut, P., and Denis, J. 2006. Spatial, temporal and structural variations of a Posidonia oceanica seagrass meadow facing human activities. Aquatic Botany, 84, 287-293.

Leujak, W. and Ormond, R.F.G. 2008. Reef walking on Red Sea reef flats – quantifying impacts and identifying motives. Ocean and Coastal Management, 51, 755–762.

Lewin, W.C., Arlinghaus, R. and Mehner T. 2006. Documented and potential biological impacts of recreational fishing: Insights for management and conservation. Reviews in Fisheries Science, 14(4), 305-367.

Lewis, J. 1996. Coastal benthos and global warming: strategies and problems. Marine Pollution Bulletin, 32 (10), 698–700.

Liddle, M.J. 1991. Recreation ecology: effects of trampling on plants and corals. Trends Ecology Evolution, 6, 13-17.

Linares, C., Coma, R., Diaz, D., Zabala, M., Hereu, B. and Dantart, L. 2005. Immediate and delayed effects of a mass mortality event on gorgonian population dynamics and benthic community structure in the NW Mediterranean Sea. Marine Ecology Progress Series, 305,127–137.

Lively, C.M., Raimondi, P.T. and Delph, L.F. 1993. Intertidal community structure: space-time interactions in the northern Gulf of California. Ecology, 74, 162-173.

Lloret, J., Marín, A., Marín-Guirao, L. and Carreño, M.F., 2006. An alternative approach for managing scuba diving in small marine protected areas. Aquatic conservation: Marine and Freshwater Ecosystems, 16, 579-591.

Lloret, J., Zaragoza, N., Caballero, D. and Riera V. 2008. Biological and socioeconomic implications of recreational boat fishing for the management of fishery resources in the marine reserve of Cap de Creus (NW Mediterranean). Fisheries Research, 91, 252–259.

Lockwood RN. 1999. Evaluation of on-site angler survey methods, study 673. Annual Reports for Projects F-35-R-24 and F-53-R-15. 125 pp.

Lockwood, R.N. 1997. Evaluation of catch rate estimators from Michigan access point angler surveys. North American Journal of Fisheries Management, 17, 611-620.

Lockwood, R.N., Peck, J. and Oelfke, J. 2001. Survey of angling in Lake Superior waters at Isle Royale National Park, 1998. Noth American Journa Fisheries Management, 21, 471-481.

Loomis, J., Kent, P., Strange, L., Fausch, K. and Covich, A.1999. Measuring the total economic value of restoring ecosystem services in an impaired river basin: results from a contingent valuation survey. Ecological economics, 33, 103-117.

Lowe, C. 2003. Sustainability and the question of “enforcement” in integrated coastal management: the case of Nain Island, Bunaken National Park.

143

____________________________________________________________Bibliography

Indonesian Journal of Coastal and Marine Resources Special Edition, 1,49–63.

Lozano, G., 1972. Nota sobre el hallazgo de dos especies de Urocordados (Tunicata), nuevas para la fauna marina de Tenerife. Vieraea, 1972, 45-51.

Lubchenco, J., Palumbi, S.R., Gaines, S.D. and Andelman, S. 2003. Plugging a hole in the ocean: the emerging science of marine reserves. Ecological Applications, 13(1), 3-7.

Luna-Pérez, B., Valle-Pérez, C. and Sánchez-Lizaso, J. L., 2009. Benthic impacts of recreational divers in a Mediterranean Marine Protected Area. ICES Journal of Marine Sciences, 66, 517-523.

Luna-Pérez,B., Valle, C., Vega Fernández, T., Sánchez-Lizaso, J.L. and Ramos-Esplá, A.A. 2010. Halocynthia papillosa (Linnaeus, 1767) as an indicator of SCUBA diving impact. Ecological Indicators, 10 (5), 1017-1024.

Lynch, T.P., Wilkinson, E., Melling, L., Hamilton, R., Macready, A. and Feary, S. 2004. Conflict and Impacts of divers and anglers in a Marine Park. Enviromental Management, 33(2), 196-211.

MacGinite, G.E. 1939. The method of feeding of tunicates. Biological Bulletin, 77, 443-447.

Machias, A., Giannoulaki, M., Somarakis, S., Maravelias, C.D., Neofitou, C., Koutsoubas, D.,

Papadopoulou, K.N. and Karakassiss, I. 2006. Fish farming effects on local fisheries landing in oligotrophic area. Aquaculture, 261, 809-816.

Machias, A., Karakassis, I., Labropoulou, M., Somarakis, S., Papadopoulou K. and Papaconstantinou, C. 2004. Changes in wild fish assemblages after the establishment of a fish farming zone along in an oligotrophic marine ecosystem. Estuarine, Coastal and Shelf Sciences, 60, 711-779.

Macpherson, E. 1998. Ontogenetic shifts in habitats use and aggregation in juvenile sparids fishes. Journal Experimental Marine Biology and Ecology, 220, 127-150.

Malvestuto, S.P. 1983. Sampling the recreational fishery. In: Nielsen, L.A. and Johnson, D.L. (Eds). FisheriesTechniques. Maryland, USA: American Fisheries Society, Bethesda, pp. 397–419.

Mann, C. 2003. Concepts for Recreation Management: Carrying Capacity & Limits of Acceptable Change. Repport of the Institute of Forest Policy, July 2003.

Manning, R. 2003. Emerging Principles for Using Information/Education in Wilderness Management. International Journal of Wilderness, 9, 20-29.

Marion, J.L. and Rogers, C.S. 1994. The applicability of terrestrial visitor impact management strategies to the protection of coral reefs. Ocean and Coastal Management, 22, 153–163.

