Recovery of red coral Corallium rubrum colonies

after fishing events in the Marseilles area

(France, NW Mediterranean)

Master’s thesis by Marina García Pizá

Master in Fundamental and Applied Ecology

20 September 2011

Project directors:

Dra. Cristina Linares Prats

Universitat de Barcelona

Dr. Joaquim Garrabou Vancells

Institut de Ciències del Mar,

CSIC

Presentation

This master’s thesis consists of an analysis of experimental data from a long-term

monitoring photographic study of red coral populations in the coast of Marseille. A

total of 351 red coral colonies have been monitored throughout seven years by

analyzing digital photographs in the video lab at the ICM, Institut de Ciències del

Mar, CSIC, Barcelona.

The project has been coordinated by MedRecover: Marine Biodiversity Conservation

Group, recognized as a consolidated research group by the Government of Catalonia’s

Agency for the Administration of University and Research Grants (Agència de Gestió

d’Ajuts Universitaris i de Recerca -AGAUR-). MedRecover studies the direct and

potential combined effects of global change on the conservation of marine

biodiversity in rocky coastal ecosystems, mainly in the Mediterranean Sea.

The master’s thesis contributes to the knowledge of the recuperation capacity of a

longevous species after a fishing event and the results will help in the improvement of

the management strategies valid until now. It forms part of the project: “Marine

conservation biology in coastal ecosystems (2009-2013)”, funded by AGAUR, and

coordinated by Joaquim Garrabou and Cristina Linares.

Marina García Pizá

Institut de Ciències del Mar, CSIC, 20 September 2011

Recovery of Red Coral Corallium rubrum colonies after fishing events in the Marseilles area (France, NW Mediterranean)

Summary

The Mediterranean red coral, Corallium rubrum, is a slow-growing longevous

gorgonian that produces a red calcium carbonate skeleton, which is being used by the

jewellery industry. Even though intensive fishing has caused a shift in size structure

for most of the populations in the Mediterranean, this species seems to persist. The

hypothesis that allows the persistence of C. rubrum despite the fishing pressure is

based on: 1) Small size at first reproduction and 2) Capacity of colonies to re-grow

new branches after being fished. From long-term photographic series monitoring two

populations in the Marseilles area, we studied the impact and recovery dynamics of

red coral after two fishing events. The results showed that >90% of affected colonies

have the capacity to recover by re-growing new branches in less than 7 years, but

colony size is far from the intact colony size before being fished. On the other hand,

scuba diving acts as an additional disturbance as it slows down this recovery rate.

Fishing practices and subsequent damage caused by diving, contribute to the

reduction of mean colony size with its corresponding effects such as decrease in larval

productivity and therefore an increase in risk of local extinction. Dramatic long-

lasting shifts in population demography highlight the need of improved management

regulations.

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Table of contents

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1. Introduction

The Mediterranean red coral, Corallium rubrum (Linaeus 1758, Gorgonacea,

Octocorallia) is a colonial sessile cnidarian with an arborescent growth form. Red

coral is an important species of the rocky bottoms of the Mediterranean and is

essentially distributed in the Western basin even though it is also present in some

areas of the Eastern basin and the neighbouring African Atlantic coast (Zibrowius

1984; Garrabou & Harmelin 2002).

Red coral is a sciaphilous species, typically associated with animal-dominated

communities growing in dimly lit habitats such as rocky cliffs, overhangs and caves

from 10 to more than 600m in depth (Laborel & Vacelet 1960; Chintiroglou et al.

1989). Water flow, food supply, temperature and light regimes are the key factors that

explain its distribution (Weinberg 1979). Its main habitat is the hard substratum in the

coralligenous where it is a characteristic species of high importance (Ballesteros

2006).

C. rubrum is an asymbiotic anthozoan species and a passive benthic suspensivore that

feeds on detritic organic particles in suspension, copepods, invertebrate’s eggs,

phytoplankton, nano- and pico-plancton (Picciano et al. 2007). Like other suspension

feeders, red coral depends on water circulation and trophic content for its energy

inputs and thus, plays an important role in pelago-benthic energy transfer processes

(Arntz et al. 1999).

