Ecological and Socio-Economical Aspects of Artificial Reefs

Ecological and Socio-economic Aspects of Artificial Reefs

Thesis submitted in partial fulfillment of the requirements for the degree of “DOCTOR OF PHILOSOPHY”

Omer Polak

Submitted to the Senate of Ben-Gurion University of the Negev

23.11.2022

Beer-Sheva

Eilat Campus, Eilat

Ecological and Socio-economic Aspects of Artificial Reefs

Thesis submitted in partial fulfillment of the requirements for the degree of “DOCTOR OF PHILOSOPHY”

Omer Polak

Submitted to the Senate of Ben-Gurion University of the Negev

Approved by the advisor, Dr Nadav Shashar

Approved by the Dean of the Kreitman School of Advanced Graduate Studies

23.11.2022

Beer-Sheva Eilat Campus, Eilat

This work was carried out under the supervision of Dr. Nadav Shashar

In the Department of Life Sciences. The program in Marine Biology and Biotechnology, Eilat Campus

Faculty of Natural Sciences

Research-Student's Affidavit when Submitting the Doctoral Thesis for Judgment

I, Omer Polak, whose signature appears below, hereby declare that:

 I have written this Thesis by myself, except for the help and guidance offered by my Thesis Advisors.

 The scientific materials included in this Thesis are products of my own research, culled from the period during which I was a research student.

 This Thesis incorporates research materials produced in cooperation with others, excluding the technical help commonly received during experimental work. Therefore, I am attaching another affidavit stating the contributions made by myself and the other participants in this research, which has been approved by them and submitted with their approval.

Date: 23.11.2012 Student's name: Omer Polak Signature:

Table of Contents

List of Figures ...... V List of Tables ...... VII 1. Introduction ...... 12 1.1 Man and the marine environment ...... 12 1.2 Marine restoration...... 12 1.3 Socio-economics of artificial reefs ...... 16 1.4 Rationale and hypotheses ...... 18 2. Fish community structure following habitat relocation ...... 22 2.1 Introduction ...... 22 2.2 Methods ...... 24 2.3 Results ...... 27 2.4 Discussion ...... 32 3. Habitat-dependent movements of a coral-dwelling fish ...... 36 3.1 Introduction ...... 36 3.2 Methods ...... 38 3.3 Results ...... 44 3.4 Discussion ...... 47 3.5 Conclusions ...... 49 4. AR with coral transplants: a conservation tool for fish enhancement? A cases study from Eilat Red Sea...... 50 4.1 Introduction ...... 50 4.2 Methods ...... 52 4.3 Results ...... 55 4.4 Discussion ...... 61 5. Can transplantation of corals on ARs in high pressure dive sites succeed? A case study from Eilat, Red Sea...... 65 5.1 Introduction ...... 65 5.2 Methods ...... 68 5.3 Results ...... 72 5.4 Discussion ...... 80 5.5 Conclusions ...... 82 6. Economic value of biological attributes of artificial coral reefs ...... 84 6.1 Introduction ...... 84

III

6.2 Methods ...... 86 6.3 Results ...... 92 6.4 Discussion ...... 96 7. Can a small artificial reef reduce diving pressure from a natural coral reef? ...... 99 7.1 Introduction ...... 99 7.2 Methods ...... 102 7.3 Results ...... 106 7.4 Discussion ...... 110 7.5 Conclusions ...... 112 8. Discussion ...... 114 Appendix ...... 121 Bibliography ...... 132 162 ...... תקציר 164 ...... תודות

List of Figures

Figure 2.1 The study site ...... 25 Figure 2.2 Rarefaction comparison of fish diversity on the amphibian ARs in the old and new locations...... 28 Figure 2.3 Fish abundance (a) species richness (b) and (c) Fisher’s α biodiversity index of nocturnal and diurnal fish assemblages ...... 29 Figure 2.4 Multivariate analysis of fish assemblages (made with unstandardized and fourth root transformed data) showing similarity differences of fish community composition in the different ARs and NR environments ...... 31 Figure 3.1 The study area (a). The experiment was conducted in areas (1) and (2)...... 39 Figure 3.2 An outline of the experimental design and the measurement of current vs. fish movement...... 42 Figure 3.3 Distribution of number of passes as a function of distance between two random corals in the natural reef ...... 44 Figure 3.4 Average number of passes (±SE) between experimental corals at different distances at high (●), artificial (○), and low (▼) complexities ...... 45 Figure 3.5 Probability to move between corals vs. distance between corals in the low, high, and artificial habitats ...... 46 Figure 4.1 Study site and location of AR deployment ...... 53 Figure 4.2 Design of the AR ...... 53 Figure 4.3 Fish count of the abundance of fish (blank circles) and species richness (black circles) on the reef throughout the monitoring period in all the fish (a) and in only juveniles (b) ...... 57 Figure 4.4 Representation of common fish abundance colonizing the AR ...... 58 Figure 4.5 Comparison of (a) species richness and (b) fish abundance after AR deployment between the AR (square) and two nearby coral outcrops ...... 59 Figure 4.6 An nMDS repeated measures analysis of fish composition on the AR (empty icons) and adjacent coral outcrops A (black triangle) and B (grey triangle) over time ...... 60 Figure 4.7 An nMDS repeated measures analysis of fish composition on the AR over time ...... 61 Figure 5.1 Study site and location of AR deployment ...... 69 Figure 5.2 Design of the AR ...... 69

Figure 5.3 Survivorship (a-c) and number of dead and missing corals (e-g) of branching corals in transplantation events 1, 2, and 3 conducted on the flat surfaces of the AR ...... 73 Figure 5.4 Comparison of (a) dead and (b) missing corals in branched and massive corals in the third transplantation...... 76 Figure 5.5 Density accumulation of coral recruits (±SE) onto the AR in 1 m2 using fluorescent light methodology (see text) in correspondence to monthly sea surface temperature in Eilat (±SD)...... 78 Figure 6.1 Representation of scenario B (coral and fish abundance) at (a) no conservation effort and at (b) low, (c) medium, and (d) high conservation efforts...... 87 Figure 6.2 Representation of the different biological scenarios ...... 89 Figure 6.3 Median willingness to pay for different biological scenarios ...... 95 Figure 7.1 Map of study site ...... 102 Figure 7.2 Design of the artificial reef (AR) ...... 104 Figure 7.3 Percent frequency of diving visitations inside the marine protected area (MPA) before (n = 35) and after (n = 30) AR deployment...... 107 Figure 7.4 Time spent in the vicinity of the artificial reef (AR) and adjacent outcrops A and B (Mean ± SD) (ANOVA p > 0.05)...... 108 Figure 7.5 Comparison between the numbers of divers hr-1 per dive type on outcrop A and on the AR...... 109

List of Tables

Table 2.1 Fish abundance, richness, and Fisher’s α diversity index, for the old and new locations of the amphibian ARs (n = 3)...... 27 Table 2.2 Contribution to species composition of the five most prominent species on fish of the amphibian ARs in their old and new locations using SIMPER analysis ...... 30 Table 2.3 An ANOSIM pair-wise comparison (r values) between the different natural and artificial habitats ...... 32 Table 3.1 Home range attributes of Dascyllus marginatus ...... 47 Table 4.1 List of corals reared on the designed coral nursery and later transplanted on the artificial reef ...... 54 Table 4.2 Physical and biological properties of the three examined outcrops: A, B, and the AR.56 Table 5.1 Details of the three transplantations including percent survival of coral at end of monitoring period...... 74 Table 5.2 Attached corals on the AR at flat and curved surfaces ...... 79 Table 6.1 Description of the different scenarios (A-G) presented to the participants ...... 90 Table 6.2 Descriptive statistics of the interviewed participants ...... 93 Table 7.1 Descriptions of dive types and their frequencies in the study area...... 103 Table 7.2 Physical and biological properties of the most visited structures in the study area. ... 105 Table 7.3 Statements expressed in four unsolicited letters sent by leading dive center managers (A-D) regarding the planned AR ...... 112 Table A.1 The location of an artificial reef possibly has an effect on the fish community structure that was formerly related to distance from the natural reefs, isolation, and the heterogeneity of the local habitat ...... 121 Table A.2 Home range in Dascyllus marginatus was previously thought to be constrained to a few meters (<3 m) from its obligatory coral ...... 124 Table A.3 Artificial reefs interact with the environment in which they are found ...... 126 Table A.4 Coral transplantation on artificial reefs is an uncommon practice and the procedure and the benefits of such restoration process are still not known ...... 128

VII

Abstract

The development of rehabilitation and restoration tools for coral reefs has been advocated over the years. One tool used for restoration is the deployment of Artificial Reefs (ARs). ARs may serve as new and sometimes surrogate habitats for the declining natural reefs and provide home for fish, corals, and other marine organisms. But, the insertion of an AR does not end in making a change in the marine environment. ARs also provide socio-economic services for coastal users including the diving industry. It was my intention to observe the ecological and social-economic aspects of ARs as one unit, and to understand patterns that link the two disciplines into a combined framework. The rationale was to find ways that an AR could be made sustainably successful while catering to the needs of fish, corals, and humans (divers and snorkelers), while concurrently diverting divers from nearby natural reefs. To understand the relationship between the AR structure and the location it is placed, I monitored the relocation of four ARs from a sandy patch close to a degraded natural reef to a sandy flat area with nearby sea grass beds. Fish were censused on the ARs, the sandy areas, the natural reef, and the sea grass beds before and after AR relocation. An increase in abundance, richness, and biodiversity of the fish was noticed. Richness was dependent on the increase in abundance. The composition of fish changed between the two locations, and the new location exhibited more diurnal fish than the former one. It was also found that the fish community on the ARs was different from that on the natural reef. As change in fish community was evident in relation to geographic location, I examined the fish behavior response to changes in coral distances on a small scale. I manipulated a set of two Stylophora pistillata corals with a pair of their coral dwelling fish Dascyllus marginatus, both very common species in the study area. The detached corals were located in the gravel area in front of the Marine Lab in Eilat. The distance between the corals was randomly changed and the number of passes (that do not include foraging bouts) was counted. The same experimental setup was conducted with the addition of artificial objects to augment complexity and in a sandy patch within the natural reef. Results showed that the fish decreased their passes between the corals as the distance between the corals increased. Similar results were observed between coral pairs in natural reef settings. Additionally, the number of passes was similar in the reef location and the augmented complexity experiments, and was different from the exposed gravel area. In

VIII

general, D. marginatus restricts its travels to 50–60 cm and further travels are rare. A one-month observation of tattooed fish in coral concentrates in the natural reef has shown that most (42%, n=40) of the D. marginatus had low activity levels (1–10% change in changes in location, n=31 samplings) and show that the fish reside mainly in one coral. A small (six units of 2×2×2 m each) concrete AR was deployed in the most heavily dived location in the coral reef of Eilat, Red Sea, and in close proximity to a thriving natural reef. Corals were transplanted on the AR to accelerate fish colonization and increase attractiveness to divers. The AR was monitored for (1) fish colonization and community composition, (2) natural coral recruitment and survival of coral transplanted on the AR, and (3) the ability of the AR to change diver’s behavior. Fish colonized quickly to the AR and surpassed in numbers those on the nearby natural coral outcrops within 063 days of deployment. The main factor that controlled fish composition was time, and it appeared that the fish composition developed in a stepwise rather than continuous manner, with a defined change attributed to a major fish recruitment event 17 months after deployment, until reaching a semi-stable state. Coral transplantation increased fish abundance and richness but did not change the overall fish community composition. Transplanted corals were severely damaged by divers who caused 46.7% of the damage to corals in the first year after transplantation, by dislodgement. Natural death was high, especially after transplantation, but in similar scale to that known in other transplantation experiments. Natural coral recruitment was constant and high (16.25±3.5 SD recruits/m2yr-1). When examined all together, divers did not show a change in behavior over time to AR development or coral transplantation. A minor change (a drop from 20–40% of time in reserve to 10% after AR deployment) was found in the category of divers spending a short amount of time of their dive in the nature reserve. These divers are suspected to be diverted out of the reserve as a result of the AR placement. No effect was attributed to coral transplantation. The amount of money divers are willing to pay to maintain a range of conservation levels in AR settings was evaluated. A survey was delivered to divers on the beaches of Eilat during a holiday vacation. The interviewees were presented with pictures of an AR with different amounts of fish and corals that were manipulated using an image processing program. The interviewees were asked to note the amount of money they were willing to pay for conserving four levels of biodiversity (no biota, low, medium, and high numbers of organisms) and in seven different scenarios (such as increase in corals only, increase in fish only, increase in both fish and corals).

Divers were able to detect the levels of conservation but only partially differentiate between the different scenarios. They were able to detect higher levels of biodiversity and ranked the fish- only scenario as least favorite. This research investigated the links between ecological and socio-economic aspects of AR deployment. With the rising trend of inserting ARs, it is important to understand the confinements in which ARs need to function, and their requirements according to the desired targets of the deployment. Sustainable integration of the ARs conservation and human appreciation functions is desirable; yet these two may be interfering with each other. Increased knowledge, understanding the needs and function of each of the reefs' parameters, and proper use of ARs with coral transplantation, allows for bridging between the two approaches.

Key Words: Conservation, Red Sea, artificial reefs, transplantation, restoration, biodiversity, community composition, habitat contrast, natural reef, habitat complexity, home range, risk, safety, shelter, predation, colonization, Eilat, succession, damage, impact, recruitment, behavior, contingent valuation, diving, willingness to pay, management.