144


Martín, M.A., Sànchez-Lizaso, J.L. and Ramos-Esplà, A.A. 1997. Cuantificación del impacto de las artes de arrastre sobre la pradera de Posidonia oceanica (L.) Delile, 1813. Publicación Especial del Instituto Español de Oceanografia, 23, 243-253.

Martínez, C., Mallol, J., Arnés, I. and Sánchez-Lizaso, J.L. 1999. Impacto del anclaje sobre la pradera de Posidonia oceanica en las zonas de fondeo autorizado de la Reserva Marina de Tabarca. I Jornadas Internacionales de Reservas Marinas. 24-26 Marzo, 1999. Murcia. Pp. 78.

McAllister, D.E. 1996. The Status of the World Ocean and Its Biodiversity. Ocean Voice International, Ottawa.

McClanahan, T.R. and Mangi, S. 2001. The effect of closed area and beach seine exclusion on coral reef fish catches. Fisheries Management and Ecology, 8, 107–121.

McCool, S.F. and Cole, D.N. 1998. Experiencing Limits of Acceptable Change: Some Thoughts After a Decade of Implementation. In: McCool, Stephen F.; Cole, David N. (Eds.), Proceedings: Limits of Acceptable Change and related planning processes: progress and future directions; 1997 May 20-22; Missoula, MT. pp 72-78

McPhee, D.P. and Skilleter, G.A. 2002. Harvesting of intertidal animals for bait for use in a recreational fishing competition. Proceedings of the Royal Society of Queensland, 110, 19–27.

McRoy, C.P. and McMillan, C. 1977. Productivity and physiological ecology

of seagrasses. In: C.P. McRoy and Helfferich, C. (Eds.). Seagrass ecosystems: a scientific perspective. Marcel Dekker, New York, pp. 53-88.

Medina, A., Abascal, F.J., Aragón, L., Mourente, G., Aranda, G, Galaz, T., Belmonte, A., de la Serna,J.M. and García,S. 2007. Influence of sampling gear in assessment of reproductive parameters for bluefin tuna in the western Mediterranean. Marine Ecology Progress Series, 337, 221-230.

Medio, D., Ormond, R.F.G. and Pearson, M. 1997. Effects of briefings on rates of damage to corals by SCUBA divers. Biological Conservation, 79, 91–95.

Meinesz, A. and Lefevre, J.R. 1978. Destruction de l’éstage infralitoral des Alpes Maritimes (France) et de Monaco par les restructurations du rivage. Bulletin Ecological, 9, 259-276.

Methratta, E.T. and Link, J.S. 2006. Evaluation of Quantitative Indicators for Marine Fish Communities. Ecological Indicators, 6,575-588.

Meyer, C.G. 2007. The impacts of spear and other recreational fishers on a small permanent Marine Protected Area and adjacent pulse fished area. Fisheries Research, 84, 301-307.

Mike,A. and Cowx, I.G. 1996. A preliminary appraisal of the contribution of recreational fishing to the fisheries sector in north-wet Trinidad. Fisheries Management and Ecology, 3, 219-228.

Milazzo, M., Anastasi, I. and Willis, T.J. 2006. Recreational fish feeding

145

____________________________________________________________Bibliography

affects coastal fish behavior and increases frequency of predation on damselfish Chromis chromis nests. Marine Ecology Progress Series, 310, 165–172.

Milazzo, M., Badalamenti, F., Ceccherelli, G. and Chemello, R. 2004. Boat anchoring on Posidonia oceanica beds in a marine protected area (Italy, western Mediterranean): effect of anchor types in different anchoring stages. Journal of Experimental Marine Biology and Ecology, 299, 51-62.

Milazzo, M., Badalamenti, F., Vega Fernández, T. and Chemello, R. 2005. Effects of fish feeding by snorkellers on the density and size distribution of fishes in a Mediterranean marine protected area. Marine Biology, 146 (6), 1213-1222.

Milazzo, M., Chemello, R., Badalamenti, F., Camarda, R. and Riggio, S. 2002b. The impact of human recreational activities in marine protected areas: what lessons should be learnt in the Mediterranean Sea? PSZNI Marine Ecology, 23, 280–290.

Milazzo, M.; Chemello, R.; Badalamenti, F. and Riggio, S. 2002a. Short-term effect of human trampling on the upper infralittoral macroalgae of Ustica Island MPA. Journal Experimental Marine Biology and Ecology, 82,745-748.

Millar, R.H. 1971. The biology of Ascidians. Advances in Marine Biology 9, 1-100.

Mistri, M. and Ceccherelli, V.U. 1994. Growth and secondary production of the

Mediterranean gorgonian Paramuniceaclavata. Marine Ecology Progress Series, 103, 291-296.

Mittermeier, R. A., Gil, P.R., Hoffmann, M., Pilgrim, J., Brooks, T., Goettsch, Mittermeier, C., Lamoreux J. 2004. Hotspots Revisited: Earth's Biologically Richest and Most Endangered Terrestrial Ecoregions. Cemex Books on Nature. Web: http://multimedia.conservation.org/cabs/online_pubs/hotspots2/MediterraneanBasin.html.

Moberg, F. and Folke, C. 1999. Ecological Goods and Services of Coral Reef Ecosystems. Ecological Economics, 29 (2) 215-33.

Moliner, R. and Picard, J. 1953. Notes biologiques a propos d’un voyage d’etude sur les cotes de Sicile. Annales de l’Institut Oceanographique, 28(4),163-188.

Monniot, F. 1972. Les ascidies littorales et profondes des sédiments. Smithsonian Contributions to Zoology, 76, 119-126.

Montefalcone, M., Lasagana, R., Bianchi, C.N. and Morri, C. 2006. Anchoring damage on Posidonia oceanica meadow cover: a case of study in Prelo Cove (Ligurian Sea, NW Mediterranean). Chemistry and Ecology, 22, 207-217.