Long-lived gorgonians such as red coral, play an essential role in marine ecosystems

because they structure and stabilize the ecosystem (Mitchell et al. 1993) by providing

three-dimensional complexity and increasing diversity (Jones et al. 1994). For these

reasons, these kind of species are often called ecosystem engineers. Beyond its

ecological importance, red coral is an emblematic species of the Mediterranean

because it has been harvested since antiquity for the high value of its skeleton in

jewellery and ornamentation (Santangelo et al. 1993; Morel et al. 2000).

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1.1. Red Coral fishing

The Mediterranean red coral is considered the precious coral par excellence

(Figure 1). In addition to its ornamental use in the Neolithic (3000-5000 years ago),

red coral also developed a tremendous cultural importance, as it appeared in

decorative arts of the Minoan and Mycenean civilization documents in the Bronze

Age. Red coral fishing became over-efficient when the Greeks and Arabs developed a

wooden dredging device known as "Saint Andrew's Cross" (Galasso 1998, 2001)

(Figure 2). It consisted of a wooden cross with attached nets to it that was used to

entangle red coral. Most fisheries depleted their stocks rapidly and moved on to new

ones afterwards. This "boom and bust" exploitation, more similar to coal mining than

to fishery, created unstable yields, with various peaks and drops (Santangelo et al.

1993, 2001). Today, these non-selective methods such as the "Saint Andrew's Cross"

are prohibited in all the Mediterranean countries because they cause serious damage

to benthic communities. After the 1950's, scuba diving found its application in red

coral harvesting as it allowed harvesting populations dwelling in protected crevices

and cave entrances that were inaccessible by dredging. Today, scuba diving remains

as the exclusive exploitation method of C. rubrum (Figure 3).

Almost all known C. rubrum stocks are overexploited (Santangelo et al. 2001;

Tsounis et al. 2007). The overall Mediterranean yield has decreased from ~100 t. yr-1

before the 1980's to an estimated ~25-30 t. yr-1 during recent years (Tsounis et al.

2007). Today, C. rubrum fisheries (legal and illegal) continue to impede the total

recovery of exploited populations and most populations are considered overfished

(Linares et al. 2010).

1.2. Fishing effects and conservation on red coral populations

Despite the fishing pressure, red coral is still found in the NW Mediterranean

sublittoral habitats. However, intensive harvesting has resulted in significant shifts in

the size structure of current C. rubrum populations (Santangelo et al. 1993; FAO

1998), causing a decrease in biomass and mean and maximum colony size (Garrabou

& Harmelin 2002; Linares et al. in press). Fishermen continue to harvest on already

decimated populations, resulting in a further reduction of abundance and mean size.

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Due to the history of intense harvesting, mankind has not had the chance to document

pristine red coral populations (Jackson et al. 2001). Private collections and museum

specimens have provided precious information regarding the maximum size this

species can potentially reach: up to 50cm and 4cm in height and diameter respectively

(Garrabou & Harmelin 2002). The oldest Mediterranean marine reserves harbouring

red coral are around 20 to 30 yr old (Francour et al. 2001) offering an excellent

opportunity to analyze the effects of protection on red coral populations. The limited

data available indicate an increase in maximum size and a shift towards more mature

populations in the oldest marine reserves (Tsounis et al. 2007; Linares et al. 2010).

Despite these positive effects, colony sizes do not reach characteristic values of

pristine populations (Garrabou & Harmelin 2002), most likely as a result of other

impacts such as poaching and diving frequentation, which do not allow their total

recovery (Linares et al. in press). Overall, despite the actual regulation efforts (see

below), protected and unprotected red coral populations are submitted to other

disturbances such as mass mortality events linked to climate change, sedimentation

increase or increased diver frequentation. This scenario could be driving C. rubrum

populations, especially in shallow water habitats, to situations of risk of local

extinction in the near future.

Therefore, management and conservation measures based on sound ecological data

are urgently needed. Until now, C. rubrum has been included in three international

conventions for the protection and conservation of species (Specifically in the 3rd

Annex of the Bern Convention in 1979, in the 5th Annex of the European "Habitats,

Fauna, Flora" in 1994, and in the 3rd Annex of the Barcelona Convention in 1995).