1. Introduction

1.1 Man and the marine environment

Coral reef ecosystems supply numerous resources beneficial to man. The ecosystem services coral reefs provide include increased biodiversity, protected shores, food resources, and social-cultural (tourism) benefits to humans (reported elaborately in Costanza et al., 1997 and Moberg and Folke, 1999). These amenities have high monetary value to mankind (Costanza et al., 1997). For example, economical benefit of Hawaiian coral reefs alone was estimated at 330 million a year (Cesar and Beukering, 2004). Coral reefs are declining around the globe (Wilkinson, 2004) as a results of both natural and anthropogenic stressors (Hoegh-Guldberg, 1999; Hughes et al., 2003). These stressors can be subdivided into global threats, such as global warming and ocean acidification (Hoegh-Guldberg et al., 2007), or to local scale stressors, such as overfishing, terrestrial runoff and coral breakage by SCUBA divers (Richmond, 1993; Zakai and Chadwick-Furman, 2002). As corals provide structural habitat for a range of organisms, their decline has cascading effects on reef associated fauna. It was therefore suggested that, rather than only stopping or reducing the causes of these damage factors, effort should be allocated to active restoration of local areas to achieve future marine sustainability (Rinkevich, 2005). A possible avenue for active restoration of coral reefs is the creation of novel habitats, pre-designed to accommodate for both ecosystem needs (biodiversity, etc.) and mankind’s socio- economic needs (Rosenzweig, 2003). Such restoration may result in a win-win situation. Studies addressing win-win ecology may provide decision makers with active restoration strategies, which bridge the needs of both man and nature under tight budgetary constraints.

1.2 Marine restoration

Two approaches are known for restoring reef environments. First, is a passive approach, which claims that the environment should be left untouched, while the stressors on the environment should be mitigated. Second, is an active approach, which claims that if the reef is in a degraded state, a direct intervention should be implemented to restore it (Rinkevich, 2008). If an active approach is taken, then one of three strategies is used for ecological recovery

(Spurgeon and Lindahl, 2000; Edwards and Gomez, 2007): 1. Restoration: The act of reinstating a habitat to its original state before the inflicted impact. 2. Rehabilitation: Partial restoration of the functionality of a state to resemble the characteristics of the original habitat, but not necessarily mimicking it in full. 3. Remediation (creation): The act of repairing damage to an ecosystem or creating a new ecosystem, such as coral transplantation on previously denuded reefs.

To date, most restoration efforts are actually acts of rehabilitation of the environment (Ferse, 2008) and in some cases remediation. Coral reef rehabilitation efforts have been executed worldwide and concentrated on mitigation of ship groundings (Jaap, 2000; Bruckner and Bruckner, 2001; Ebersole, 2001; Hudson and Goodwin, 2001; Miller and Barimo, 2001; Schmahl et al., 2006), Acanthaster planci (crown of thorns ) damage (Harriott and Fisk, 1988, Hudson and Goodwin, 1997), effects of tourism (Rinkevich, 1995), coral mining, dredging and destructive fishing (Auberson, 1982; Gabrie et al., 1985; Clark and Edwards, 1994; Fox et al., 2001), and coastal pollution and development (Plucer-Rosario and Randall, 1987; Newman and Chuan, 1994; Muñoz-Chagín, 1997; Raymundo et al., 1999).

1.2.1 Artificial Reef deployment

Insertion of artificial reefs (ARs) is considered an active restoration measure (Edwards and Gomez, 2007). Artificial reefs are traditionally defined as submerged man-made structures susceptible to fouling (Svane and Petersen, 2001). A narrower definition refers to them as "submerged structure[s] placed on the seabed deliberately to mimic some characteristics of natural reefs" (Pickering et al, 1998, see also Jensen, 1997). For over 200 years there has been intentional introduction of ARs to the sea to increase fisheries catch. In present day however, ARs are designed for multiple functions, including; sea bed protection from trawl nets, attracting divers for the tourism industry, forming new habitats for biological conservation, reef restoration, reef rehabilitation, divers' mitigation off natural reefs (NRs), beach protection and scientific research (Pickering et al., 1998; Baine, 2001; Miller, 2002). Due to the rapid degradation of reefs worldwide, the use of ARs for reef restoration, by providing new habitats, is becoming more common (Perkol-Finkel and Benayahu, 2004; Miller, 2002; Pickering et al., 1998).

Most submerged ARs for tourism purposes comprise decommissioned ships and materials of opportunity, or unintentionally sunken objects (wrecks). Very few ARs, however, were pre-designed to accommodate both an increase in fish and invertebrate biodiversity and for increasing aesthetics for the purpose of attracting divers. As the need for this pre-designed ARs increases, the knowledge involved in designing, deploying, and monitoring ARs that currently serve both biological and socio-economic needs is greatly needed. Different commercially designed ARs are now commonly available. The use of Reefball™, small domed-shaped cement casted modules, has been extensively used in over 59 countries with over half a million modules placed worldwide (Reefball Foundation, 2012; www.reefball.org). Other common commercial AR companies include EcoReef™ that mimics branching corals (Moore and Erdman, 2002) and BioRock™, which uses electrolysis of seawater to enhance aragonite and brucite precipitation on iron structures (Hilbertz et al., 1977). Recent knowledge has been acquired to help design ARs that will attract fish, corals, and other invertebrates (Bohnsack and Sutherland, 1985; Perkol-Finkel and Benayahu, 2007; Spanier, 1994). Recommendations for increase in fish biodiversity include increase of complexity, high relief, void space, shaded area, and use of stable structures that will withstand storms (Kojima, 1957; Bohnsack and Sutherland, 1985; Harriot and Fisk, 1987; Rilov and Benayahu, 1998; Seaman, 2000; Edelist and Spanier, 2009). Increased attractiveness of the natural coral reef features, like increased corals and fish, may also increase attractiveness of the ARs to divers as well (Fitzhardinge and Bailey-Brock, 1989). Previous studies suggest that ARs, as popular dive sites, may reduce divers damage on natural reefs (Abelson, 2006; Leeworthy et al., 2006). The appearance of ARs can be further enhanced by transplantation of corals (Abelson, 2006). Hence, ARs can be considered visitor management tool.

1.2.2 Coral transplantation

Coral transplantation for restoration purposes is primarily used to enhance recovery of natural reefs at damaged locations. It usually results in increased coral cover and diversity, increased coral recruitment (derived from the transplanted colonies), reintroduction of coral species to damaged areas, and an increase in complexity that, in turn, serves as structure for reef-associated organisms (Abelson, 2006). Additionally, it has been suggested that coral transplants enhance the aesthetic value of an area for divers (Shinn, 1976; Fitzhardinge and

Bailey-Brock, 1989). Two approaches are used to transplant corals. (1) A one-step approach where corals are taken from one location and directly transplanted into the damaged area, usually located nearby. (2) A two-step approach in which coral fragments are drawn from natural colonies (or by collection of propagules), reared in coral nurseries and later transplanted in the target site. Regardless of the approach, transplantation efforts are usually limited to confined areas and their costs are fairly high (Ferse, 2008). Furthermore their success, measured by survivorship and growth, is not always clear (Edwards and Clark, 1998). Some studies suggest that reducing the local stress is the first measure to be done prior to transplantation (Abelson, 2006). However, at a local habitat rehabilitation scale, coral transplantation has been demonstrated to be helpful (Rinkevich, 2005; Horoszowski-Fridman et al., 2011). As ARs and coral transplants both serve as restoration measures to replace habitat loss and rehabilitate local fauna, their use in tandem is surprisingly rare. The use of ARs with coral transplants was criticized as being expensive (Ferse, 2008), time and effort consuming (Rinkevich, 2005), and not having added value, as natural coral recruits have resulted in similar coral populations over time (Edwards and Clark, 1998). Therefore, it was suggested that use of ARs with coral transplants be used only in extreme cases where natural recruitment is low (Edwards and Clark, 1998), or not be used at all (Rinkevich, 2005). In contrast, Abelson (2006) recommended the use of both methods in tandem, as it creates complex habitats for reef associated organisms and can increase the attractiveness of the site for tourism.

1.3 Socio-economics of artificial reefs

1.3.1 Economic aspects of AR deployment

In recent years, the sinking of ships to increase diving tourism has gained popularity worldwide. However, meticulous research on the cost-benefits of these deployments is scarce and complicated (Sutton and Bushnell, 2007). Ships intended for deployment as ARs for diving attraction are required to withstand rigorous checks to ascertain that they are clean for the environment and safe for diving. The cost of cleaning, preparing, and attaining all the needed permits for these ships can be very costly ranging between $46,000 to $2 million USD (Hess et al., 2001). However, even though empirical data on AR valuation is limited, many dive operators insert ARs to increase their income (Pendleton, 2004). This is partly because a one-time investment is seen as having the potential for long-term revenue. For example, the sinking of the warship USS Yukon, which cost over $400,000 to clean and sink, is expected to return an annual revenue of $5.7 million from tourism (Pendleton, 2004). These values pertain both to market (expenditure) and non-market values (see below). Introduction of ARs can increase fisheries financial output substantially, for example in Southeast Florida expenditure on ARs for fisheries and diving amounted 1.7 billion per year (Adams et al., 2009). Although traditionally ARs were deployed for fish many ARs are lately inserted for tourism purposes enhancement (Baine, 2001). Even though strict economical valuation of such acts are not well documented, various AR deployment events have documented to incur high monetary value. For example, the submersion of the USS Oriskany, an aircraft carrier, off the Florida shores was very costly ($20 million USD) but the revenue in the following years were documented to be at around $3.4 million per year for direct use and $11.5 million per year for overall economical impact (Adams et al., 2009). Similar results were exhibited in the submersion of the USS Yukon ($2.6 million per year) and Vandenberg (6.5 million increase in recreational expenditure per year: Adams et al., 2009). Economic valuation of ARs is challenging since ecosystems provide many services that cannot be bought and sold in markets. Ecological services provided by ARs may include added biodiversity, enhanced fish production and mitigation of habitat loss and more (Adams et al., 2009). To overcome the difficulties of valuating non-market products or public goods, like environmental benefits, economists have developed methods to either indirectly or directly

assess non-market value. One method commonly used is the Travel cost, where the price of the visit to a public commodity, like a coral reef, is estimated indirectly by quantifying the expenses incurred to travel and stay in the location of the attraction (Freeman, 2003). Another method used to determine directly the value of ARs is the Contingency Valuation Method (CVM). The Contingency Valuation Method estimates the value of a non-market product or service (e.g., AR) by asking people how much they are willing to pay for specific environmental services, for example the beauty of a coral reef (Lipton et al., 1995). The advantage of this methods is that it can be applied to real but also hypothetical situations that describe potential differences in the public good, like varied levels of biodiversity. These methods for estimating the value of a non- market service provided by ARs can provide additional information on the biological and aesthetic factors of ARs that people are most willing to pay for.

1.3.2 Social aspects of AR deployment

To attract divers to artificial reefs we need previous knowledge of the reasons a diver chooses to spend a dive on an AR rather than on the NR (Shani et al., 2012). Previous surveys conducted in Eilat revealed that divers preference for AR diving sites in the region are similar to natural ones. has found that AR diving sites are preferred similarly to natural ones (Shani et al., 2012). Additionally, evidence supports the notion that ARs can draw divers off NRs (Leeworthy et al., 2006). The use of marine organisms (primarily fish and corals) as reef attraction for divers was rarely addressed in the literature. In the Red Sea, South Sinai, Leujak and Ormod (2007) recorded the reasons snorkelers were attracted to natural reefs, and found that more experienced snorkelers preferred less crowded reefs with better reef health. Overall, snorkelers preferred reefs with both corals and fish the most (51%), only fish secondly (36.5%), and thirdly, reef with only corals (5.2%) and the rest were interested in other “things” on the reef. They mention that 29.8% of the visitors complained of the low number of large fish, and <16.1% of the people complained of the low abundance of fish and corals. Similar results were obtained by Wielgus et al. (2003), who found that the most important variable considered when choosing a dive site was the abundance and diversity of both corals and fish. Conversely, Shafer and Inglis (2000) found that coral size and abundance were more important than fish diversity or abundance, or fish size in dive sites located in the Great Barrier Reef. Results from these studies suggest that when selecting an AR design that will be attractive to divers, the design should foster the colonization

of a diverse assemblage of both fish and corals.

1.4 Rationale and hypotheses

To date most AR deployment initiatives that were executed for scientific purposes were examined solely on their ecological aspect, or in a few instances for their social and economic values. However, the integration of ecological and socio-economic aspects concurrently has not yet been undertaken. Therefore, my study examined the impact ARs have on the environment and the community by studying both ecological (chapters 2-5) and socio-economic (chapter 6-7) aspects of ARs deployments in Eilat, Red Sea.

Ecological section. In this section I monitored the development of marine life around and on the inserted ARs. Chapter 2 discusses the influence of distance from natural reefs on AR community composition and how relocation of ARs to new habitats (sea grass and sandy bottom) farther from natural reefs alters community composition of the AR. I examined how fish community structure, abundance, richness and biodiversity changed between old and new locations of ARs and how community composition compared to the surrounding habitat type. Specifically, I compared fish community structure, abundance, richness and biodiversity between (1) ARs and coral reefs, and (2) ARs and sandy bottom habitat. I used the principals of distance from the source population, described in the Theory of Island Biogeography (MacArthur and Wilson, 1967), to explain the community composition of reefs as distance increased between the AR and natural reefs. Contrary to the principles of Island biogeography, numerous records attest that isolated reefs are independent of distance from source population and support higher diversity indices (Shulman, 1985; Ault and Johnson, 1998; Belmaker et al., 2005). In this chapter I tested the following hypotheses:

 Fish richness, abundance and biodiversity on the AR will vary with distance of the AR from the natural reef. Here, following the Island Biogeography theory, I predicted that: o Fish richness, abundance and biodiversity will be higher at AR sites closer to the natural reef compared to AR sites further from the natural reef.

o Differences in the fish community structure between the natural reef and the AR will increase with distance of the AR from the natural reef.  Relocation of ARs to isolated sandy bottom habitat will alter the fish community structure, abundance, richness and biodiversity of the AR. Here, I predicted that fish richness, abundance, biodiversity and community structure will be higher in more isolated reef patches.