MOPU. 1987. Estudio geofísico marino en la costa de la provincia de Alicante. Tramos Puerto de Vilajoyosa-Punta Albir. Subdirección General de Costas y Señales Miritimas. Dirección General de Puertos y Costas, Madrid.

146


Morales-Nin, B., Moranta, J., García, C., Tugores, M.P., Grau, A.M., Riera, F. and Cerdà, M. 2005. The recreational fishery off Majorca Island (western Mediterranean): some implications for coastal resource management, ICES Journal Marine Science 62, 727–739.

Morey, E.D., Shaw, W.D. and Rowe, R.D. 1991. A discrete-choice model of recreational participation, site choice, and activity valuation when complete trip data are not available. Journal of Enviromental Economics and Management, 20, 181-201.

Morganti, C., Cocito, S. and Sgorbini, S. 2001. Contribution of biocostructor to coralligenous assemblages exposed to sediment deposition. Biologia Marina Mediterranee, 8, 283–286.

Muoneke, M.I. and Childress,W.M. 1994. Hooking mortality: a review for recreational fisheries. Reviews Fisheries Science, 2, 123–156.

Murray-Jones, S. and Steffe, S.A. 2000. A comparison between the comercial and recreational fisheries of the surf clam, Donax deltoides. Fisheries Research,44, 219-233.

Myers, N., Mittermeier, R.A., Mittermeier, C.G., da Fonseca, G.A.B., and Kent, J. 2000.Biodiversity hotspots for conservation priorities. Nature 403:853-858

Myers, R.A. 1997. Comment and reanalysis: paradigms for recruitment studies. Canadian Journal of Fsheries and Aquatic Sciences, 54(4), 978-981.

Nagelkerken, I., Buchan, K., Smith, G.W., Bonair, K., Bush, P., Garzón-Ferreira, J., Botero, L., Gayle, P., Harvell, C.D., Heberer, C, Kim, K., Petrovic, C., Pors, L. and Yoshioka, P.M. 1997. Widespread disease in Caribbean sea fans. II. patterns of infection and tissue loss. Marine Ecology Progress Series, 160, 255-263.

Naranjo, S.A., Carballo, J.L. and Garcia J.C. 1996. The effects of environmental stress on ascidian populations in Algeciras Bay (Southern Spain). Possible marine bioindicator? Marine Ecology Progress Series, 144, 119-131.

Neis, B., Schneider, D. C., Felt, L., Haedrich, R. L., Fischer, J., and Hutchings, J. A. 1999. Fisheries assessment: what can be learned from interviewing resource users? Canadian Journal of Fisheries and Aquatic Sciences, 56, 1949–1963.

Neis, B., Schneider, D.C., Felt, L., Haedrich, R.L., Fischer, J. and Hutchings, J.A. 1999. Fisheries assessment: what can be learned from interviewing resource users? Canadian Journal of Fisheries and Aquatic Sciences, 56: 1949–1963.

Norse, E.A. 1993. Global Marine Biological Diversity. A Strategy for Building Conservation into Decision Making. Island Press, Washington, USA.

Notarbartolo di Sciara, G. 2003. Establishing Marine Protected Areas for Cetaceans in the ACCOBAMS Area. Meeting of the Scientific Committee of ACCOBAMS, Istanbul, 20-22 November 2003.

147

____________________________________________________________Bibliography

Oracion, E. 2003. The dynamics of stakeholders participation on marine protected area development: a case of study in Batangas, Philippines. Silliman Journa, 44 (1): 95-137.

Orams, M. 1999. Marine tourism : Development, impacts and management. Routledge Ltd. London. 115 pp.

Ormond, R. F. G., Gage, J. D. and Angel, M. V. 1997. Marine biodiversity: patterns and processes. Cambridge University Press, New York, USA.

Ott,W. 1978. Environmental Indices: Theory and practice. Ann Arbor Science Publishers, Miami.

Palermo, V. and Thompson, A.S. 2000. Angler effort and match in the 1998 sport fisheries of five lower Fraser River tributaries. Canadian Manuscript Fisheries Aquatic Science No. 2530, 55 pp.

Pawson, M.G., Glenn, H. and Padda, G. 2008. The definiton of marine recreational fishing Europe. Marine Policy, 32(3), 339-350.

Peluso, N. 1992. Coercing conservation: the politics of state resource control. Global Environmental Change, 4,199–218.

Pendleton, L.H. 1994; Environmental quality and recreation demand in a caribbean coral reef Coastal Management, 22 (4), 399 – 404.

Pérès, J.M. and Picard, J. 1964. Nouveau manuel de bionomie benthique

de la mer Méditerranée. Recueil des Travaux de la Station Marine d’Endoume, 31, 5-137.

Pérès, J.M. and Picard, J. 1975. Causes de la raréfraction et de la dispartion des herbies de Posidonia oceanica sur les côtes françaises de la Mediterranée. Aquatic Botany, 1, 133- 139.

Pérès, J.M., 1967. The Mediterranean benthos. Oceanographic and Marine Biology Annual Review, 5, 449-533.

Pérès, J.M., 1991. Les ‘violets’, ascidies comestibles du genre Microcosmus, espèces à protéger pour une meilleur gestion du stock. In: Boudouresque, C.F., Avon, M. and V. Gravez (eds.). Les Espèces Marines à Protéger en Méditerranée. GIS Posidonie publ., France, pp. 211-218.

Pérez, T., Garrabou, J., Sartoretto, S., Harmelin, J. G., Francour, P., and Vacelet, J. 2000. Mortalité massive d’invertébrés marins: un evénement sans préceédent en Méditerranée nord-occidentale. Life Sciences, 323: 853–865.