Moreover, during recent years, some attempts have been made to include the genus

Corallium in the Appendix II of CITES (Convention on the international trade in

endangered species of wild fauna and flora).

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Figure 1: Red coral colony and small pieces of jewellery in a souvenir at L’Estartit, Costa

Brava, Spain. (Photo: Marina García Pizá).

Figure 2: Non-selective fishing of red coral under a rock using the “St Andrews cross”

(Illustration: L.F Marsigli 1725).

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Figure 3 Selective fishing by scuba diving remains the exclusive exploitation method of

C.rubrum today (Photo: J.G. Harmelin).

1.3. Red coral persistence

Red coral displays high longevity (>200 years), very low growth rates (0.20 mm.

year-1 for basal diameter) and low mortality rates (<1% in absence of fishing)

(Torrents 2007; Garrabou & Harmelin 2002; Linares et al. in press). Bearing in mind

these life history traits, it seems a paradox that this slow-growing species has not

undergone extinction yet despite being exposed to intensive harvesting since ancient

times. Therefore, the question on how do these shallow populations persist and

replenish arises. A study analyzing the size and age at first reproduction of red coral,

partially contributed to answering this question. This study showed that the size of the

smallest fertile colonies is surprisingly small, only 2.4 cm in average height, which

corresponds to an age of 7-10 years (Torrents et al. 2004). This small size would have

allowed the colonies to avoid from being harvested, while they could contribute to the

reproductive output of the population and reveal at least partially, a potential

persistence mechanism.

On the other hand, red coral as other clonal organisms, has the capacity to recover

after events of partial mortality (e.g. breakage of a branch). In fact, red coral colonies

with signs of breakage and recent re-growth of new branches have been observed

(Hereu et al. 1999; Rossi et al. 2008).

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Taking into account that harvesting methods do not always harvest the whole colony,

re-growth of the parts of the colony still attached to the substrate could be a potential

source of recovery for harvested populations. However, this process has never been

studied in red coral populations and it is unclear up to what extent this mechanism

may contribute to the recovery after a fishing event. Under the framework of the study

of red coral population dynamics, two fishing events were observed, which offered

the opportunity to analyze for the first time, the recovering capacity of two

populations of the coast of Marseille.

Given this lack of knowledge over the response of this species to fishing events, the

aim of this study was to asses the capacity to recover of two C. rubrum populations of

the coast of Marseille, by analyzing long-term photographic series and quantifying

mortality and branch re-growth rates during five and seven years after each fishing

event.

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2. Materials and Methods

2.1. Study species: Reproduction, growth and mortality rates

Red coral is a gonochoric species (Lacaze-Duthiers 1864) that reproduces annually

during late spring-early summer (Santangelo et al. 2003; Tsounis et al. 2007).

C. rubrum is an internal brooder and planulae have limited dispersal capabilities

(Weinberg 1979; Bramanti et al. 2005). Studies on the genetic structure of red coral

populations have uncovered genetic differentiation at small scales (~10m),

corroborating the short dispersal of this species (Ledoux et al. 2010). The dispersal of

red coral planulae is restricted to short distances because their free life is brief,

reaching 4-12 days under laboratory conditions (Weinberg 1979; Grillo et al. 1992),

but probably much less in the field (Ledoux et al. 2010). Low natatory capacity and

geonegative behaviour tend to confine larvae to the vicinity of parental colonies

within cavities (Weinberg 1979). These philopatric dispersal features are reflected in

the frequent aggregative distribution of red coral colonies (Harmelin 1985) and by the

limited gene flow between populations (Ledoux et al. 2010).

Size at first reproduction in C. rubrum is much smaller compared to other gorgonian

species. As commented above, the colony height is 2-3 cm at first reproduction

corresponding to an age of 7-10 years (Torrents et al. 2004). Despite the small size at

first reproduction, it would be large enough to reduce most mortality risks by

overgrowth (specially by sponges) regardless the age at first reproduction (Torrents et

al. 2007).