Chapter 3 investigates how AR structure and the placement of coral transplants on ARs influences fish movement patterns. In ARs where coral transplantation occurs, distance between corals may affect fish movement behavior and ultimately predation, foraging, mating and more. In this chapter, I observed the movement patterns of reef fishes between coral colonies that varied in both distance between coral patches and complexity of the surrounding environment (coral reef or artificially augmented complexity) between coral patches. In this chapter I tested the following hypotheses:

 Distance between coral patches influences the number of times fish move between coral patches. Here, I predict that fish will move less between corals separated by greater distance and the rate of movement will be determined by the size of the home range.  Habitat complexity between coral patches influences the number of times the fish move between coral patches. Here, I predicted that in high-sheltered (complex) environment fish will move more than in less complex environments.

Chapter 4 examines changes to the fish community on a small AR inserted in Eilat, Red Sea over a four year period following its deployment. In this chapter, I investigated how s fish recruitment and fish community composition on the AR changed over time and how these factors of the AR compared with natural reefs over time. In this chapter I tested the following hypotheses Time since deployment will maintain the difference in fish communities between the AR and the natural reef. Here, I predicted that long term monitoring will result in two distinct fish community compositions on the AR and the natural reefs.

 Time since deployment of the AR influences the fish composition of the AR. Here, I predicted stochastic factors, like addition of fouling organisms and recruitment events will decrease similarity in the fish community composition over time.  Coral transplantation on an AR will influence the richness and abundance of reef fishes on the AR. Here I predicted that fish richness, abundance and the similarity in the fish community composition towards natural reefs will increase following coral transplantation.

Chapter 5 describes the success of coral transplantation on an AR located in a very heavily dived location. It is documented that divers can inflict damage to natural reef corals (Zakai and Chadwick- Furman, 2002). Thus, diving activity on AR may influence the success of coral transplants on ARs. In this chapter I investigated the success of such restoration efforts on coral survivorship by measuring coral death, dislodgement by divers and evaluate the coral transplantation techniques I tested the following hypotheses:

 Species of the coral transplants will influence its overall survival and the coral’s ability to self-attach to the AR. Here I predicted that o massive corals will survive better than branching corals o the self attachment rate of branching corals will be faster. o branching corals will have lower dislodgement rate than massive corals.  Physical protection of corals from diver dislodgement will influence coral survival rates. I predicted that protected areas will reduce coral dislodgement.

Socio-economic section: In this section I tried to evaluate the relative importance of coral and fish diversity on ARs and evaluate if such ARs are cost-beneficial. I monitored the relationship between diver’s attractiveness to ARs and the ability to reduce diving pressure off natural reefs using ARs as attractions. Chapter 6 examines how different biological attributes influence the willingness to pay for protecting ARs at various biological states and in various scenarios. Here, I determined what biological attributes divers want to see when diving ARs, by making associations between biodiversity of ARs and biological attractiveness for diving. Previous evidence show that divers are able to differentiate between various ecological attributes, but the documentation is

inconsistent and the different attributes were difficult to isolate and thus evaluate. In this section I tested the hypothesis that:

 Biodiversity of the coral reef community (e.g., or biological scenario) will influence divers willingness to pay for AR, and thus the non-market value of ARs to the community. I predicted that divers will able to monetarily differentiate between different biological scenarios and will be willing to pay more for environment with high biodiversity and they will be able to detect varying degree of conservation levels.

Chapter 7 depicts social changes in the behavior of divers. It is suggested that ARs have the potential to decrease diving pressure off natural reefs (Abelson, 2006; Leeworthy et al., 2006). Yet scarce evidence support this theory. I tested the ability of a small AR to divert divers from a nearby NR. In this chapter I tested the following hypotheses :  Distance of the ARs from the natural reef will influence the intensity of the use of the natural reef for diving Here, I predicted that ARs placed in close proximity to a natural reef will reduce the dive time on the natural reef.  The use of coral transplantation on ARs will influence the time divers use the AR and the time divers use the natural reef. I predicted that the use of coral transplants on ARs will decrease dive time off nearby natural reefs.

This study aims to provide further knowledge to reef ecologists, conservationists, reef managers and stakeholders. In the light of increasing coral reef decline, new methods, approaches, and empirical evidence are necessary to devise new options for mitigating coral reef decline.

2. Fish community structure following habitat relocation

2.1 Introduction

Reef structure and its related abiotic and biotic parameters play important roles in determining reef fish species composition and abundance (Öhman and Rajasuriya, 1998; Nanamy and Nishihara, 2003; Brokovich et al., 2006). Small-scale, local attributes are also known to affect the community structures of reefs (Seaman, 2000; Baine, 2001; Belmaker, 2009). Processes that can affect fish species composition on a reef include physical characteristics like reef relief, location, connectivity, and structural complexity (Gladfelter et al., 1980; Spanier et al., 1990; McCormick, 1994; Chabanet et al., 1997; Schmitt and Holbrook, 2000; Brokovich et al., 2006; Belmaker et al., 2011) and ecological processes such as competition, predation, recruitment, migration, and disturbances (Hixon and Menge, 1991; Mora et al., 2003). Natural coral reefs differing in their physical and ecological attributes or that are situated in different surroundings, are known to have distinct fish species compositions (Sale, 2004). Micro- and meso-scale (few meters and tens-hundreds of meters, respectively) interactions between reef structure and local environment lead to the development of specific fish communities on specific reefs (Jones and Syms 1998; Holbrook et al., 2002; Grober-Dunsmore et al., 2008). Understanding these interactions is an important step in our attempt to decipher the processes shaping species composition in marine habitats (Holbrook et al., 2002; Brokovich et al., 2006). McArthur and Wilson (1967) described the Island Biogeography theory that dictates biodiversity change between isolated habitats. They claimed that the closer an isolated habitat (an island) to a mainland the more similar will be its species composition. It was therefore hypothesized that fish richness, abundance, biodiversity and fish composition will be higher at locations closer to continuous natural reef. To the contrary, it was shown that isolated reef patches support higher fish biodiversity and biomass (Shulman, 1985; Ault and Johnson, 1998; Belmaker et al., 2005) presumably as a result of reduced predation and competition (Walsh, 1985; Belmaker at al., 2005). Therefore it was alternatively hypothesized that fish richness, abundance, biodiversity and community structure will be higher in more isolated reef patches. This hypothesis is strengthened by the concept of contrast between patch structure and the local

environment. Tews et al.(2004) referred to distinct structure within the local environment (like tree in a desert) as “keystone structures” that provide vital resources, shelter and goods for the local fauna. Introduction of an AR in a location denuded of large objects may behave similarly and increase fish abundance and richness. In ecological studies, artificial reef (AR) structures can serve as surrogates for natural reef (NR) patches. While their attributes are similar to NRs, the fish community structure of ARs depends on the properties of the submerged object (Baine, 2001) and on the surrounding environment (Ambrose and Sawbrick, 1989; Burt et al., 2009). For example, it was shown that the extent of AR isolation from NRs may either increase (Walsh, 1985; Jordan et al., 2005) or decrease (Herrera et al. 2002) fish abundance on the AR, through processes such as predation (Beukers and Jones, 1997; Belmaker et al., 2005; Overholtzer-McLeod, 2006; Leitão et al., 2008a), competition (Almany, 2004a), recruitment (Ault and Johnson, 1998; Jordan et al., 2005), or translocation (Belmaker et al., 2011). In turn, an AR can also influence the community of an adjacent NR. For example, it was shown that distance from an AR affects algae grazing pressure on NRs (Einbinder et al., 2003) and that adjacent artificial structures may decrease fish abundance and biomass in natural habitats (Strelcheck et al., 2005). However, the species composition of an AR often remains distinct from that of the local NRs even after many years (Arena et al., 2007; Burt et al., 2009). For researchers, ARs provide an opportunity to experimentally test ecological questions in the marine environment (Seaman, 2000; Spanier, 2000; Bortone, 2006), as they can be designed to fit specific experimental needs, and also be moved and manipulated as desired (e.g., Belmaker et al., 2005). As such, ARs present a means to separate between structural and environmental effects on local fish communities. In this study I examined changes in fish species composition on 17-year-old ARs following the translocation of ARs from a coral reef environment to an area characterized by sand and sea grass. Fish species compositions on the ARs prior to and following reef relocation were compared to each other and with those of the surrounding natural environments. The relative impact of adjacent natural environment and the physical properties of the AR on the fish community structure of an AR were examined. As described earlier in the hypotheses, I predicted that the ARs prior to translocation to a sandy environment will host higher diversity and abundances of fish than post translocation. Alternately, lack of processes such as predation

and competition might increase both abundances and diversity in the new isolated fish communities.

2.2 Methods

2.2.1 Study sites

Four former WWII amphibian duck boats were submerged (to function as ARs) in the northern Red Sea area of the Gulf of Aqaba (29° 32' 85'' N, 34° 57' 47' E) in 1991 at depths of 22-24 m and distances of 15-30 m from a continuous NR (see Golani and Diamant, 1999; Fig 2.1b). The ARs dimensions were 11×2×1.5 m-2 (L×W×H) and are characterized by low relief (>1.5 m), low complexity (Rugosity index of 1.25, following Luckhurst and Luckhurst, 1978) and low coral coverage (11.15 ± 1.06%, mean± SD, n = 2; measured using 10 m-long line transects following Loya, 1972). After placement in May 1991, the fish communities on the ARs were monitored for three years by Golani and Diamant (1999). Between July and September 2008, about 17 years after their deployment, the ARs were relocated a distance of 3 km and placed on the sea floor at depths of 14-16 m in flat, sandy areas with mats of the sea grass Halophila stipulacea (Fig.2.1c). For their relocation, the four amphibians and other artificial objects were each lifted to a depth immediately below the surface with two 2-ton lift bags, towed to the designated immersion area, and slowly submerged (Fig. 2.1b). Very few fish (< 50) remained in contact with the ARs during their relocation (I. Zarhi and G. Zur pers comm). With the exception of two 9 year- old artificial objects (referred to here as OTARs) located 200 and 400 m away, there are no other artifacts adjacent to the new AR location. Fish diversity measurements included the OTARs, which have similar outer surface areas (21 m2 compared with 33 m2 of the ARs) but higher reliefs and structural complexities than the relocated ARs (Angel and Spanier, 2002). Previous studies found that the new AR location (AKA the northern beach) is a nursery for many fish species (Golani and Lerner, 2007).

Israel Jordan Eilat

Old New

N N 1 km

N a C2 A4 b R B A3 A1 A2 C1 A3 C2 C1 A4

R A1 N

grass Sea B

10 m A2 Damaged coral reef

Figure 2.1 The study site. (a) Old AR location (b) new AR location. A1-A4 amphibian vehicles. C1-2 metal construction beams, B- circular bellow, R- steam roller. Parallel lines indicate the location of transects. Adapted from Golani and Diamant (1999).

2.2.2 Censuses

Fish censuses of the original AR location and the designated new location were conducted two months (May-June, 2008) before the ARs were relocated. The first census of the new AR setting occurred in April 2009 eight months after relocation and the second was in December 2009. Fish recruitment season in this area typically lasts from July until December (Ben-Tzvi et al., 2008). Fish were censused on the ARs using a belt transect of 11 m × 2, which is similar to the dimensions of the ARs’ top view (adapted from Brock, 1954). Four identical

transects spaced 5 m apart were examined in the surrounding areas (natural reef and sea grass) 15 -30 m from the AR’s location (Fig. 2.1). Transects were performed using SCUBA by swimming at a speed of ca. 0.5 m sec-1 and 1 m above substrate while manually recoding the fish occurring within the belt transect. Each transect was examined twice. Transient fish and fishes swimming in the water column were counted during the first pass, while the second pass was used to tally sedentary and cryptic fish. Individuals that appeared to have moved across the transect were counted only once. Only fish larger than 2 cm were counted (Roberts and Ormond, 1987).

2.2.3 Data analysis

Fish communities were described using species richness, abundance, and Fisher’s α diversity index. Since one AR had an extreme outlier pertaining to a large school of Parapriacanthus ransonneti, it was removed from the comparison. To control for variation in abundance, species richness was further examined by a rarefaction analysis using EstimateS 8.2 (Colwell, 2006). Additionally, data was examined to evaluate differences between nocturnal (aggregated nocturnal fish that were hiding in the AR structures during the day) and diurnal species composition (classified according to www.fishbase.org), juvenile presence, and trophic affiliation (trophic levels were determined according to www.fishbase.org). Similarity between communities was examined using non-metric multidimensional scaling (nMDS) with Bray- Curtis similarity (Clarke, 1993) and was accompanied with a cluster analysis as recommended by Clarke, (1993). Data was fourth-root transformed to accommodate for the existence of rare or large schools of fish belonging to the same species. Analysis of similarity (ANOSIM) was conducted to assess the similarity in species composition between habitats (Clarke, 1993). Samples that did not contain any fish were omitted. A similarity percentage (SIMPER) analysis was performed on non-transformed data to identify the dominant species in each habitat (Clarke, 1993). Paired t-tests were used to compare fish diversity indices between the old and new AR locations and between the results of rarefaction curves comparing richness for abundances of 403 and 337 individuals in the old and new locations, respectively. When percentages were examined, an arcsine transformation was used (Sokal and Rohlf, 1995). Statistical analyses were conducted mainly in SPSS™ v 13.0, and Primer-E™ v 5.0 was used to perform a multivariate analysis of species composition.