Petchey, O.L., Hector, A. and Gaston, K.J. 2004. How do different measures of functional diversity perform? Ecology,85, 847–857.

Petersen, J.K. 2007. Ascidian suspension feeding. Journal of Experimental Marine Biology and Ecology, 342, 127-137.

Petersen, J.K. and Riisgard, H.U. 1992. Filtration capacity of the ascidian Cionaintestinalis and its grazing impact in a

148


shallow fjord. Marine Ecology Progress Series,. 88, 9-17.

Petersen, J.K., Mayers, S. and Knudsen, M.A. 1999. Beat frequency of cilia in the branchial basket of the ascidian Ciona intestinalis in relation to temperature and algal cell concentration. Marine Biology, 133, 185-192.

Petersen, J.K., Schou, O. and Thor, P. 1995. Growth and energetics of the ascidian Ciona intestinalis (L.). Marine Ecology Progress Series, 120, 175-184.

Peterson, C.H., and Estes, J.A. 2001. Conservation and Management of Marine Communities. In: Bertness, M.D., Gaines S.D. and Hay, M.E. (Eds.), Marine Community Ecology. Sunderland, Massachusetts: Sinauer Associates, Inc.

Pitcher, T.J and Hollingworth, C.E. 2002. Recreational Fisheries: Ecological, Economic and Social Evaluation. Fish and Aquatic Resources Series 8, Blackwell Science, Oxford, England. 225 pp.

Plan Development Team 1990. The potential of marine fishery reserves for reef fish management in the U.S. Southern. Atlantic. NOAA Technical Memorandum.

Planes, S., Galzin, R., Bablet, J.P. and Sale, F.. 2005. Stability of coral reef fish assemblages impacted by nuclear tests. Ecology, 86 (10), 2578-2585.

Plathong, S., Inglis, G.J. and Huber, M.E. 2000. Effects of self guided snorkeling trails in a tropical marine

park. Conservation Biology, 14, 1821–1830.

Policansky, D. 2002. Catch-and-release recreational fishing: a historical perspective. In: Pitcher, T.J., Hollingworth, C.E. (eds). Recreational fisheries: ecological, economic and social evaluation. Fish and Aquatic Resources Series, 8. Oxford, EN: Blackwell Science, 74-94 pp.

Pollnac, R. B., Crawford, B. R. and Gorospe, M. L. G. 2001. Discovering factors that influence the success of community-based marine protected areas in the Visayas, Philippines. Ocean and Coastal Management, 44,683–710.

Pollock, B. 1980. Surprises in Queensland angling study. Australian Fisheries, 39(4), 17-19.

Pollock, K.H., Hoenig, J.M., Jones, C.M., Robson, D.S and Greene, C.J. 1997. Catch rate estimation for roving and access point surveys. North American Journal of Fisheries Management, 17 11-19.

Pollock, K.H., Jones, C.M. and Brown, T.L. 1994. Angler survey methods and their applications in fisheries management. American Fisheries Society Special Publication. No 25, 371 pp.

Post, J.R., Sullivan, M., Cox, S., Lester, N.P., Walters, C.J., Parkinson, E.A., Paul, A.J., Jackson, L. and Shuter, B.J.. 2002. Canada’s recreational fishery: the invisible collapse? Fisheries, 27(1), 6–17.

149

____________________________________________________________Bibliography

Povey, A. and Keough, MJ. 1991. Effects of trampling on plant and animal populations on rocky shores. Oikos, 61 (3), 355-368.

Prior, M., Ormond, R., Hinchen, R. and Wormald, C. 1995. The impact of natural resources of activity tourism: a case of study of diving in Egypt. International Journal of Environmental Studies, 48, 201–209.

PSICO. 2007. Partnership for Interdisciplinary Studies of Coastal Oceans. The Science of Marine Reserves (2nd Edition, International Version). http://www.piscoweb.org.

Pulfrich, A., Parkins, C.A. and Branch, G.M. 2003. The effects of shore-based diamond-diving on intertidal and subtidal biological communities and rock lobsters in southern Namibia. Aquatic Conservation: Marine and Freshwater Ecosystems, 13, 233–255.

Quinn, G., Keough, M. 2002. Experimental Design and data analysis for biologists. Cambridge University Press, Cambridge.

Ramos Esplá, A.A. 1992. Impact biologique et économique de la Réserve marine de Tabarca (Alicante, Sud-Est de l’Espagne). Ajaccioi, 26-28 septembre 1991. Medpan News, France. Vol 3, pp 59-66.

Ramos Esplá, A.A., Valle, C., Bayle, J.T. and Sánchez-Lizaso, J.L. 2004. Áreas marinas protegidas como herramientas de gestión pesquera en el Mediterráneo (Área COPEMED). Serie Informes y Estudios COPEMED, 11. 156 pp.

Ramos-Esplá AA. 1985. La Reserva Marina de la Isla Plana o Nueva Tabarca (Alicante). Alicante: Universidad de Alicante-Ayuntamiento Alicante, 196 pp.

Ramos-Esplá, A.A., 1991. Ascidias litorales del Mediterráneo ibérico. Faunística, ecología y biogeografía. Publicaciones Universidad de Alicante, Alicante.

Ramos-Esplá, A.A., Moreno, D. and De La Linde, A. 2008. Halocynthiapapillosa (Linnaeus, 1767), Microcosmus sabatieri (Roule, 1885) y M. vulgaris (Heller, 1877). In: Barea-Azcón, J.M., Ballesteros-Duperón, E. and Moreno, D. (eds). Libro Rojo de los Invertebrados de Andalucía. Consejería de Medio Ambiente, Junta de Andalucía, Tomo II: 652-656, y IV: 1357-1358.

Randløv, A. and Riisgard, H.U. 1979. Efficiency of particle retention and filtration rate in four species of ascidians. Marine Ecology and Progress Series, 1, 55–59.