Recruitment rates of red coral vary form one site to another (different colony

abundance, density and size, crevice size and shape and exposure to currents)

(Bramanti et al. 2007; Linares et al. in press) and high recruitment pulses seem to

combine with many years of near zero recruits (Garrabou & Harmelin 2002; Linares

et al, in press) or be very regular from 5-10 colonies. year-1. dm-2 (Bramanti et al.

2005). Nevertheless, the unique two long-term studies on this species: Garrabou &

Harmelin 2002 (22-year study) and Linares et al. in press (13-year study), suggest that

in most cases, recruitment of this species is very low (<1 recruit/300 cm2), especially

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over periods longer than 5 years. This is consistent with the life-history traits, which

state that adult survival in long-lived species such as red coral, is often associated

with infrequent and low recruitment (Stearns 1992). Annual mortality rates for these

recruits are high (4% in colonies less than 5 years old) resulting form the competition

for space with other sessile organisms. Mortality rates then decrease progressively up

to a value of about 1% for colonies over 15 years old (Garrabou & Harmelin 2002).

Growth ring analysis and experimental studies have determined very slow growth

rates of this long-lived organism that reaches 50 cm of height and reaches more than

100 years of age (Garrabou & Harmelin 2002; Marschal et al. 2004). This species has

a growth rate of about 0.20 mm.year-1 in diameter and 1.78 ± 0.7 mm.year-1 in height

(Garrabou & Harmelin 2002; Marschal et al. 2004; Torrents 2007; Linares et al.

2010).

Red coral annual mortality rate is usually very low (less than 1%) when populations

are not subjected to human-induced disturbances. The principal sources of natural

mortality include competition for space with sponges and other sessile biota, removal

from the substrate due the action of boring species (such as the clionaidae family,

which perforate corals by secreting acid over the calcified skeletons), storms,

increased sedimentation of fine particles (Harmelin 1985) and predation of the

coenchyme by the small gastropod Pseudosmnia carnea and the crustacean Balssia

gasti (Abbiatti et al. 1992). However, predators of red coral are scarce (Santangelo et

al. 1993) and as in other sessile invertebrate dominated communities, predation

appears to have a minor impact on the red coral community (Garrabou & Harmelin

2002). The principal sources of anthropogenic mortality are due to professional

fishing or poaching (Hereu et al. 1999; Tsounis et al. 2007) and massive mortality

events linked to positive temperature anomalies (Garrabou et al. 2009). Another

source of anthropogenic mortality is the increase of pollutants and water from

desalination plants (Harmelin 1985; Garrabou et al. 2001).

Besides overfishing, during the last decade, climate-induced mortalities have caused

the most important impacts on red coral populations. In the NW Mediterranean area,

at the end of the summers of 1999 and 2003, two mass mortality events affected

marine rocky benthic communities on large spatial scales (more than 500 km)

(Garrabou et al. 2001). These two large-scale mortality outbreaks have been related to

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positive temperature anomalies (Garrabou et al. 2001), severely affecting red coral

populations. In the Marseilles region during the summer of 2003, 40% of the

populations of red coral were affected with a medium degree of impact (Garrabou et

al. 2009). As over-fishing could now be coupled with mass mortality events (i.e. due

to thermal anomalies), then the co-occurrence of these two mortality sources could

dramatically depress the recovery of red coral populations (Santangelo 2006).

2.2. Study area

The study area is the limestone rocky coast of Massif des Calanques in the SE of

Marseilles, France (Figure 4). Massif des Calanques is located in a eutrophic area near

a urbanized coast, influenced by wind-induced local upwelling (Millot et al. 1980)

and is affected by the Ligurian current. Due to the specific habitat characteristics of

the area: like submerged cavities and overhangs (Laborel & Vacelet 1960), along with

instability of the water column during summer, the development of red coral at

shallow depths (20-22m) is favoured.