2.3 Results

2.3.1 Comparison of ARs in old and new locations

2.3.1.1 Fish abundance and diversity

Fish abundance on the ARs was greater and amounted more than twice in the new than the old location (Table 2.1; Paired t-test, t2 = −4.69, p < 0.043). Similarly species richness was significantly higher in the new site (Table 2.1; Paired t-test, t2 = −5.12, p < 0.035), and a comparison of biodiversity using Fisher’s α index revealed that fish diversity was higher but marginally insignificant in the new location (Table 2.1; Paired t-test, t2 = −3.18, p = 0.086). A rarefaction comparison of the ARs in the two locations demonstrated slightly higher species richness at each abundance level in the new location, but the two rarefactions appear to conform to a single curve (Fig. 2.2).

2.3.1.2 Nocturnal- diurnal fish distribution

In their old location, the ARs supported similar numbers of diurnal and nocturnal fish (Fig. 2.3a; Paired t-test, t = −1.29, p > 0.25). In the new AR location, diurnal fish abundance was significantly higher than that of nocturnal fish (Paired t-test, t = 4.94, p < 0.02). The overall abundance of diurnal fish significantly increased in the new AR locations (Paired t-test, t = −3.5, p < 0.04). In both locations, diurnal fish exhibited significantly higher species richness than nocturnal species (Fig. 2.3b; Paired t-test, t = 3.51, p < 0.04 and t = 15, p < 0.001 for old and new locations, respectively). Similar pattern was observed for biodiversity (Fig. 2.3c; Paired t-test, t = 4.12, p < 0.03 and t = 5.52, p < 0.02).

Table 2.1 Fish abundance, richness, and Fisher’s α diversity index, for the old and new locations of the amphibian ARs (n = 3).

Old New Sig. Abundance 78.3 (27.1) 366.6 (102.6) **

Richness 15.6 (4.7) 33.6 (4.0) **

Fisher’s α 5.99 (1.8) 9.12 (1.0) N.S. ** represent paired t-test differences, p < 0.03. Standard deviation is presented in parenthesis

70 Old 60 New

Observed species Observed 20

0 0 200 400 600 800 1000 1200 1400 1600 Individuals

Figure 2.2 Rarefaction comparison of fish diversity on the amphibian ARs in the old and new locations. Confidence intervals reveal that the two curves are not significantly different (paired t-test, p < 0.05).

2.3.1.3 Feeding guilds

Fish trophic positions, based on feeding guild representation, were similar in the old and new AR sites (Appendix Table A.1). Planktivores and zoobenthivores were the most abundant (96.4%, n=613 and 75.8%, n=1356 of total fish sampled, in the old and new ARs respectively) and had the highest species richness (78.9%, n=38 and 75.8%, n=58 of total fish sampled, in the old and new ARs respectively). All other feeding guilds combined amounted to less than 7% of the fish counted on the ARs. Absent from the ARs at their original location, herbivorous fish accounted for only 6% of all fish (n=1356 fish) observed on the relocated ARs.

2.3.2 Comparison of the ARs to the surrounding environments in the old and new locations

2.3.2.1 Sandy areas do not support many fish

In both their old and new locations, the ARs were situated on flat, sandy areas of the sea floor surrounded further away either by a natural coral reef (old) or by sea grass beds (Fig. 2.1).

At both sites, the sandy areas (before AR deployment in the new location and in the old location 10 months after AR relocation) were extremely poor in fish. The sandy section of the original AR location contained a total of four species comprising seven individuals, while in that of the new location I did not record any fish at all prior to AR deployment. The most prevalent species (four out of the seven individuals) in the original area was the burrowing Goby Amblyeleotris steinitzi . All other counted individuals were situated on small, leftover human debris resting on the sea floor.

Figure 2.3 Fish abundance (a) species richness (b) and (c) Fisher’s α biodiversity index of nocturnal and diurnal fish assemblages (±SD) of the amphibian ARs in the old and new locations. Letters represent significance levels (p<0.05).

2.3.2.2 Comparing ARs to natural reefs and Sea grass beds

The similarity in fish assemblages between the ARs and their surrounding environments was examined in both locations (Fig. 2.4 a,b). The most prevalent species on the AR in its original location was the nocturnal fish Sargocentron diadema , while in its new location a diurnal species, Scolopsis ghanam , dominated (Table 1.1). S. diadema constituted 40.4% of the species composition in the old location where each of the other species contributed up to 15%. Only Pomcentrus trichrorus was found in the top five common species in both ARs locations (Table 2.2).

Table 2.2 Contribution to species composition of the five most prominent species on fish of the amphibian ARs in their old and new locations using SIMPER analysis. D/N= Day or Night.

Species Avg. Avg. % D/N Feeding guild abunda similarity contribution nce Old location Average similarity 21.53 Sargocentron diadema 28.75 8.7 40.39 N Zoobenthivore Pomacentrus trichorus 4.75 3.14 14.57 D Planktivore

Cheilodipterus macrodon 5.75 1.95 9.06 N Planktivore Dascyllus marginatus 3 1.23 5.73 D Planktivore Chaetodon paucifasciatus 1.75 1.12 5.19 D Corallivore

New location Average similarity 61.58

Scolopsis ghanam 118.75 31.9 51.85 D Zoobenthivore Neopomacentrus miryae 36.25 5.28 8.57 D Planktivore Parachelinus octotaenia 15.75 2.98 4.84 D Planktivore

Pomacentrus trichorus 11.75 2.87 4.66 D Planktivore Siganus rivulatus 12.5 2.73 4.44 D Herbivore Cluster analysis revealed that sea grass beds formed a distinct out-group (Fig. 2.4a, Table 2.3, ANOSIM, r > 0.974, p < 0.02) separated from all other habitats as a result of their small yet distinct fish community comprising five species and 0.08 fish m-2. The natural reef differed from the artificial structures that were close to each other regardless of time in the new location or structural design (Fig. 2.4; ANOSIM, r>0.587, p<0.007). Multivariate analyses revealed that among the artificial structures, the relocated ARs and the OTARs had the most closely related fish assemblages (Fig. 2.4b, ANOSIM, r = 0.981, p < 0.03) while the old and new AR locations were farther apart (ANOSIM, r = 0.583, p < 0.03).

0 a 20

100

b NR

OAR

NAR

OTAR

Figure 2.4 Multivariate analysis of fish assemblages (made with unstandardized and fourth root transformed data) showing similarity differences of fish community composition in the different ARs and NR environments (a) Cluster analysis and (b) non Metric Multi-dimensional Scaling (nMDS). Dotted lines indicate 15% similarity, and solid lines represent 35-38% similarity. Sand habitats were omitted since no fish were found during transects on the sand in the new location. NR=Natural Reef, OAR= Old Artificial Reef location, NAR= New Artificial reef location, OTAR= Other Artificial Reefs (see section 2.2.1), SG= Sea Grass. Stress = 0.06.

Table 2.3 An ANOSIM pair-wise comparison (r values) between the different natural and artificial habitats. All comparisons were significant (p < 0.03). Old AR New AR OARs Natural Sea location location Reef Grass Old AR location

New AR location 0.583

OARs 0.741 0.981 Old Reef 0.587 0.809 0.911

Sea Grass 1 1 1 0.974

2.4 Discussion

AR fish composition was altered in the wake of the relocation. Fish abundance, species richness and biodiversity were higher on the ARs in their new location (Table 2.1). Rarefaction analysis (Fig. 2.2) attributed this difference to the higher abundance of fish species in the new location, as the rarefaction curves indicate that the fish richness between the two AR locations were similar. Indeed, in their old location, the ARs supported lower fish biodiversity, possibly as a consequence of their proximity to predators, known to travel between prey patches, in the adjacent NR (Bohnsack, 1989; Belmaker et al., 2005; Overholtzer-McLeod, 2006). Since AR complexity and structure, including the added complexity of corals and other fouling organisms, were identical in the two locations, differences in the fish species richness, and biodiversity must be attributed to the effects of the surrounding area. My results suggest that within a given local- scale environment, AR structure has greater influence on the fish community composition than the surrounding environment and biodiversity differences could be attributed to local ecological processes such as predation. ARs located in sandy/sea grass area may be referred as islands with available niches in an area deprived of such attributes. Isolation of these “islands” is most likely dependent on the distance from the source of the main reef to the isolated AR or the size of the AR, as stated by the Island biogeography theory (McArthur and Wilson, 1967) and may be controlling the availability of fish. Since the ARs in the old location were located in close proximity to the NR it was predicted that the abundance and richness of fish would be higher. However, results indicate of an opposite response. I found higher species richness and abundance in the more

distant ARs. This can be explained by two main possibilities. (1) ARs in the new location received immigrant fishes from a source other than a NR, creating a point of attraction within a non-habitat matrix comprised of sandy bottom and sea-grass bed; or (2) Isolation from NR source in the new location reduced predation pressure, as a result of the relatively low arrival of predatory fish (following Belmaker et al., 2005), and (3) the lower structural complexity of the ARs compared to the NR retained less amount of fish in the old AR’s location. The fish community in the new ARs location was different from the old AR fish community thus providing some evidence that the natural continuous reef is not the only source of fish available. Other sources for fish replenishment to the new location can derive from old, small artificial objects left after the removal of fish cages two years earlier (that were known to support a great abundance and diversity of fish not local to the sand/ sea grass beds; S.Shafir personal communication) and smaller artificial objects like pipes and old fishing nets (O.Polak personal observation). Strelcheck et al, (2005) also found that ARs with other ARs in their proximity influence fish abundance and biomass. Predation is a major factor structuring fish communities (Hixon and Beets, 1993; Sandin and Pacala, 2005). It is possible that proximity to natural reef may increase predation pressure resulting in reduced richness and abundance (Belmaker et al., 2005). Predators may be mobile, unlike site-attached species, thus fish inhabiting ARs close to the NR are exposed to higher predation. Additional explanation can be increased structural complexity. The structure of the AR is less complex than the NR, thus fish on the AR may have higher mortality rates due to reduced shelter and increased predation, even at the more isolated locations that attract less predators. Richness and abundance on the old location was lower than in the isolated new location. It was additionally found that predation in the old location was similar to the predators species in the new location (only two predator species on the old natural reef, but eight predator species in both old and new ARs, Table 2.1) therefore my results support the former hypothesis. Thus ARs may serve as high quality fish aggregating devices. There is abundant evidence that ARs support high fish biodiversity (Bohnsack, 1989; Spanier et al., 1990; Rooker et al., 1997; Arena et al., 2007). In the old location, AR proximity to the NR is assumed to have affected their species compositions. NRs are structurally complex and provide fish a variety of niches (Luckhurst and Luckhurst, 1978; Brokovich et al., 2006; Walker et al., 2009). It is therefore assumed that in their original location, surrounded by a complex coral reef, the ARs did

not attract as many fish (but see Arena et al., 2007 and Hunter and Sayer, 2009). Once relocated to a site surrounded by sand and sea grass, the ARs could provide new and rare niches for a large variety of local fish (Spanier et al., 1990). Indeed, Walsh (1985) reported that fish seeking shelter may travel large distances to find suitable habitats. Processes affecting changes in species composition may be found by separately examining the contribution of different fish functional groups. While nocturnal fish abundance and richness on the ARs were similar in the old and new locations, the diurnal fish community after AR relocation increased both in abundance and richness. Thus the ARs served as an aggregating attractor to the latter group. Alternatively, it is also possible that NR proximity to the original AR location allowed diurnal fish to easily migrate between the AR and the NR in response to, for example, increased density on the AR. For the nocturnal fish community, its lack of change between the old and new AR locations could be attributed to habitat specificity rather than to habitat availability. Fish habitat selection is species-dependent (Chabanet et al., 1997). During the day, nocturnal fish usually need shaded, partially closed shelters (Kojima, 1957; Greenfield and Johnson, 1990; Spanier et al., 1990; Cocheret de la Morinèire et al., 2004), which are less common on NRs but prevalent on ARs. Finally, with the exception of lack of herbivorous fish in the reef location (but see Einbinder et al., 2006), no difference between fish dietary guilds in the two AR locations was found. Overall, fish numbers increased following AR relocation. In the sandy/sea grass area of the new AR location, the ARs, whose characteristics differ markedly from those of the local environment, may function as artificial “keystone structures.” Defined as distinct spatial structures that provide resources, shelter, or goods and services for other species, keystone structures are known to increase local biodiversity (Tews et al., 2004). It may be possible that the contrast between the more heterogeneous physical properties (such as complexity and height) of the ARs and those of the more homogeneous immediate surroundings, and the benefits these offer to the fish (such as shelter) are important for maintaining this large fish community. An alternative hypothesis could be related to habitat isolation. It is known that fish community structure can be affected by the extent to which the community's habitat is isolated (Schroeder, 1987; Jordan et al., 2005). Continuous reefs may reach community stability faster than isolated reefs and may have more predictable community structures (Nanamy and Nishihara, 2003; Belmaker et al., 2009; Mellin et al., 2010; Belmaker et al., 2011). However,

isolated patches of reef can support a larger biomass (Belmaker et al., 2005). It was predicted that local scale processes such as competition (Almany, 2004a), predation (Belmaker et al., 2005) and recruitment (Ault and Johnson, 1998) are major determinants that regulate fish assemblages on isolated reefs. The physical attributes and location of the reef may both determine the characteristics of fish community composition on an AR. The study showed that AR structure has a stronger influence on fish community structure than does its natural surroundings as all ARs were grouped together in the multivariate analysis and were distinct from surroundings including NRs (Fig. 2.4). As the ARs at both sites already functioned as well developed reefs with corals and attached fouling, it further emphasizes that when the physical structure of an AR is different from its environment, the communities that develop on it will ultimately differ from those that develop on the NR. Similar results showing that ARs and NRs are distinct from each other were observed elsewhere in fish (Arena et al., 2007; Burt et al., 2009) and in corals (Perkol-Finkel et al., 2006), and it is presumed that these differences are maintained over long periods of time (Burt et al., 2009). My findings showed that fish community composition was different on ARs in different locations and since AR structure remained similar, only local differences altered fish community composition, and these changes were attributed to ecological processes such as predation. It is therefore concluded that fish community composition is determined primarily by the AR structure than the local environment. Further investigation positioning both identical and different AR structure at various environments will better explain the connection between structure and environment on fish community composition.