Rapport, D.J. 1992. Evaluating ecosystem health. Journal of Aquatic Ecosystem Health, 1, 15-24.

Reid, D.D., Montgomery, S.S. 2005. Creel survey based estimation of recreational harvest of penaeid prawns in four southeastern Australian estuaries and comparison with commercial catches. Fisheries Research,74, 169-185.

Ribera, M.A. 1991. Medes Island nature reserve and regional tourism visitation. In: Olivier, J., Gerardin, N. and de

150


Grissac, J. Proceedings of the Economic Impact of the Mediterranean Coast Protected Areas Meeting. MEDPAN Secretariat Publications, Paris, pp. 51–58.

Ribera-Siguan, M.A. 1992. La réserve marine des ýˆles Medes. Bilan d’un succès imprévu. In: Parchi marini del Mediterraneo, Atti del 2° Convegno internazionale, I.CI.MAR, San Teodoro, 17–19 May 1991; 152–161 pp.

Ribes, M., Coma, R. and Gili, J.M. 1998. Seasonal variation of in situ rates by the temperate ascidians Halocynthiapapillosa. Marine Ecology and Progress Series, 175, 201-213.

Richez, G. 1991. Visitation during summer 1990 by scuba divers (snorkeling excluded) in Port Cros National Park (France). In: Olivier, J., Gerardin, N. and de Grissac, J. Proceedings of the Economic Impact of the Mediterranean Coast Protected Areas Meeting. MEDPAN Secreatariat Publications, Paris, pp. 85–89.

Richez, G. 1992. La navigation de plaisance dans l’anse d’Elbu (Réserve naturelle de Scandola, Cose du Sud): étes 1988 et 1989. Travaux Scientifiques du Parc Naturel Régional et des Réserves Naturelles de Corse, 36, 35–64.

Richez, G. 1993. La plongée sous marine de loisir en Corse. Apnée exclude, durant l´été 1991. Travaux Scientifiques du Parc Naturel Régional et des Réserves Naturelles de Corse, 45, 1–65.

Riegl, B. and Cook, P.A. 1995. Is damage susceptibility linked to coral community structure? A case study from South Africa. In: Piller, W.E. (ed). Proceedings of the 1st European Regional Meeting of the International Society for Reef Studies. Beiträge zur Paläontologie, Denmark, pp. 51–63

Riegl, B., and Riegl, A. 1996. Studies on coral community structure and damage as a basis for zoning marine reserves. Biological Conservation, 77,269–277.

Rius, M. 2007. The effect of protection on fish populations in the Ses Negres Marine Reserve (NW Mediterranean Spain). Scientia Marina,71(3), 499–504.

Roberts, C. M. and Hawkins, J. P. 2000. Fully-protected marine reserves: A guide. WWF Endangered Seas Campaign, Washington D.C., and Environment Department, University of York, York, UK.

Roberts, C.M., Bohnsack, J.A., Gell, F., Hawkins, J.P. and Goodridge, R. 2001. Effects of Marine Reserve on Adjacent Fisheries. Science, 294(5548), 1920-1923.

Roberts, L. and Harriott, V.J. 1994. Recreational scuba diving and its potential for environmental impact in a marine reserve. In: Bellwood, O., Choat, H. and Saxena, N. Recent Advances in Marine Science and Technology. James Cook University of North Queensland, Townsville, Australia, pp. 695–704.

Rocha, F., Gracia, J., González, A. F., Jardón, C. M., and Guerra, A. 2004.

151

____________________________________________________________Bibliography

Reliability of a model based on a short fishery statistics survey: application to the Northeast Atlantic monkfish fishery. ICES Journal of Marine Science, 61, 25–34.

Rocha, F., Gracia, J., González, A.F., Jardón, C.M., and Guerra, A. 2004. Reliability of a model based on a short fishery statistics survey: application to the Northeast Atlantic monkfish fishery.ICES Journal of Marine Science, 61, 25–34.

Rochet, M.J. and Trenkel, V.M. 2003. Which community indicators can measure the impact of fishing? A review and proposals. Canadian Journal of Fisheries 60(1), 86-99.

Rogers, C.S. 1990. Responses of coral reefs and reef organisms to sedimentation. Marine Ecology Progress Series, 62, 185-202.

Rogers, S. I. and Greenaway,B. 2005. A UK perspective on the development of marine ecosystem indicators . Marine Pollution Bulletin, 50: 9-19.

Roncin,N., Alban, F., Charbonnel,E., Crechriou, R., la Cruz Modino, R., Culiolie, J.M., Dimech,M., Goñi,R., Guala,I., Higgins, R., Lavisse, E., Le Direach, L., Luna, B., Marcos, C., Maynoum, F., Pascual, J., Person, J., Smithn, P., Stobart, B., Szelianszky, E., Valle, E., Vasellio, E., Boncoeur,J. 2008. Uses of ecosystem services provided by MPAs: How Much do they impact the local economy?. Journal of Nature conservation; 1(16), 256-270.

Ros, J. D., Romero, J., Ballesteros, E. and Gili, J.M. 1985. Diving in blue

water. The benthos. In: Margalef, R. (Ed.). Key Environment: Western Mediterranean. Pergamon Press, Oxford, pp. 233–295.

Roth, E., Toivonen, A.L., Navrud, S., Bengtsson, B., Gudbergsson, G., Tuunainen, P., Appelblad, H. and Weissglas, G. 2001. Methodological, conceptual and sampling practices in surveying recreational fisheries in the Nordic countries-experiences of a valuation survey. Fisheries Management and Ecology, 8, 355-367.

Rouphael, A.B. and Hanafy, M. 2007. An alternative management framework to limit the impact of SCUBA divers on coral assemblages. Journal of Sustainable Tourism, 15, 91–103.