The two shallow red coral populations studied are located in two different islands;

Riou Sud located in Riou Island is characterized by a vertical wall and Grotte Corail,

located in Maire Island, which consists of a cave-like tunnel. At these depths, both

sites are mainly characterized by the presence of perennial animals associated to

coralligenous assemblages (Coma et al. 1998; Garrabou & Harmelin 2002). In the

study area, red coral populations have been exploited since the medieval ages using

different kinds of techniques (Morel et al. 2000). At present, harvesting is permitted

to 10 authorised red coral fishermen and 2 of them are present in the Calanques area,

even though poaching is likely to be common. Harvesting by divers threatens mainly

the largest and most attractive colonies.

Lastly, scuba divers frequentation is almost null in Riou Sud, while Grotte Corail is

one of the most commonly frequented diving sites in the Marseille’s area, with more

than 5000 dives per year. At this site, the near presence of detached branch fragments

on the sea bottom seems to be the evidence of high diving pressure.

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Figure 4: Map of the Calanques region in the coast of Marseille. The two dots in red

represent the study sites: Riou Sud in Riou Island and Grotte Corail in Maire Island.

2.3. Photographic survey of permanent transects: Red coral

monitoring

When studying a slow-growing species, long-term photographic series are useful to

detect and quantify changes over spatial and temporal scales. During this long-term

study on the population dynamics of red coral in the coast of Marseille, two fishing

events took place, affecting the populations of Riou Sud and Grotte Corail. We

analyzed the photographic series obtained yearly during seven years.

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A permanent plot was set up in each population to monitor the colonies

photographically over time. Each plot was 2-4m long and PVC screws were attached

to the wall every 40-50cm. An elastic band was placed along each transect and around

each screw in order to guide the photographer through the transect. The pictures were

taken above and below the band following the transect. A fixed frame was used to

ensure a same study area in each photograph of 20x20cm. This method allowed a

monitoring of each colony individually throughout the years.

In Riou Sud, two permanent plots were established at 20 and 22m in 2005.

Photographs were analyzed from 2005 to 2011. Throughout the monitoring of the

photographs, a fishing event was detected between April and June 2006, therefore a

comparison could be done before and after the fishing event. Moreover, the fished

colonies could be identified.

In the Grotte Corail population, two permanent transects were established at 15 m just

after a fishing event was observed during June 2002. Photographs were analyzed from

2002 to 2009, but due to logistic constraints, pictures were not taken during 2007 and

2008. Among the colonies observed in 2002, some were half broken with hanging

broken branches, as well as recent wounds with tissue coming out. This gives us the

evidence that the colonies had just been fished. However, we have no before-after

information on this site because the plots were set up just after the fishing event.

2.4. Photographic analysis: Survival and re-growth study

All the analyzed colonies were individually identified and followed annually

throughout the study period. Mortality and branch re-growth was surveyed on

photographs corresponding to 1-year intervals whenever it was possible. A total of

351 colonies were monitored during the 7-year study: 211 colonies in Riou Sud and

140 in Grotte Corail. After the fishing event in each site, the studied colonies were

divided into two groups; affected and unaffected by fishing. This allowed us to

explore potential differential mortality traits among groups.! For each colony, we

determined the type of mortality suffered or not and the presence and number of new

branches. After the fishing event, the affected colonies were monitored not just to

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study new branch re-growth but also to see if being affected by fishing in previous

years, could influence into a future mortality.

The following parameters were defined from the photographs:

Re-growth of a branch was defined as the first bump observed on the base of the

affected colony. Figure 5 shows the evolution on new branches.

Partial mortality was defined when a part of the colony was lost between the time

intervals considered. After the fishing event, the colony base could either have a

smooth surface if no branches were present (e.g. after the fishing event) or a rough

surface when a branch was beginning to grow or some branches had been left (e.g.

due to a more conscious fishing extraction). Figure 6 shows an example of partial

mortality.

Total mortality was defined when the colony disappeared and was not seen again in

the following years. Total mortality could either be due to type of extraction when the

whole colony was removed with the substratum still attached to the base, or to the

death of the colony some time after the extraction (e.g. the colony base is epiphyted

by sponges and eventually disappears). Figure 7 shows an example of total mortality.