3. Habitat-dependent movements of a coral-dwelling fish

3.1 Introduction

Coral reefs are composed of mosaics of habitats that differ in complexity. For example, most coral reefs contain both areas of dense coral cover and patches with relatively low cover, such as sand and gravel. Several studies have shown that the spatial organization of habitat patches influences reef fish populations and communities (Shulman, 1985; Sweatman and Robertson, 1994; Ault and Johnson, 1998; Nanami and Nishihira, 2003; Overholtzer-McLeod, 2006; Belmaker et al., 2011). In many cases patch reefs contain greater fish richness and diversity than contiguous reefs (Shulman, 1985; Nanami and Nishihira, 2003; Belmaker et al., 2011). This is often interpreted as affected by predation, which is a major structuring force, especially affecting small fish (Hixon and Beets, 1993; Hixon, 1998; Nanami and Nishihira, 2001; Sandin and Pacala, 2005; Almany and Webster, 2006). Specifically, it is believed that dense coral cover offers refuge from predation for many obligate coral reef fish, while open habitats with low cover are exposed to high predation pressure (Nanami and Nishihira, 2001; Belmaker et al., 2005; Overholtzer-McLeod, 2006; Turgeon et al., 2010). Fish movement appears to be spatially restricted. Site attached fish are usually small and often live within or in close proximity to corals. While reef fish are safer in complex habitats such as corals, most fish species cannot stay confined to a coral indefinitely, as they need to forage, find mates, etc. (Fricke, 1980; Wilson and Godin, 2009). Thus, some site attached fish were recorded to move in scales of up to 3-4 m, like Gobiids (Munday, 2000; Feary, 2007; Wall and Herler, 2008), damselfish and chromis (Fricke, 1980; Belmaker et al., 2007b; Rilov et al., 2007). Several factors may influence travel patterns between patches including, connectivity between patches (Turgeon et al., 2010; Belmaker et al., 2011), presence of territorial fish that hinder movement (Turgeon et al., 2010), exploration behavior (Nomakuchi et al., 2009), diel migration between feeding and sleeping sites (Holland et al., 1993), spawning migrations (Warner, 1995) and ontogenic shift in habitat type (McAfee and Morgan, 1996). Based on the optimal foraging theory, travel probability between patches is expected to decrease with increased distance between patches as well as with increased predation risk (MacArthur and Pianka, 1966). Site attached fish, like damselfish, are usually restricted to one coral and, similarly to central foragers, their movement patters decreases with distance from shelter (Wall

and Herler, 2008). It is therefore assumed that such fish will decrease their movement as the distance from their home coral increases until reaching the boundary of their home range. Dascyllus fish’s home range was recorded at various locations and habitat types and its movement from shelter appears not to exceed 3-4 m (Sale, 1971; Shpigel, 1980; Rilov et al., 2007). Hence, the maximum distance in movement of site-attached fish is determined by the distance to the boundary of their home range. Movement can also be affected by the surrounding environment. The complexity of the settings, (i.e., open sandy spaces vs. continuous reefs) can influence the activity levels of fish (Turgeon et al., 2010). Differences in movement patterns between patches within a habitat are often driven by predation risk, which is mediated by shelter availability (Lima and Dill, 1990; Kotler, 1994). For example, Shuai and Song (2011) have examined the effects of predation risk on the foraging behavior of gerbils in complex and simple microhabitats in two different distances. They found that gerbils foraged more in closer and more complex (sheltered) patches. Several predator-prey fish experiments conducted in simple and complex environments (open space and vegetation filled ponds) have documented behavioural change in both predator and prey (Werner et al, 1983; Savino and Stein, 1989). These studies have documented that predator success is reduced in increased habitat complexity. Similarly, complexity of reef settings (i.e., open sandy spaces vs. continuous reefs) was observed to influence fish distribution and movement behaviour (Ross et al., 2007; Turgeon et al., 2010, Schrandt et al., 2012). It was therefore hypothesized that for site-attached fish (1) in close distances between patches complex environments will demonstrate higher fish movement than simple habitats. (2) movement will decrease with distance between patches until fish will reach the boundary of its home range. The damselfish Dascyllus marginatus is common in the northern part of the Red Sea (Fricke, 1980; Brokovich, 2001). It is symbiotically associated with branching corals, and in the Gulf of Aqaba and it is particularly associated with Stylophora pistillata and Acropora spp (Fricke, 1980; Ben-Moshe, 2007). The species is traditionally regarded as residing within a single coral with a home range of a few meters around it (Fricke, 1980; Shpigel, 1980). However, observations on the reef of Eilat suggest that there may be large variability in the number of fish residing in the same coral over time (Ben-Tzvi and Polak unpublished data). Therefore D. marginatus may have a greater home range than previously hypothesized (sensu; Belmaker et al., 2009).

In this study, I used observations in the natural reef and relocation of detached corals to examine how the movements of D. marginatus are modified by distance between coral patches and change in the surrounding environment from simple to complex. If habitat complexity decreases predation risk due to availability of shelter, I predict that: (1) at close distances movement will be higher in complex vs. simple habitats (2) the movement frequency will be similar at farther distances close to the boundary of the fish’s home range. This study increases the mechanistic understanding of the relationship between habitat structure and fish behavior, that is likely to influence fish growth rates and demography. These questions are particularly important given the fast pace of habitat degradation in coral reefs, resulting in the transformation of complex habitats into simple ones (Pandolfi et al., 2003; Hoegh-Guldberg et al., 2007).

3.2 Methods

3.2.1 Study Site

The study was conducted in the northern Gulf of Aqaba, Red Sea, at the city of Eilat, Israel, in front of the Inter University Institute for Marine Sciences (29º32'85''N, 34º57'47'E, Fig. 3.1). The location can be separated into two areas: the northern area has a narrow fringing reef and a few corals ending 10 m from the shore with an adjacent massive area of gravel devoid of any corals and predators. The southern area, located only 100 m to the south has a flourishing fringing reef and a reef slope that reaches 25 m in depth.

Figure 3.1 The study area (a). The experiment was conducted in areas (1) and (2). In area 1 I conducted the low complexity (b) and the artificial complexity (c). And in area 2 I conducted the high complexity treatment within a small sandy patch inside the reef (d). In (c), corals are covered with a net that prevents the inserted fish to escape and helps them acclimatize to the corals during their first 24 hours after placement. Natural reef measurements were conducted in the area south of experimental station (2).

3.2.2 Natural reef measurements

The relationship between movement and distance between coral was first examined in the natural reef. I snorkeled over the reef and looked for D. marginatus movement between two corals. This was done since many fish do not often move between corals; hence fish movement indicated fish activity and ensured (in many cases) fish movement during the sampling period. I then counted the number of times fish pass between the two corals during 10 minutes. Lastly, I recorded the two corals’ volume (Length×Width×Height), the number of D. marginatus individuals in the two corals, and the distance between corals.

3.2.3 Experimental manipulation of shelter isolation

3.2.3.1 Habitat selection and preparation

To experimentally test the relationship between the distance between corals and habitat complexity on Damselfish movement, I tested fish behavior in three coral environments (Fig. 3.1). I used two Stylophora pistillata corals of similar volume (6070–8900 cm3) attached onto metal frames (Fig. 3.1d). The detached corals were placed at 3–5 m depth in three habitats that differ in shelter availability: (1) gravel (low complexity); (2) gravel with artificially increased shelter availability (artificial complexity) and (3) small sandy patch within the reef (high complexity). Artificial complexity was created by adding the detached corals with cement blocks, pipes, buoys and dead S. pistillata corals (Fig. 3.1b). Outline of the experiment is presented in figure 3.2.

3.2.3.2 Fish collection and acclimatization

Two fishes were collected using clove oil anesthesia (Munday and Wilson, 1997). Total fish lengths were measured and one fish was distinguished from the other with a visible implant elastomer tattoo (Northwest Marine Technology Inc.). To prevent escape, both fish were released into a mesh enclosure that covered the detached corals, and left to acclimatize to the corals. After 24 hours, the mesh net was removed and the fish were left to further acclimatize to the local environment for an additional 4–24 hours before the beginning of the experiment. Preliminary tests showed an acclimatization time of 2 hours, after mesh removal, is sufficient for normal behavior. Both fish were replaced after all three coral environments (see above) were tested or if at least one fish escaped the experimental setup beforehand. Overall I had 41 replications (13 low complexity, 15 high complexity, and 13 artificial complexity) with 20 different fish pairs, and in each replicate I had six different distances of coral separation.

During the experiment, the corals were repositioned in randomly ordered distances of 5, 10, 15, 20, 30, 50, and 100 cm from each other. A video camera was positioned to record fish movement without human presence. A three-minute interval was used to let the fish acclimatize to the new distance followed by 10 minutes in which the number of times the two fish passed

between corals was measured. A pass was defined as a rapid direct movement between the two corals and does not include any foraging behavior. Since fish movement might be influenced by current (A. Genin personal communication), additional measurements were made at 12 different dates and at various current conditions. Four pairs of fish were used in the artificial complexity habitat and with a constant distance of 10 cm between corals. A current meter (S4 ADW, InterOcean Systems Inc.) was placed 5 m from the experimental setup and the fish were filmed as mentioned. Sixty-five replicates of 10-minute intervals were analyzed from the videos. No correlation was found between fish passes and ambient current (r2=0.013, Spearman coefficient=-0.163, p=0.193). Size of fish could be a factor influencing fish behavior. Therefore, correlation between fish size and movement was conducted on 22 individuals from the low complexity habitat. The coefficient of the slope of movement as function of distance was extracted from each individual and later correlated the coefficients by the fish size. No correlation was found between the two factors (r2=0.016, Spearman coefficient=-0.146, p=0.518). To estimate vigilance, I measured the time the fish spent inside the corals from analysis of the video recorded. I measured the time in high and low complexity habitats only, and using one pair of fish to reduce variance in fish personality.

3.2.4 Home range assessment

To assess the use of corals within natural settings and to test the hypothesis that D. marginatus is a central place forager using a single home coral, I conducted a monitoring experiment in the natural reef. This experiment was set in the natural reef location in which four relatively close (<1.5 m) coral concentrations that are isolated (>2 m) from other potential corals were identified. Each concentration contained 4–8 corals. Volume (L×W×H) of all coral was measured as well as distance between them. Within each coral all individual fish were marked as mentioned above. After recovery, fish were immediately returned to the coral they were removed from. Every 1–5 days (between 9 August and 4 September 2011) the corals in each concentration were revisited in the morning (08:00–09:00), midday (12:00–13:00), afternoon (16:00–17:00), and night (19:00–20:00), and the location of each individual fish was recorded. The surrounding area (<2 m beyond the coral concentration’s borders) was monitored for marked fish escaped

Figure 3.2 An outline of the experimental design and the measurement of current vs. fish movement.

from coral concentrations. Night monitoring was conducted with both LED and UV torch (Northwest Marine Technology Inc.) that enables better identification of the fluorescent marks.

3.2.5 Data analysis

The number of passes between two corals was analyzed using a quasi-Poisson model, which allows for over dispersion. The predictors were distance between coral (log transformed for the manipulated experiments), and habitat type was the grouping variable. This analysis was performed using R version 2.15.1 (The R foundation for Statistical Computing, http://www.r- project.org). Frequency of passes was measured to calculate the number of times a fish passed/did not pass between the corals within the designated 10-minute interval. A binary logistic regression with distance and habitats as the covariate and pass/no pass as the dependent variable was used to analyze frequency of passes. To calculate the probability to move between corals the following probability model was used: P (1.1) ln x 1 P

1 (1.2) P  x 1 e where P is the probability that one of the fish passed between corals and x is the distance between the corals (cm). For deviations see Rosenfeld et al. (2008). I found two outliers in the natural reef coral concentrations that were greater than three times the average and were therefore discarded. For the home range analysis, only fish that remained in the coral concentration for the whole period of examination (36 days) were used in the analysis. Fish that disappeared from the tagged corals but later returned were considered as changing location but their travel distance was disregarded. Fish movement within the designated coral concentrations was calculated by summing the number of times a marked fish changed location between two consecutive sampling visits and calculating the percent frequency of its change (n=31 samplings). All statistical data were analyzed with SPSS™ v 13.0 and the formulation of the logistical regression model was conducted in Kaleidoscope.

3.3 Results

3.3.1 Natural reef measurements

Movement of fish between corals in the continuous natural reef revealed a negative logarithmic pattern (Fig. 3.3). Fish movement between coral decreased substantially with increased distance between coral pairs (Quasi-Poisson regression model, β=-0.2416, t=-2.379, p=0.0214). I found that 50% (n=299) of the passes occurred within 17 cm, and 75% occurred up to 30 cm distance. The number of occurrences documented after this distance was lower as the occurrence of encountering fish travelling distances greater than 30 cm in the natural reef is uncommon.

of passes Number 5

0 0 20 40 60 80 Distance between corals (cm)

Figure 3.3 Distribution of number of passes as a function of distance between two random corals in the natural reef. Logarithmic regression of all data is presented (due to over dispersion Quasi-Poisson regression was conducted for statistical analysis, β=-0.2416, t=-2.379, p=0.0214).

3.3.2 Experimental manipulation of isolation and complexity

The number of passes between pairs of corals decreased logarithmically as the distance between the corals increased (Fig. 3.4), exhibiting a similar pattern to that observed in the natural reef. Only 19.1% of the overall passes (n=2768) were made to ≥30 cm distance, 8.16% to ≥50 cm distance, and a mere 1.3% for ≥100 cm. Importantly, controlling for distance, I found a significant difference between the low complexity and the high and artificial complexity

treatments (Fig. 3.4; Quasi-Poisson regression, t=-5.689, p<0.001) but not between artificial complexity and high complexity (t=0.663, p=0.508). Additionally, I did not observe a difference in the rate of decrease in passes between the different habitat treatments, this is attributed to the similar number of passes exhibited at farther distances.