Rouphael, A.B. and Inglis, G.J. 1997. Impacts of recreational SCUBA diving at sites with different reef topographies. Biological Conservation, 82(3), 329–336.

Rouphael, A.B. and Inglis, G.J. 2001.’Take only photographs and leave only footprints’? An experimental study of the impacts of underwater photographers on coral reef dive sites. Biological Conservation, 100, 281–287.

Rouphael, A.B. and Inglis, G.J. 2002. Increased spatial and temporal variability in coral damage caused by recreational scuba diving. Ecological Applications,12 (2), 427-440.

Rouphael, T. and Inglis, G.J. 1995. The effects of qualified recreational SCUBA divers on coral reefs. Technical report 4. CRC Reef Research Centre Ltd, Townsville, Australia, 39 pp.

152


Ruiz, J.M. 2000: Respuesta de la fanerógama marina Posidonia oceanica (L.) Delile a perturbaciones antrópicas. Tesis doctoral. Universidad de Murcia. 212 pp.

Sala, E. 2004. The past and present topology and structure of Mediterranean subtidal rocky-shore food webs. Ecosystems, 7, 333–340.

Sala, E., Garrabou, J., Zabala, M. 1996. Effects of diver frequentation on Marine sublitoral population of the bryozoan Pentapora fascialis. Marine Biology, 126, 451-459.

Salm, R.V., Clark, J. and Siirila, E. 2000. Marine and Coastal Protected Areas: a Guide for Planners and Managers, 3rd edn. IUCN, Washington DC. 371 pp.

Sánchez Lizaso, J.L., Goñi, R., Reñones, O., García Charton, J.A., Galzin, R., Bayle, J.T., Sánchez Jerez, P., Pérez Ruzafa, A. and Ramos, A.A. 2000. Density dependence in marine protected populations: a review. Environmental Conservation, 27, 144-158.

Sanchez-Jerez, P., Bayle-Sempere, J. Arechavala-López, P., Luna-Pérez, B., Vázquez-Luis, M., Ojeda-Martínez, C., Valle-Pérez, C., Forcada-Almarcha, A. and Fernández-Jover, D. 2007. Evaluación de la biodiversidad de peces marinos e impacto de la pesca deportiva en el Parc Natural de Serra Gelada. Universidad de Alicante y Conselleria de Territori i Habitage. Generalitat Valenciana, 163 pp.

Sánchez-Lizaso, J.L., 1993. Estudio de la pradera de Posidonia oceanica (L.) Delilde de la Reserva Marina de Tabarca (Alicante): Fenologia y producción primaria. Tesis Doctoral, Universitat d’Alacant: 121pp.

Sandersen, H. and Koester,S. 2000. Co-management of tropical coastal zones: the case of Sourfriere Marine Management Area, St Lucia, WI. Coastal Management, 28, 87-97.

Saphier, A. and Hoffmann, T.C. 2005. Forecasting Models to Quantify Three Anthropogenic Stresses on Coral Reefs From Marine Recreation: Anchor Damage, Diver Contact and Copper Emission From Antifouling Bottom Paint. Marine Pollution Bulletin, 51,590-598.

Sartoretto S., Verlaque, M. and Laborel, J. 1996. Age of settlement and accumulation rate of submarine “coralligène” (-10 to -60 m) of the northwestern Mediterranean Sea; relation to Holocene rise in sea level. Marine Geology, 130, 317-331.

Schiel, D.R. and Taylor, D.I. 1999. Effects of trampling on a rocky intertidal algal assemblage in southern New Zealand. Journal of Experimental Marine Biology and Ecology, 235 (2), 213-235.

Schleyer, M.H. and Tomalin, B.J. 2000. Damage on South African coral reefs and assessment of their sustainable diving capacity using a fisheries approach. Bulletin of Marine Science, 67, 1025–1042.

153

____________________________________________________________Bibliography

Schoeman, D.S. 1996. An assessment of a recreational beach clam fishery: current fishing pressure and opinions regarding the initiation of a commercial clam harvest. South African Journal of Wildlife Research, 26(4), 160-170.

Shackley, M. 1998. ‘Stingray City’—Managing the impact of underwater tourism in the Cayman Islands. Journal of Sustainable Tourism, 6, 328–338.

Shafer,C.S. and Inglis, G.J. 2000. Influence of social, biophysical, and managerial conditions on tourism experiences within the Great Barrier Reef World Heritage Area. Environmental Management, 26, 73–87.

Shi, H., Singh, A., Kant, S., Zhu, Z.L. and Waller, E. 2005. Integrating habitat status, human population pressure, and protection status into biodiversity conservation priority setting. Conservation Biology, 19(4), 1273-1285

Shrestha S, Bell MA, Marcotte PL. 2002. An economic impact assessment of Myanmar-IRRI country programs. In: IRRI country report. International Rice Research Institute, Los Baños, Philippines.

Shurin, J.B., Borer, E.T., Seabloom, E.W., Anderson, K., Blanchette, C.A., Broitman, B., Cooper, S.D. and Halpern, B. 2002. A cross-ecosystem comparison of the strength of trophic cascades. Ecology Letters,5, 785–791.

Smith, A. 1999. Analysis of the vessels over 100 tons in the global fishing fleet. FAO, Circular de Pesca. No 949. Roma.

Spaling, H. and Smit, B. 1993. Cumulative environmental change: conceptual frameworks, evaluation approaches and institutional perspective, Environmental Management, 17, 587–600.

Stankey, G.H. and Mc Cool, S.F. 1984. Carrying capacity in recreational settings. Lisure Sciences, 6(4), 453-473.