Figure 5: Branch re-growth. The colony on the left shows two small bumps and a clear

branch, while the colony on the right shows a group of higher developed branches.

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Figure 6: Partial mortality: The circles indicate a smooth base for each harvested colony.

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Figure 7: Total mortality: The circles show the three missing colonies on the left, which

disappear after the fishing event in 2006.

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Figure 8: Evolution of fished colonies in Riou Sud from 2005 to 2007 (top to bottom). The

fishing event is clearly observed in 2006, where partial mortality is shown for all the colonies.

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2.5. Statistical analysis

The differences between affected and unaffected colonies and the differences between

sites were analyzed using non-parametric analysis, since our data violated normality

and homocedasticity.

A Kolmogorov-Smirnov two-sample test was used to compare affected and

unaffected colonies between the two sites, and to compare the average number of

branches in each site. The data representing the number of branches in affected

colonies contained a large number of zero values, especially during the first two years

of monitoring, when many colonies still had no branches. Even when the data were

log transformed, the data did not follow a normal distribution. Thus, an ANOVA

repeated measures test could not be applied in order to compare differences between

sites and years. Finally, a non-parametric Kolmogorov-Smirnov two-sample test was

applied in order to compare the number of branches between sites during the last year

of monitoring (five and seven years after the fishing event). Meanwhile, to compare

affected and unaffected colonies as well as the type of mortality suffered by the

colonies within the same site, a Mann-Whitney U test was applied.

3. Results

3.1. Fishing impact

Figure 9 shows the percentage of affected and unaffected colonies by a fishing event

in each site. The fishing event that took place in Riou Sud during 2006 affected 38%

of the colonies within the transect. Meanwhile the fishing event in Grotte Corail

during 2002, affected 57% of the colonies. The differences among affected and

unaffected colonies between sites were not significant (Kolmogorov-Smirnov test,

p=0.10). The type of mortality caused by fishing was mainly partial mortality.

Between 90-100% of the colonies were affected by this type of mortality in both sites

(Figure 10). Total mortality represented only 7% of the affected colonies in Riou Sud,

meanwhile, it was not present in Grotte Corail.

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Figure 9: Percentage of red coral colonies affected and unaffected by fishing in Riou Sud

during 2006 and in Grotte Corail during 2002. N is the number of colonies studied in each

site.

Figure 10: Percentage of affected colonies by fishing that suffer partial mortality and total

mortality. N is the number of affected colonies in each site.

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3.2. Long-term mortality rates of affected and unaffected colonies

3.2.1. Mortality rates of colonies affected by a fishing event

Colonies affected by fishing in Riou Sud showed very low values of new mortality

“post-affection” from 2007-2011: only 1.2% of the affected colonies suffered total

mortality (Figure 11 a). In contrast, the monitoring from 2003-2009 in Grotte Corail

showed a higher percentage of affected colonies suffering from mortality “post-

affection”: approximately 15% of the affected colonies showed partial mortality and

1.2% total mortality (Figure 11b).

3.2.2. Mortality rates of colonies unaffected by fishing

In Riou Sud, a small percentage (5%) of unaffected colonies showed partial mortality

and no total mortality was observed during 2007-2011 (Fig 11a). However, in Grotte

Corail, up to 60% of the unaffected colonies suffered partial mortality, while total

mortality was recorded on 4% in these colonies (Figure 11b).

3.2.3. Comparison between affected and unaffected colonies

In Riou Sud, partial mortality was significantly higher in unaffected colonies (Mann-

Whitney U test, p<0.05). Meanwhile, differences in total mortality between affected

and unaffected colonies were not significantly different (Mann-Whitney U test, p

0.86). However, in Grotte Corail, both differences in partial and total mortality

between affected and unaffected colonies showed significant differences (Mann-

Whitney U test, p<0.05).

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Figure 11 a) New mortality of affected and unaffected colonies by fishing in Riou Sud after

the fishing event from 2007-2011. b) New mortality of the affected and unaffected colonies

by fishing in Grotte Corail after the fishing event from 2003-2009.

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3.3. Colony recovery from fishing impact !