16 High 14 Artificial Low 12 10

4 of Number passes 2

-2 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 2.2 Log distance between corals (cm)

Figure 3.4 Average number of passes (±SE) between experimental corals at different distances at high (●), artificial (○), and low (▼) complexities. Passes between high and artificial complexities were similar to each other (t=-5.689, p<0.001) and different from low complexity (t=0.663, p=0.508).

Since many fish did not cross the distance between the corals during the 10-minute interval of the experimental runs, I also reduced the data to a binomial representation of pass/no pass. The probability to move between corals in all habitats was similar and was around 50 cm (Fig. 3.5). The fish in the low complexity habitat exhibited fewer passes than the two other habitats (Logistic binary regression, p<0.001 for low complexity and p=0.198 between the latter habitats). Only 14.8% of the fish moved distances of ≥50 cm and there was a mere 4.1% (n=324 passes) of occurrences where fish moved between corals 100 cm apart. Video recording analysis indicated that fish in the high complexity habitat spent more time outside their corals than the same fish in the low complexity habitat (Mann-Whitney U test, u=140064, p<0.001).

Figure 3.5 Probability to move between corals vs. distance between corals in the low, high, and artificial habitats. Dashed lines represent ±SE of mean (solid lines) values. For calculation of deviations see Rosenfeld et al., (2008).

3.3.3 Home range analysis

In natural observations, 20% of the fish (n=40) did not move at all, most fish (42%) had low activity level (1–10% change in location, n=31 sampling times), 22.5% had moderate activity level ranging between 10% and 20% changes for the period of examination, and 15% were highly active (>20% changes in location). No correlation was found between maximum distance travelled by the fish and the variation in location (Appendix Table A.2; Spearman r2=0.397, p=0.16, n=14 fish that travelled known distances), also no correlation was found between variation in fish location and distance to nearest coral (Spearman r2=-0.109, p=0.502, n=40). Additionally, no correlation was found between number of passes and fish size (Appendix Table A.2; Spearman r2=-0.089, p=0.595, n=40).

Table 3.1 Home range attributes of Dascyllus marginatus. Error indicates standard deviations.

No. of Plot No. of Avg. Distance to Distance Avg. Avg. corals marked sighting closest travelled change in percent fish incidents coral (cm) (cm) location change

13 1 16 28.5±2.5 18-300 10-572 3.3±2.63 7.46±6.32 13 2 8 48.13±55.67 10-175 35-275 5.62±2.77 18.58±9.57 14 3 6 33.83±10.67 23-52 111 3.67±3.2 11.82±10.33 6 4 10 53.7±10.14 32-53 50-57 1.4±2.22 4.52±7.17

3.4 Discussion

Predation plays a major role in the structuring of coral fish communities (Hixon and Beets, 1993; Hixon, 1998; Nanami and Nishihira, 2001; Sandin and Pacala, 2005; Almany and Webster, 2006). It is known that risk avoidance impacts fish behavior (Werner et al., 1983; Rilov et al., 2007). In the experimentally detached corals and in the natural reef I found that fish significantly reduce movement between two adjacent corals with increased separation. This is attributed to increased predation risk. Complexity is known to increase fish abundance and richness in natural (Brokovich et al., 2006) and artificial reefs (Almany, 2004b; Gratwicke and Speight, 2005). While this is believed to be caused by decreased predation in complex habitats (Savino and Stein, 1982; Manatunge et al., 2000), this has been rarely demonstrated in natural reef setting. In the detached corals experiment, fish changed their behavior from the denuded gravel habitat to match the behavior found on the reef just by increasing habitat complexity. This suggests that this response is due to visual perception Few fish studies have recorded effects of increased habitat complexity on fish behavior such as mating behavior (Myhre et al., 2012), and most studies investigate the negative effect of complexity on predatory success (Savino and Stein, 1982; Nelson and Bonsdorff, 1990; Manatunge et al., 2000) due to decreased obstruction of movement and visibility. The fish used in the experiments were not practicing mating behavior and, while I did not directly examine predation, I did observe in the recorded videos less predators in the gravel location. Additionally, after artificially increasing complexity in the simple (gravel) location the behavior equaled that

in the complex (reef) location where predator presence was higher. I therefore assume that change in behavior is a result of perceived risk of predation. Reduction in movement rate was found similar at either habitats and was attributed to similar number of passes at distances >50 cm between corals. This appears to be as a result of the innate home range boundary of D. marginatus. This means that at close distances habitat complexity, as proxy for shelter availability, helps modify fish behavior while at farther distances perceived predation risk is independent of habitat availability. The results conform to the proposed hypothesis regarding habitat availability as a major factor eliciting change in fish behavior at continuous vs. patchy reef patches. While I found that D. marginatus is not entirely confined to a single coral and that this species exhibits constant low mobility in the natural reef, D. marginatus appears to have a small home range and is active mainly within 30–50 cm from its “home coral”. Similarly, Rilov et al., (2007) found that the damselfish Stegastes partitus travels a maximal distance of 3 m and spends over 70% of its time within a distance of 50 cm from its nest. D. marginatus usually resides in one single “home” coral and occasionally visits several nearby corals. This use of occasional corals may enable D. marginatus to extend his home range to farther distances from its “home” coral. Rilov et al. (2007) suggested that the male S. partitus resides within an ecological neighborhood (an area in which an organism spends most of his time, sensu Addicott et al., 1987) and is constrained to less than 80 cm. It is important to mention that mobility of fish can be age dependent, as post-recruitment D. marginatus are known to change corals while mature individuals move less (Belmaker et al., 2009). Small reef fish are believed to have limited home ranges, with movements that rarely exceed 10 m (Einbinder at al., 2006). Although movement distances of D. marginatus are usually low, some individuals displayed high movement levels. These individual differences, contributing to the high variance in the data (average passes range between 20.6±15.8 SD at 5 cm distance to 1±2.3 at 100 cm), may be the result of boldness behaviors where some fish are willing to take more risk than others (Harcourt et al., 2009; Wilson and Godin, 2009; Wilson et al., 2010). Other studies demonstrated similar intraspecific variance in other fish traits (Bell and Stamps, 2004). Conversely, individual differences may be attributed to location within the group hierarchy (Fricke, 1980), which was not accounted for in this study.

3.5 Conclusions

Reduction of fish movement as a consequence of reduced habitat and shelter availability (Wall and Herler, 2009) may have negative consequences on fish’s foraging, breeding, nest maintenance and ultimately their fitness (Wilson and Godin, 2009). This may have wide scale implications as coral reefs are declining rapidly, resulting in decreased live coral cover (Pandolfi et al., 2003; Feary et al., 2007; Hoegh-Guldberg et al., 2007). It is already established that there is a strong connection between coral volume and fish density (Schmitt and Holbrook, 2000), but a change in behavior of fish as a consequence of spatial distribution of corals is yet to be documented. My study recorded that D. marginatus reduces movement between corals greater than 50 cm apart and it seldom travels more than 100 cm between corals; additionally, its movement frequency is substantially reduced when coral cover is low. Therefore, reduction in shelter availability may impact fish in ways that are unaccounted for by measurements of live coral cover that do not account for the coral spatial distribution. The reduced state of reefs led to transplantations of corals as a remediation measure. The knowledge that some coral dwelling fish possess different behavior as the distance between corals changes is an important aspect in the design of coral transplantations. Notably, reef restoration and rehabilitation have to take into account not only the wellbeing of corals but also the supply of proper environment that will maximize the biological functions and ultimately the fitness of the coral’s obligate dwelling fish.

4. AR with coral transplants: a conservation tool for fish enhancement? A cases study from Eilat Red Sea.

4.1 Introduction

Coral decline is measured globally and most coral reefs are expected to be at risk of long tern degradation by year 2050 (Frieler et al., 2013). This massive decline in coral reef entails conservation measures. The first measure is the “passive approach” that includes creation of management tool to protect the habitat at risk, such as the allocation of Marine Protected Areas (Fernandes et al., 2005) and associated measure such as increased legislation, enforcement, involvement of stakeholders etc (Epstein et al,, 1999; Rinkevich, 2008). The second is the “active” approach that includes rehabilitation and restoration of degraded sites (Clark and Edwards, 1995; Epstein et al.2001; Rinkevich, 2005) and the creation of alternative sites like artificial reefs (Pickering et al., 1998; Miller, 2002; Abelson, 2006). Artificial reefs (ARs) are known to aggregate fish from the environment and create a novel habitat that may increase local fish biota (Bohnsack, 1989; Rilov and Benayahu, 2000). On the other hand ARs are considered in many cases as an ecological trap where aggregated fish are fished regularly (Bohnsack et al., 1994). ARs are known to either be a productive or attractive habitat and therefore their role as conservation tool is debated (Bohnsack, 1989; Pickering and Whitmarsh, 1997; Brickhill et al., 2005). As most deployed ARs are intentionally sunk for enhancing fish catch (Svane and Petersen, 2001) little is known about the effects of deploying and AR in no-take zones on the fish community composition and dynamics. Therefore, the introduction of new artificial environment in close proximity to local fragile environment, such as coral reefs, demands monitoring the effects of artificial reefs on the local natural reefs (Carr and Hixon, 1997). The success of ARs for fish production is dependent on several factors (see elaboration in chapter 2) and can either be controlled by the AR designer like location of AR, AR structure design and complexity, or controlled by natural occurrences like ecological processes such as colonization (mostly recruitment of juveniles and immigration of adults), and other post- settlement processes like predation and competition . The location of AR deployment plays a

major role in the creation of stable fish community (Walsh, 1985; Herrera et al., 2002; Jordan et al., 2005). ARs isolated from continuous coral reefs are known to increase fish abundance (Shulman, 1985; Ault and Johnson, 1998), presumably as a result of reduced predation and competition (Walsh, 1985; Belmaker et al, 2005). The ARs structure has unique features like vertical surfaces, increased shaded area, and its substrate composition. These attributes corresponds with specific fish communities (Pickering and Whitmarsh, 1997). Habitat complexity can be further enhanced by ARs and thus enhance fish indices (Sherman et al., 2002; Gratwicke and Speight, 2005). Colonization of new habitats, such as novel ARs, is dependent on several factors that may affect it. Seasonality of recruits causes change in the adult fish population (Doherty, 1983). Colonization may be dependent on the distance from a continuous natural reef, where closer ARs are expected to receive greater emigrants and therefore colonize faster than farther more isolated ARs (MacArthur and Wilson, 1967). On the contrary colonization through immigration of adults and recruitment of juveniles to isolated ARs was found to sustain higher fish biomass (Ault and Johnson, 1989; Belmaker et al., 2005) . Other factors that may influence colonization rates include predation pressure and availability of shelter (Hixon and Beets, 1993), AR structure (Walsh, 1985; Herrera et al., 2002; Rilov and Benayahu, 2002) and AR location (Herrera et al., 2002). Stochastic occurrences can also influence fish community over time. Incidents such as strong storms and predation episodes were recorded to change fish community composition (Walsh, 1983; Ault and Johnson, 1989) as well as regular recruitment events (Sale, 2004). These processes are dependent on time. Hence, it was hypothesized that time since deployment of the AR will influences the fish composition of the AR. While it was proposed that ARs can serve as surrogates for natural coral reefs it was demonstrated that AR and natural reefs fish composition remain different over long periods of time >30 years (Arena et al., 2007: Burt et al, 2009; chapter 2). It was therefore hypothesized that time since AR deployment will maintain the difference in fish communities between the AR and the natural reef. Transplantation of coral on ARs was conducted sporadically and research on such conservation actions is limited (Ortiz-Prosper et al., 2001; Ferse, 2008). Transplantation increases habitat complexity and creates new niches for fish, ultimately increasing the overall fish richness, and abundance and possibly change their community composition. Thus, coral transplantation on the AR is hypothesized to influence the richness and abundance of reef fishes

on the AR. In this study I monitored the fish composition on a small AR located in close proximity to two similar sized natural reef outcrops. I aimed to observe the colonization patterns and evaluate the ability of such AR to serve as fish production site without endangering the local fish community on the nearby natural reef outcrops. I predicted that (1) fish diversity indices (i.e. fish richness and abundance) and composition on ARs will be different from natural reefs (2) these measurements will change over time as consequence of stochastic events and (3) the transplantation of corals on ARs will increase diversity indices and fish community composition. ARs with coral transplants is a new conservation tool and its use, especially in light of the fast decline of coral reefs (Pandolfi et al., 2003; Frieler et al., 2013), appears important.

4.2 Methods

4.2.1 The study site

An artificial reef (AR) was submerged in the locality of the Coral Reef Nature Reserve, Eilat, Gulf of Aqaba, in April 2006 (Fig. 4.1; Polak and Shashar, 2012). The AR was immersed at a depth of 7 m on a flat sandy bottom 10 m from the reef slope and in the vicinity of two similar sized coral outcrops, A and B. The site is found approximately 70 m from a continuous fringing reef. The general area is the most visited dive site in Eilat with an estimate of over 120,000 dives yr-1 (accumulated reports from dive clubs 2011, unpublished data).

4.2.2 AR design

The AR was composed of six 2×2×2 m units of concave hollow concrete Ocean Brick™ blocks (Fig. 4.2). Each module weighed 4.2 tons (weight in air) and was pre-drilled to accommodate coral transplants of matching stub sizes. The AR design was planned to allow for passing of currents and waves, for great stability, for increase in void space, and lighted and shaded areas. Additionally, the four-meter height of the final AR design was determined to allow enough vertical relief for fish aggregation. To increase complexity and shelter, a metal “tree” composed of 40×5 cm diameter tubes was inserted in the void space directly on the bottom. All passages were sealed with metal frames to deny SCUBA and snorkeling access, hence creating a non- disturbed refuge for fish.