Steffe, A.S., Murphy, J.J., Chapman, D.J., Tarlinton, B.E., Gordon, G.N.G. and Grinberg, A. 1996. An assessment of the impact of offshore recreational use of fishing in NSW waters on the management of commercial fisheries. NSW Fisheries Research Institute. FDRC Report 94/053.

Steneck, R.S. and Carlton, J.T. 2001. Human alterations of marine communities: students beware! In: Bertness, M., Gaines, S., and Hay, M. (Eds.), Marine Community Ecology . pp. 445–468. Sunderland, MA, USA: Sinauer Press.

Storey, D.A. and Allen, G.P. 1993. Economic impact of marine recreational fishing in Massachusetts. North American Journal of Fisheries Management, 13,698–708.

Stuart, V., Klumpp, D.W., 1984. Evidence for food-resource partitioning by kelp-bed filter feeders. Marine Ecology Progress Series, 16, 27-37.

Sullivan, M.G. 2003. Active management of walleye fisheries in Alberta: dilemmas of managing recovering fisheries. North American Journal of Fisheries Management, 23, 1343-1358.

154


Tabata, R.S. 1989. The use of nearshore dive sites by recreational dive operations in Hawaii. Coastal Zone, 89, 2865–2875.

Tabata, R.S. 1992. Scuba-diving holidays. In: Weiler, B. and Hall, C.M. (Eds.). Special Interest Tourism, Belhaven Press, New York, pp. 171–184.

Talge, H. 1992. Impact of recreational divers on scleractinian corals at Looe Key, Florida. In: Proceedings of the Seventh International Coral Reef Symposium, 2 University of Guam Press, Guam, pp. 1077–1082.

Talge, H., 1990. Impact of recreational divers on coral reefs in the Florida Keys. In: Jaap, W.C. (Ed.), Diving for science. Proceding American Academy Underwater Science. 12th Annual Scientific Diving Symposium., Costa Mesa, California, USA: pp 365-372.

Taylor, R.G., Whittington, J.A., and Haymans, D.E., 2001. Catch-andrelease mortality rates of common snook in Florida. North American Journal Fisheries Management, 21, 70–75.

Thorstad, E.B., Hay, C.J., Naesje, T.F., Chanda, B. and Okland, F. 2004. Effects of catch-and-release angling on large cichlids in the subtropical Zambezi River. Fisheries Research, 69, 141-144.

Tinsman, J.C. and Whitmore, W.H. 2006. Aerial flight methodology to estimate and monitor trends in fishing effort on Delaware artificial reefs sites. Bulletin of Marine Science, 78(1), 167-176.

Toivonen, A.L., Roth, E., Navrud, S., Gudbergsson, G., Appelblad, H., Bengtsson, B. and Tuunainen, P. 2004. The economic value of recreational fisheries in Nordic countries. Fisheries Managment and Ecology, 11, 1-14.

Tratalos, J.A. and Austin, T.J. 2001. Impacts of recreational SCUBA diving on coral communities of the Caribbean island of Grand Cayman. Biological Conservation, 102, 67–75.

Trist, C. 1999. Recreating ocean space: recreational consumption and representation of the Caribbean marine environment. Professional Geographer, 51,376–387.

Tudela, S. 2004. Ecosystem effects of fishing in the Mediterranean: an analysis of the major threats of fishing gear and practices to biodiversity and marine habitats. Studies and Reviews. General Fisheries Commission for the Mediterranean. No.74. Rome, FAO.

Tudela, S., Kai Kai, A., Maynou, F., El Andalossi, M. and Guglielmi, P. 2005. Driftnet fishing and biodiversity conservation: the case study of the large-scale Moroccan driftnet fleet operating in the AlboranAlborán Sea (SW Mediterranean) Biological Conservation, 121(1), 65-78.

Underwood, A.J. 1981. Techniques of analysis of variance in experimental marine biology and ecology. Oceanography and Marine Biology. An Annual Review, 19, 513-605.

Underwood, A.J. 1997. Experiments in ecology: their logical design and interpretation using analysis of

155

____________________________________________________________Bibliography

variance. Cambridge University Press, Cambridge, U.K., 504 pp.

UNEP/MAP/RAC/SPA, ACCOBAMS, IUCN, WWF MedPO, WWF MedPAN. 2009. Supporting the development of a representative, effective network of MPAs in the Mediterranean Sea. 15th UNEP Conference of Parties to the Barcelona Convention Almería, 16 January 2008.

Uyarra, M.C. and Côté, I.M. 2007. The quest for cryptic creatures: impacts of species-focused recreational diving on corals. Biological Conservation, 137, 77–84.

Vahl, O. 1980. Seasonal variations in seston and in the growth rate of Icland scallop, Chlays islandica (O.F. Muller) from Balsfjord, 70 N. Journal of Experimental Marine Biology and Ecology, 48, 195-204.

Valle, C., Bayle, J.T., Ramos, A.A. 2002. Weight-length relationships for selected fish species of the western Mediterranean Sea. Journal Application Ichthyology,19, 261-262.

Vandermeulen, H., 1998. The development of marine indicators for coastal zone management. Ocean and Coastal Management, 39, 63-71.

Vigliano, P.H., Lippolt, G., Denegri, A., Alonso, M., Macchi, P. and Dye, C.O. 2000. The human factors of the sport and recreational fishery of Sna Carlos de Bariloche, Rio Negro, Argentina. Fisheries Research. 49, 141-153.

Vigliano,P.H. and Lippolt, G.E. 1991. El factor humano de la pesca deportiva

y recreacional de salmónidos en el lago Fonvk, Provincia de Rio Negro, Argentina. Medio Ambiente, Chile. 11, 69-78.

Vogel, S. 1994. Life in moving fluids. The physical biology of flow. Princeton University Press, Princeton.