All the colonies that survived the fishing event show branch re-growth during the

seven-year monitoring. Figure 12 shows the percentage of affected colonies with

presence of new branches over the study time. A small percentage of affected

colonies died in the following years after the fishing event and therefore, branch

presence was not reached in 100% of the affected colonies.

During the first year after the fishing event, colonies in both sites started growing new

branches. The greatest difference in branch presence was during the first year after the

fishing event, in which Riou Sud had 75% of its colonies with presence of new

branches, while Grotte Corail showed new branch presence in 15% of its colonies

(Figure 12). During this first year, colonies in Riou Sud had from one to four branches

as well as a small percentage (15%) of colonies with more than four branches (Figure

14 a). Meanwhile, Grotte Corail had colonies ranging form one to three branches

(Figure 14 b).

The overall number of branches marked another difference between sites. In Grotte

Corail, colonies with 4 branches were less abundant and were not observed until 4

years after the fishing event, whereas in Riou Sud, these colonies were already present

since the first year after the event. Five years after the event in Riou Sud, 55% of its

colonies grew more than four branches compared to seven years after the event in

Grotte Corail, where only 35% of its colonies grew more than four branches. Both

sites reached a similar degree of colonies with branch presence, even though there

were differences between the number of branches per colony.

Figure 13 shows the evolution of the average number of branches per colony during

the consecutive years after the fishing event for each site. Riou Sud displayed a higher

mean number of branches throughout the study compared to Grotte Corail, reaching

an average of six branches per colony, while Grotte Corail reached an average of four

branches per colony. The average number of branches between sites; five years after

the event for Riou Sud and seven for Grotte Corail, was significantly different

(Kolmogorov-Smirnov test, p<0.05).

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Figure 12: Percentage of affected colonies with presence of new branches, five and seven

years after the fishing event in Riou Sud and in Grotte Corail respectively. Note that there is

missing data in Grotte Corail during 2007 and 2008.

Figure 13: Average (±SD) number of branches per colony during five years after the fishing

event for Riou Sud and seven for Grotte Corail.

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Figure 14: Number of branches in affected colonies in Riou Sud (a) and Grotte Corail (b)

during the consecutive years after each fishing event.

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4. Discussion

This study showed that most colonies (>90%) could recover from a fishing event

when fishermen did not harvest the whole colony. Between five and seven years were

necessary to observe at least one new branch in colonies affected by the fishing

events. However, recuperation of the affected colonies at the end of the study was not

complete. This could be observed in Riou Sud where colony size after five years was

far away from the size of intact colonies before the event. Nevertheless, quantitative

measurements would be necessary to establish a more precise recuperation status.

Previous studies have shown the slow growth rates of red coral colonies being around

0.20 mm.year-1 for basal diameter and 1.78 mm.year-1 for colony height (Garrabou &

Harmelin 2002, Torrents 2007, Linares et al. 2010). Bearing in mind these growth

rates and the size of the larger colonies found in studied populations (97.2 mm in

height, Garrabou & Harmelin 2002), it is clear that a full recovery will require at least

60 years of effective absence of harvesting.

Going back to the paradox commented in the introduction, concerning the persistence

of a slow dynamic species such as red coral despite being exposed to intensive

harvesting since ancient times, we could consider the re-growth of new branches as a

meaningful potential mechanism to explain today’s abundance of populations. This is

clearly illustrated by comparing the growth dynamics of colonies either through re-

growth of new branches in adults or by growth patterns observed in sexual recruits,

which is the alternative mechanism available for the recovery of populations. With

reference to the mean number of branches, our results showed that five and seven

years after fishing, colonies had a mean number of 5 branches. Meanwhile, the

monitoring of colonies recruited on experimental panels showed about 3 branches

after 22 years of growth (Garrabou & Harmelin 2002). Branch growth rates are 8

times higher in affected adult colonies than in recruits. Moreover, mortality rates in

recruits are remarkably high (4% in colonies less than 5 years old, Garrabou &

Harmelin 2002) compared to mortality rates observed in affected adult colonies

quantified in this study (0.30%). This contrast between branch growth and mortality

in recruits and adult colonies should be taken into account when studying

recuperation capacity.