Figure 4.1 Study site and location of AR deployment. The AR (shaded) was located near the nature reserve and in the vicinity of coral outcrops A and B.

Figure 4.2 Design of the AR. The structure is composed of modular units that can be configured according to desire. The structure encompasses vertical, shaded, and curved micro-habitats (following Polak and Shashar, 2012).

4.2.3 Transplantation

Concurrent with the preparation for AR deployment, 14 hermatypic coral species and two soft coral species (Table 4.1) were reared in a specially designed mid-water floating coral nursery (Shafir et al., 2006). Five months after AR deployment corals were transplanted on the AR. Overall, 230 stony corals primarily of the branching species Stylophora pistillata, Pocillopora veruucosa, and Acropora eurystoma, and the hydrocoral Millepora dichotoma were transplanted. Coral sizes ranged from small fist-size corals to large corals approximately 20 cm in diameter.

Table 4.1 List of corals reared on the designed coral nursery and later transplanted on the artificial reef. Coral class Species Hard corals Acropora spp. Pocillopora verrucosa Stylophora pistillata Favia favus Porites lutea Millepora dichotoma Cyphastrea serralia Cyphastrea chalcidicum Platygyra daedalea Plerogyra sinuosa Goniopora sp. Goniastrea sp. Soft corals Lithophyton arboreum Dendronephthya hemprichii

4.2.4 Data collection

Four months prior to the AR deployment, fish and coral monitoring of the adjacent coral outcrops were conducted. Total number of fish on the coral outcrops (and later also on the AR) were counted for a maximum time of 30 min. The first 2 minutes were dedicated to counting only transient and moving fish within a 2 m radius from the outcrop’s edge and later counts disregarded this group of fish to eliminate recounting. All counts were made at same time of day (10:00-13:00) to avoid change in fish diurnal behavior and expect maximum activity level (sensu Rickel and Genin, 2005). Live coral cover was measured in each coral outcrop with six 0.5×0.5 m quadrates placed at north, south and top at the inner and leeward sides. Corals positioned at

the rim of the quadrate and intruded more than 50% within the quadrate were included in the count. After the initial count of all the coral outcrops in the area, two outcrops (A and B) of similar size and close proximity to the AR were chosen for further monitoring. For comparison between outcrops A, B, and the AR, two counts were made within a week’s time and the date comprising the greatest fish richness and abundance was selected for analysis. In case results differed by more than 10% the outcrop was recounted. The two outcrops and the AR were censused for fish abundance and diversity every three months. For change in the fish community, including colonization, the AR was exclusively counted as mentioned above on days 1, 2, 8, 16, and 31, and approximately every month thereafter (at which only a single count was done).

4.2.5 Data analysis

Similarity between NR and AR was examined with repeated measures non-metric multidimensional scaling (nMDS) with Bray-Curtis similarity (Clarke, 1993). Data was fourth- root transformed to attenuate the large school of fish arriving, especially during the summer season. Grouping was done using percent similarity derived from cluster analysis. A PERMANOVA analysis was conducted with pair wise comparisons to differentiate between the data, and SIMPER analysis was used to compare the dominant species and their relative contribution. Grouping Analysis was conducted using Primer-E V. 6.0. Parametric statistics were conducted using SPSS™ v. 13.0.

4.3 Results

First, fish of the species Pseudoanthias squamipinnis arrived during the assembly of the AR and before the official monitoring period began. Overall fish colonization increased until reaching an asymptote at approximately 17 month (Fig. 4.3). Fish abundance and richness in both overall fish count (Fig. 4.3a) and in juveniles alone (Fig. 4.3b) had increased in a stepwise manner. The first step (steps are demonstrated in Fig. 4.3a) leveled off at approximately 4 month; another increase was detected after 6 month (154 days) that leveled by two weeks later, and a third increase was noticed 17 months (509 days) after deployment. Juveniles had a similar pattern but only two phases appear: after coral transplantation on month 6 (154) and 17 months (day 509) after AR deployment. The dates of the first and third phases showed an increase in the

overall fish and in juvenile counts corresponding to each other, and suggesting a seasonal effect. Detailed observation of the juvenile (Fig. 4.3b) shows that a peak in the abundance of juveniles is followed by an increase in juvenile richness. Additionally, coral transplantation 6 months after AR deployment (day 154) was also followed by an increase in juvenile fish richness. Although data shown is of all adults and juveniles combined, it is evident (Fig. 4.3b) that the seasonal patterns are related to juvenile recruitment. Physical attributes of the three outcrops were found to be similar (Table 4.2). Natural outcrop A was most complex (rugose) and also had more small holes and crevices than the other outcrops. The AR was the tallest structure and had more shaded areas and void space while outcrop B was the least unique but with similar rugosity to the AR. Coral live cover was highest in outcrop A followed by B and the AR, whose coral cover was mostly a result of coral transplantation. Table 4.2 Physical and biological properties of the three examined outcrops: A, B, and the AR. Attribute A B AR Rugosity 1.63±0.26 1.29±0.06 1.19±0.18 Projected area (L×W) (m-2) 23.6 21.8 18 Height (m) 3 2.2 4 Volume (L×W×H) (m-3) 71 48 72 No. of small holes +++ + ++ No. of crevices +++ + ++ Void space ++ + +++ Shaded area ++ + +++

Live coral cover (%) 53.9±4.3 36.9±4.3 5.1±0.5

+low, ++ medium, +++ high.

Recordings of the common species on the AR revealed variable patterns coinciding with the fish’s habitat preference (Fig. 4.4). (1) Structure-associated fish conformed to regular seasonality with no relationship to coral transplantation: P. squamipinnis exhibited moderate seasonality and its peaks coincided only in the 1st and 3rd recruitment periods, while Neopomacentrus miryae showed a distinct recruitment pattern early in the summertime (March– June); (2) coral dwelling fish reacted to both seasonality and transplantation but not to AR structure;. The coral dwelling species Dascyllus marginatus and D. trimaculatus displayed similar seasonality at the 1st and 3rd recruitment seasons, and both synchronized their recruitment

to the beginning of the winter (November–January). Additionally, both species appear to have a high peak at the 1st season coinciding with the addition of coral transplants. (3) Nocturnal fish responded intensively to seasonality and to the presence of the AR, but did not show a relationship to coral transplantation; Parapriacanthus ransonneti and Apogon cyanosoma are both nocturnal fish and both species comprise the most abundant fish during recruitment periods. P. ransonneti presents a yearly recruitment peak, while A. cyanosoma showed a high peak only in seasons 2 and 3 (4) herbivores responded to structure presence that may provide it substrate for food but not to seasonality or transplantation; The herbivore Acanthurus nigrofuscus showed a notable increase in abundance after coral transplantation and remained constant thereafter. (5) lastly, the corallivorous fish Chaetodon paucifasciatus showed no trend in abundance related to coral transplantation or time.

Figure 4.3 Fish count of the abundance of fish (blank circles) and species richness (black circles) on the reef throughout the monitoring period in all the fish (a) and in only juveniles (b). Small graph (c) presents the data of the first 261 days after deployment and shows increase in both richness and abundance. Arrows denote time of transplantation event. Gray bars represent steps in development (see text).

Figure 4.4 Representation of common fish abundance colonizing the AR. Fish that are aggregative and are located at the top of structures (P. squamipinnis and N. miryae), coral obligates (D. marginatus and D. trimaculatus), nocturnal (A. cyanosoma and P. ransonneti), the herbivores (A. nigrofuscus and the corallivore C. paucifasciatus). Note the difference in scale in the abundance axis. Arrow marks time of first transplantation event.

Figure 4.5 Comparison of (a) species richness and (b) fish abundance after AR deployment between the AR (square) and two nearby coral outcrops: A (triangle), which has the highest coral diversity, and B (circle), which has moderate coral diversity.

A comparison between the AR and the natural coral outcrops A and B (Fig. 4.5) revealed that the AR species richness already resembled the other coral outcrops 10 months after deployment and increased until leveling off after 25 month, though not surpassing the natural reefs (Fig. 4.5a). The abundance of fish was similar 15 months after transplantation, and from 23 months and after, the abundance of fish on the AR was greater than on the two natural reefs. Additionally, fish on the AR displayed a multi-peaked change in abundance unlike the two natural reef outcrops that remained relatively constant. Species composition was different between the natural reef outcrops A and B and the AR

(Fig. 4.6); PermANOVA, Pseudo F(1,.60)=15.212, p(permutated)=0.001). The two natural outcrops were also distinct from one another (PermANOVA, pairwise comparison, t=3.58,

p(permutated)=0.001) but showed relative similarity (57%, using Bray-Curtis similarity index) in their community composition. The AR was set apart and had a similarity of 44%. Two distinct groups in the AR fish assemblage were identified during development (PermANOVA, Pseudo

F(1,.60)=15.212, p(permutated)=0.001). The first group was found to be the assemblage of <17 months since deployment, which had a similarity of 53%; the second group, >17 months, was more similar to itself (61%). SIMPER analysis of dominant species revealed that the species similarity was higher in the NR than in the AR. P. squamipinnis was the most dominant species in both the AR and NR (Appendix Table A.3); in both locations the first five dominant species shared three identical species.

Figure 4.6 An nMDS repeated measures analysis of fish composition on the AR (empty icons) and adjacent coral outcrops A (black triangle) and B (grey triangle) over time. Each sample represents a census of the structure. Boundaries indicate the percent of similarity as derived from cluster analysis.

A long-term monitoring nMDS analysis of the fish species composition (Fig. 4.1) revealed a difference between the three groups (PermANOVA, Pseudo F(2,.53)=19.85, p(permutated)=0.001). A SIMPER analysis showed that the similarity increases over time (Appendix Table A.3). In addition, other than P. squamipinnis, the dominant fish composition changes in each of the groups as the time of the AR in the water progresses. Furthermore, PCA1 explained 32.1% of the data.

Figure 4.7 An nMDS repeated measures analysis of fish composition on the AR over time. Time of sampling is grouped to before transplantation (open triangle), <17 months (full triangle), and >17 months (square).

Examination of the effect of transplantation on fish species composition revealed an effect both in the similarity between the two time frames (49.1% similarity before transplantation and 30.76% after) and in species composition (PermANOVA, pairwise comparison, t=4.69, p(permutated)=0.001).

4.4 Discussion

Colonization to the AR followed a three step increase in fish diversity indices (richness and abundance). The first step was associated with immigration of adult fish to the AR, the second one to increased complexity and new habitat niches and the third to a large juvenile recruitment event. Colonization was rapid and stabilized after 17 months thus resembling other studies from the Red Sea (Ben-Tuvia et al., 1983; Golani and Diamant, 1999).

The initial recruitment of fish is attributed to immigration of adults and some juveniles from the local environment. It was observed, even before the official counting of the AR began, that Pseudoanthias squamipinnis adults were already present on the AR. Aggregation of adults from the local environments in matter of hours after deployment was recorded elsewhere (Cummings, 1994; Golani and Diamant, 1999; Leitão et al., 2008). It was documented that fish can travel considerable distances to reach new habitats (Alevizon and Gorham, 1989; Folpp et al., 2011), therefore crossing ca. 15 m of distance between the AR and the nearby natural coral outcrops appears feasible. Interestingly, during the initial recruitment phase, the diversity indices recorded on the nearby natural outcrops remained similar to each other but different from the AR. This indicates that either the fish immigrated to the AR arrived from farther distances or their emigration rate was low enough not to be noticed in the natural outcrops. Movement of reef fish at distances >500 m was recorded to be common (Chapman and Kramer, 2000) and support the attraction approach of ARs of fish, at least in the initial stages of deployment. Recruitment of fish at this stage is associated with properties of AR structure. It is assumed that the height of the AR, the augmentation of shaded areas and presence of flat-vertical surfaces contributed to the quick colonization of fish to equal the diversity indices found on the nearby natural reefs. These same AR properties are also attributed to recruitment of rare species (Pickering and Whitmarsh, 1998) causing the difference in the fish community assemblage that was formed on the AR from as early as the initial phases of colonization. The second recruitment phase was associated with transplantation of corals. Addition of corals to the AR caused immediate increase in structural complexity. This increase in complexity is associated with increase of fish abundance (Gratwick and Speight, 2005). While dead corals are often documented not to affect the increase in fish diversity indices (Holbrook et al., 2008) attachment of live corals, and especially coral species that can harbor coral-dwelling fish, was formerly predicted to increase fish diversity indices (Schmitt and Holbrook, 2000; Abelson, 2006). Undoubtedly, coral transplantation caused an increase in fish richness and abundance and caused change in fish community composition as well. Similarly to rehabilitation efforts conducted elsewhere (Cabaitan et al., 2008; Yap, 2009), transplantation of corals appear to be an effective tool for fish conservation in no-take zones. The third step in colonization was caused by a large recruitment event of juveniles. Recruitment of this last stage resulted in an increase in the diversity indices and in change in