Vredenburgh, A.G. and Cohen, H.H. 1993. Compliance with warnings in high risk recreational activities: skiing and SCUBA. In Proceedings of the 37th Annual Meeting of the Human Factors and Ergonomics Society, 2. Human Factors and Ergonomics Society, Santa Monica, California, pp. 945–949.

Walker, D.I., Lukatelich, R.J., Bastyan, G., McComb, A.J., 1989. Effect of boat mooring on seagrass beds near Perth, Western Australia. Aquatic Botanic, 36, 69-77.

Walters, R. and Samways, M. 2001. Sustainable dive ecotourism on a South African coast reef. Biodiversity and Conservation, 10, 2167–2179.

Ward, P., Myers, R.A. 2005. Shifts in open-ocean fish communities coinciding with the commencement of commercial fishing. Ecology, 86(4), 835-847.

Werner, E. and Werner, B., 1954. Über den Mechanismous des Nahrungserwerbs der Tunicaten, speziell der Ascidien. Helgol. Wiss. Meerresunters, 5, 57-92.

Westera, M., Lavery, P., Hyndes, G. 2003. Differences in recreationally targeted fishes between protected and fished areas of a coral reef marine park.

156


Journal of Experimental Marine Biology and Ecology, 294, 145-168.

White, A.T., Vogt, H.P. and Arin, T. 2000. Philippine coral reefs under threat: The economic losses caused by reef destruction. Marine Pollution Bulletin, 40, 598-605.

Wielgus, J, Chadwick-Furman, N.E. and Dubinsky, Z. 2004. Coral cover and partial mortality on anthropogenically impacted coral reefs at Eilat, northern Red Sea. Marine Pollution Bulletin, 48, 248-253.

Williams, I.D. and Polunin, N.V.C. 2000. Differences between protected and unprotected reefs of the western Caribbean in attributes preferred by dive tourists. Environmental Conservation, 27, 4:382-391.

Winer, B.J. 1971. Statistical principles in experimental design, McGraw-Hill, Tokyo.

Winer, B.J., Brown, D.R., Michels, K.M., 1991. Statistical Principles in Experimental Design. McGraw-Hill, New York.

Wong, G. (1997). The need for economic valuation of marine and coastal tourism benefits in Malaysia. Report produced under Project MY 0062. WWF Malaysia, Petaling Jaya.

Wong, J.L.P. 1996. Marine Parks Malaysia: Tourism, impact and conservation awareness. Maritime Institute of Malaysia (MIMA), Kuala Lumpur.

Wong, P.P. 1998. Coastal tourism development in Southeast Asia : relevance and lessons for coastal zone management. Ocean and Coastal Management, 38 (2), 89-109.

Wood, L., Fish, L., Laughgren, J. and Pauly, D. 2008. Assessing progress towards global marine protection targets: shortfalls in information and action. Oryx 42, 340–351.

Worm, B., Edward, B.B., Beaumont, N., Duffy, J.E., Folke, C., Halpern, B.S., Jackson, J.B.C., Lotze, H.K., Micheli, F., Palumbi, S.R., Sala, E., Selkoe, K.A., Stachowicz, J,.J. and Watson, R. 2006. Impacts of biodiversity loss on ocean ecosystem services. Nature, 314, 787-790.

Worm, B., Sandow, M., Oschlies, A., Lotze, H.K. and Myers, R.A. 2005. Global patterns of predator diversity in the open oceans. Science, 309(5739), 1365-1369.

Wynberg, R.P. and Branch, G.M. 1997. Trampling associated with baitcollection for sandprawns Callianassa kraussi Stebbing: effects on the biota of an intertidal sandflat. Environmental Conservation, 24,139–148.

Yandell, B.S. 1997. Practical Data Analyses for Designed Experiments. Chapman and Hall, London.

Zabala, M. 1996. Impacto biológico de la creación de una reserva marina: el caso de las Islas Medes. In: Instituto de estudios Almerienses (eds) La Gestión de los espacios Marinos del Mediterráneo Occidental. Almeria,

157

____________________________________________________________Bibliography

Spain: Diputación de Almeria, pp. 55-103.

Zabala, M. and Ballesteros, E., 1989. Surface-dependent strategies and energy flux in benthic marine communities or why corals do not exist in the Mediterranean Sea. Scientia Marina, 53, 1-15.

Zakai, D. and Chadwick-Furman, N.E. 2002. Impacts of intensive recreational diving on reef corals at Eilat, northern Red Sea. Biological Conservation, 105, 179–187.

Zieman, J.C. 1976. The ecological effects of physical damage from motor boats on turtle grass beds in southern Florida. Aquatic Botany, 2,127-139.

158


Reunido el Tribunal que suscribe en el día de la fecha acordó otorgar, por a la Tesis

Doctoral de Don/Dña. Beatriz Luna Pérez la calificación de .

Alicante de de

El Secretario,

El Presidente,

UNIVERSIDAD DE ALICANTE Comisión de Doctorado

La presente Tesis de D. ________________________________________________ ha sido

registrada con el nº ____________ del registro de entrada correspondiente.

Alicante ___ de __________ de _____

El Encargado del Registro,

La defensa de la tesis doctoral realizada por D/Dª Beatriz Luna Pérez se ha realizado en las siguientes lenguas: castellano y inglés, lo que unido al cumplimiento del resto de requisitos establecidos en la Normativa propia de la UA le otorga la mención de “Doctor Europeo”.

Alicante, de de

EL SECRETARIO EL PRESIDENTE


Universitat d’Alacant

Universidad de Alicante

Documents

Anthropic impacts in Mediterranean Marine Protected Areasrua.ua.es/dspace/bitstream/10045/18846/1/master_Luna.pdfDesprés voldria agrair als meus directors per haver-me donat l’oportunitat