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Differences in partial mortality rates found in this study between Riou Sud and

Grotte Corail could be attributed to the difference in diver frequentation in both sites

(see Material and methods). A recent study based on 13 years of red coral monitoring,

showed a significant negative correlation between the annual number of recreational

dives and the basal diameter of red coral colonies, suggesting that divers may increase

the mortality rates of the largest red coral colonies (Linares et al. in press) and

therefore impede the total recovery of the affected colonies. In our study, despite the

high scuba diving pressure over Grotte Corail, the affected colonies seem to be

“protected” due to their small size and are thus, less susceptible to breakage by divers

in contrast with the unaffected colonies which are larger in size. Overall, we

demonstrated differences between sites by quantifying partial mortality rates due to

mechanical breakage of branches, which occurred at higher rates in frequented diving

sites and slowed down recovery rates. In fact, red coral colonies in Riou Sud have a

higher mean height (37.8mm) than the colonies in Grotte Corail (26.5mm) (Linares et

al. 2010).

The conservation of Mediterranean red coral is a worldwide concern as it has been

proved by the recent struggle to include the Genus Corallium in the Appendix II of

CITES (Bruckner 2009, Tsounis et al. 2009). This attempt failed because

Mediterranean red coral cannot be considered an endangered species due to the

extensive distribution of dense populations in the Mediterranean Basin. It should be

taken into account that using density and abundance seems an incorrect assessment of

the state of this species, because changes in these parameters are inadequate measures

of the decline for colonial organisms such as C. rubrum, as they exhibit a size-based

exponential increase in reproductive output (Bruckner 2009).

New management and conservation plans should focus on enhancing the persistence

of populations. In the case of red coral, it seems urgent to promote the increase of the

abundance of large colonies, as these colonies have enhanced survival likelihood and

thus increased future reproductive success (Szmant-Froelich 1985). For example, in a

C. rubrum population, a young colony produces only tens of planulae, while an old

and large colony, produces more than 2000 planulae (Santangelo et al. 2003).

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Red coral populations have been historically submitted to harvesting and have coped

with the effects of this strong disturbance. However, disturbances such as the increase

in diver frequentation, mass mortality events linked to climate change and

sedimentation increase (Ballesteros 2006; Garrabou et al. 2009; Linares et al. in

press), may act synergically jeopardizing the persistence of red coral populations.

Overall, this new disturbance regime could reduce the capacity of colonies to recover.

From this study, it is clear that new management regulations which pursue a

sustainable fishery should contemplate measures such as; at least, leaving the colony

base intact and if possible, leaving some branches attached to the base, in order to

help the recovery of the affected colonies. Another point to consider is to limit the

access of scuba divers to sites were red coral dwells, such as submarine caves, which

represent the most vulnerable habitat to scuba diving activity. Combining both

recommendations and applying a reasonable rotational harvest (over periods of

several decades) to allow the stocks to recover before re-harvesting, would be a

precautionary approach for the protection of red coral and would facilitate the

recovery of the populations.

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5. Conclusion

Fishing practices and subsequent damage caused by diving, contribute to the

reduction of mean colony size with its corresponding effects such as decrease in larval

productivity and therefore an increase in risk of local extinction.

Re-growth of new branches acts as an important mechanism in the recovery of

affected colonies, allowing the persistence of red coral despite the harvesting

pressure. Scuba diving activities can affect branch re-growth rate, and thus slow down

the recovering capacity of the colonies. Future management plans should recommend

fishermen to keep the colony base intact and if possible leave at least half of the

branches, which will ensure a faster and more robust recovery of the colonies. On the

other hand, imposing a strict regulation plan over diving activities, especially in

vulnerable sites, will ensure more efficient protection plans.

Acknowledgements

I would specially like to thank Cristina Linares and Joaquim Garrabou for giving me

the opportunity to be part of their research group and for their support during these

months. I would also like to thank my video lab mates for sharing such a pleasant

atmosphere of work and for their advice while writing this project. And finally, thank

Luis for his help with my difficulties with the computer and for his infinite patience.

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