fish species composition. Stochastic processes such as recruitment are known to effect fish species assemblage on natural and artificial reefs (Chesson, 1998; Syms and Jones, 2001; Sale, 2004). The magnitude of the recruitment pulse is considered stochastic and, at least at the immediate time scale, affects the species composition (following the recruitment-limitation hypothesis; Doherty, 1983; Doherty and Fowler, 1994). Indeed, I also observed varied degree of recruitment levels and only the larger recruitment event was able to shift the AR fish community to a higher level of increased biodiversity and increased species composition similarity. It is believed that the large abundance of several species in these recruitment event contribute greatly to the overall species composition similarity. Species composition differed between AR and NRs. Diversity indices on the AR quickly resembled those of the NR outcrops and in some cases exceeded them (Fig. 4.4b). Greater species diversity and abundance coincided with greater ecological and physical attributes of the NR outcrop A. The AR, on the other hand, had lower biological attributes (such as coral cover) but consisted of greater physical attributes like shade and void space (Table 4.2). It is likely that these last attributes are lacking in the local natural reef and therefore are responsible for the increase in biological indices. My results conform to other studies that found differences between ARs and NRs (Arena et al., 2007; Burt et al., 2009). Interestingly, even very short distances (<15 m) between natural and artificial habitats did not mask differences in fish composition entailing that either: (1) migration of (adult) fish is restricted to certain species only (see chapter 3) or (2) recruitment of juveniles is different due to the AR’s structural and biological attributes (difference in coral composition; see Perkol-Finkel et al., 2006). Different fish species responded differently to the presence of the AR (Fig. 4.3) and its unique attributes like vertical planes, void space, and shaded areas to settle. Some site-attached fish require the presence of live branching corals (Bell and Galzin, 1984; Croker at al., 2012) and herbivorous fish use the algae growth on the AR’s substrate as feeding ground (Williams and Polunin, 2001). The nocturnal species Apogon cyanosoma and Parapriacanthus ransonneti formed the vast majority of individuals during the peak recruitment seasons yet were much less abundant in nearby coral outcrops A and B. Similarly, high abundance of the nocturnal species Sargocentron rubrum was recorded on ARs in the Mediterranean Sea (Spanier et al., 1990). This may indicate that there is limited niche availability to these species in the natural reef. The corallivorous species Chaetodon. paucifasciatus, which is considered a reef health indicator

species (Crosby and Reese, 2005), surprisingly did not respond to any of the changes on the AR. It is assumed that its lack of response to coral transplantation is related to the relatively few corals transplanted (Table 4.1). Fish community composition on the AR became similar over time (Fig. 4.7). Here again the three stages of fish colonization can be seen, where each step increases similarity in the fish composition. To my knowledge no such distinct colonization stages have been documented elsewhere, but evidence exists to change in community composition with increase in benthos growth (Leitão et al., 2008). Additionally, increase in similarity of species composition over time has been detected (Bohnsack 1983; Leitão et al., 2008). The use of ARs with coral transplants is a novel method in marine conservation. Here I show that coral transplantation on ARs is only partially affecting the fish species community on ARs by increasing species richness and abundance. Despite that, no distinct change was detected at the species composition level. I found, as predicted, that fish composition was different between NR and AR. And lastly, unlike my hypothesis which claimed that stochastic factors will cause uncertainty in fish community composition over time, I found that the fish community composition has become more similar as time progressed. Other considerations such as substrate type, AR structure, AR location, biological and ecological influences such as predation and competition and environment factors are also likely to determine species composition.

5. Can transplantation of corals on ARs in high pressure dive sites succeed? A case study from Eilat, Red Sea.

5.1 Introduction

Coral reefs globally are in state of decline (Wilkinson, 2008). One of the factors attributed for this decline is physical damage to corals inflicted by SCUBA divers. In the last 20 years a dramatic 78.5% increase in diving certifications yr-1 has been recorded worldwide (www.padi.com). Consequently, different mediation approaches have been adopted to slow or perhaps even reverse the damage caused by divers. One of these approaches is the use of ARs to deviate divers from natural reefs to artificial ones (Abelson, 2006; Leeworthy et al., 2006). Divers are selective regarding their dive sites and are known to prefer natural reefs over artificial ones (Johns et al., 2001; Pendleton, 2004; Sutton and Bushnell, 2007). Nonetheless, divers’ use of ARs is substantial. For example, in Eilat (Red Sea) divers were found to visit a single wreck dive site, located at close proximity to natural reefs, at approximately 16,000 dives (7%, n= 250,000) of their dives conducted yearly in the region (Wilhelmsson et al, 1998). It was also found that divers are choosy about their AR sites and prefer large, themed structures (Shani et al., 2012). Divers are also known to appreciate biological factors such as fish and corals (Wielgus et al., 2003; Leujak and Ormond, 2007). Therefore it was suggested that integration of coral transplants on ARs may increase AR attractiveness and eventually draw divers to these more attractive sites (Fitzhardinge and Bailey-Brock, 1989; Abelson, 2006).

5.1.1 Coral transplantation and recruitment

Transplantation of corals is employed in restoration efforts, and is primarily used to accelerate natural recovery (Rinkevich, 2005; Abelson, 2006) and to bypass coral’s early slow growth rates (Harriot and Fisk, 1988). Additionally, coral transplants were suggested to enhance the aesthetic value of an area for divers (Shinn, 1976; Fitzhardinge and Bailey-Brock, 1989). The main method employed by most researchers and conservationists is a one-step direct transplantation of coral fragments (Rinkevich, 2005) by drawing fragments from donor colonies in the nearby area and transferring them to denuded reefs needing to be recolonized. Another approach pursues a two-step protocol and follows the “coral gardening” concept (Rinkevich,

1995). The first step after the collection of fragments, nubbins (Shafir et al., 2001), or sexual recruits, is growing them in either in situ or ex situ dedicated nurseries, where in the second step they are transplanted to the reef. Shafir et al, (2006) successfully tested a mid-water nursery that proved to be more effective than bottom-anchored coral nurseries (but see Shaish et al., 2008).

Success of coral transplantations is indiscriminate and evidence supports a wide range of coral survivorship rates (5–51%, reviewed in Rinkevich, 2005) and growth rates (Edwards and Clark, 1998; Yap, 2004), yet size-dependent survivorship is confirmed (Edwards and Clark, 1998; Bowden-Kerby, 2001; Soong and Chen, 2003 Transplantation success is also known to depend on the species of coral used (Rinkevich, 1995; Yap, 2000; Epstein et al., 2001), location (Raymundo, 2001), type of substrate (Yap, 2004), and fecundity and sexuality of transplants and donor colonies (Rinkevich, 1995; Epstein et al., 2001; Horoszowski-Fridman et al., 2011). Many coral transplantation efforts are conducted in open seas where wave action is considerable (Edwards and Clark, 1998). Such a location often increases transplant dislodgement and reduce overall transplant survivorship (Clark and Edwards, 1995). Therefore outcome of transplantation events also depends on coral self-attachment to substrate (i.e. growth of coral tissue on the substrate). An experiment conducted in Bolinao, Philippines, found that coral self- attachment depends on the species of coral used, its growth form and its life history (Guest et al., 2011). It was therefore hypothesized that species of the coral transplants will influence its overall survival and the coral’s ability to self-attach to the AR. Coral recruitment rate differs in various locations. Mortality of new recruits is associated with predation (Acanthaster planci, corallivorous and herbivore fishes and sea urchins) and physical abrasion, and is usually reduced in more complex habitats enabling escape from these two factors (Miller et al., 2000). Coral transplantation has, in most cases, been shown to have no effect on coral recruitment (Ferse, 2008, but see Baird and Hughes, 2000).

Transplantation of corals onto ARs is a relatively new restoration approach in conserving coral reefs (Oren and Benayahu, 1997). ARs with coral transplants are usually used either for restoring damaged reefs or for reducing tourism pressure off natural reefs (Edwards and Clark, 1998; Rinkevich, 2005; Leeworthy et al, 2006). The concept of coral transplantation to ARs follows that of natural reefs. Edwards and Clark (1998) have demonstrated that fragments of fast growing coral species are easily dislodged, thus having high mortality rates, while natural

recruitment produces similar growth results in a similar time frame, at least in branching corals. They, therefore, proposed to adopt the transplantation of massive corals but recommend using coral transplantation as a last resort. Consequently, Rinkevich (2005) concluded that use of transplants on ARs has not proven to be successful, but in contrast, Abelson (2006) has recommended the use of this method for conservation purposes.

5.1.2 Divers impact on artificial reefs

The use of SCUBA has been increasingly popular (Davis and Tisdell, 1995) and concurrently the damage divers inflict on coral reefs has increased as well. Divers are known to have major impact on natural coral population (Hawkins and Roberts, 1997; Rouphael and Inglis, 1997; Hawkins et al., 1999; Zakai and Chadwick-Furman, 2002). In Eilat, Zakai and Chadwick-Furman, (2002) documented a positive relationship between coral damage and increased diving activity. They found that branching corals were the most damaged growth form. Consequently, diver management efforts were suggested to reduce diving pressure off natural reefs. One of these measures is the use of ARs (Wilhelmsson et al., 1998; Zakai and Chadwick-Furman, 2002). As mentioned before, self-attachment of corals was suggested to increase coral survivorship (Guest et al., 2011) particularly in areas associated with strong currents and wave action (Clark and Edwards, 1995). Corals transplanted on ARs will therefore be more loosely attached than naturally growing corals and similarly are likely to be damaged and dislodged by divers. To date all transplantation acts were conducted in areas with low human impact. But, as Abelson (2006) mentioned, in cases where stressors persist the outcome of restoration will probably not succeed as expected. Therefore use of transplantation in heavily dived areas may give different results from those previously observed. In this study I monitored three consecutive transplantation events on an artificial reef positioned in the most dived area in Israel. I aimed at (1) observing coral survivorship and (2) proposing alternatives to reduce coral damage. I hypothesized that species of the coral transplants will influence its overall survival and the coral’s ability to self-attach to the AR. My predictions were: (1) that branching corals, having fast growth rates, will have increased self- attachment over massive corals. (2) massive corals will have lower natural (i.e. not anthropogenic) mortality rate and (3) if prediction (1) is correct than branching corals will have lower dislodgement rate than massive corals. It was also hypothesized that Physical protection

of corals from diver dislodgement will influence coral survival rates. It was, therefore, predicted that the use of protective coral measures will decrease their direct effect. This study increases the knowledge for reef conservation and refines the technical knowledge needed for coral transplantation.

5.2 Methods

5.2.1 Study site

A small artificial reef was immersed in October 2006 at the northern Gulf of Eilat, Red Sea (29°32’85’’N, 34°57’47’E; Fig. 5.1). The AR was placed on a flat sandy bottom 100 m from shore and 15 m from nearby coral outcrops, and is located in the vicinity of the actively protected nature reserve. This dive site supports the largest number of divers in Israel and is characterized by beginner divers (Polak and Shashar, 2012). To assess conditions for potential coral recruitment, water currents were measured using an S4 ADW current meter (InterOcean Systems Inc.) at mid-height (2.5 m above sea-bottom) and at the top of the AR (4.5 m above sea-bottom). In general, the area sees little surf or strong current action other than periodic southern storms. Sea surface temperature was measured 3 m beneath the sea surface and data were obtained courtesy of the Israel National Monitoring Program at the Gulf of Eilat.

5.2.2 Artificial reef design

The AR was composed of six 2×2×2 m units of concave hollow concrete Ocean Brick™ blocks (Fig. 5.2). Each module weighted 4.2 tons (weight on land) and was pre-drilled to accommodate coral transplants of matching stub size. The AR design was planned to allow for passing of currents and waves, for great stability, for increase in void space, and lighted and shaded areas. To favor coral settlement, the concrete’s pH was treated to reduce acidity, pumice of high microhabitat complexity was inserted, and the faces of the substrate were sandblasted to increase microhabitat complexity.

Figure 5.1 Study site and location of AR deployment. The AR was located near the nature reserve. House icons represent diving centers.

a b c

Figure 5.2 Design of the AR. The structure (a) is composed of modular units that can be configured according to desire. The structure encompasses vertical, shaded, and curved micro-habitats. (b) Surface with pre-drilled holes that match the pipes corals grow on at the nursery. A close-up of the rough micro- texture of the AR is seen in the front. (c) An image of a typical peg deterrence experimental area on the AR used to exclude divers from transplanted corals.

5.2.3 Coral preparation

Corals taken from objects intended for removal from the water and remains of coral experiments were transplanted between the years 2004-2008 onto a coral nursery located near

the Inter-University Institute for Marine Sciences in Eilat, under a special permit from the Israeli Nature Reserves and Parks Authority. The corals were first placed in a large water container (1 m3) in the laboratory where they were trimmed to fit into irrigation pipes, 16 mm in diameter. This irrigation pipe’s diameter matches the diameter of holes pre-drilled in the AR to accommodate the coral transplants. The corals were grown in situ on suspended floating mid- water nurseries in the sea at a depth of 3–8 m below water surface (see Shafir et al, 2006). Some colonies that grew to approximately 10 cm in diameter were transferred to nearby anchored nursery tables 1 m above the sand flat to avoid dislodgement. Corals that grew tissue on the pipe and reached >7 cm in diameter were deemed proper for transplantation.

5.2.4 Transplantation

Three transplantation events were conducted between 2007- 2010 (Table 5.1). Corals were transplanted in densities of 5 and 10 corals m2. All transplantations were conducted on flat surfaces of the AR (Fig. 5.2) and in all orientations (North, South, East, and West). Coral configuration varied between surfaces and was classified to four configuration treatments: SP- low containing five Stylophora pistillata colonies per m2, SP-high containing 10 S. pistillata colonies per m2, PV-high containing 10 Pocillopora verrucosa colonies per m2, and a mix containing five corals of each species (10 colonies per m2). The first transplantation was carried out in September 2007, six months after AR deployment. Each coral configuration treatment had four replicates, each located at a different orientation. The second major transplantation was conducted in January 2008 on the same location of the first transplantation with the addition of 36 corals to accommodate for dead and missing corals. Additionally, 20 more P. verrucosa corals (two replicates of the PV-high treatment) were transplanted on two new surfaces. Since such large replacement of corals was conducted I decided to refer to this coral replenishment as an independent transplantation event. The third transplantation was conducted in March 2010 on 36 flat surfaces, either horizontal or vertical and representing all locations in the AR. In all transplantations I recorded the number of live, dead (<10% live tissue cover), and missing corals on the examination plots. Additional 428 branching, 93 massive and 52 soft corals, were transplanted outside the experimental surfaces including curved surfaces (Appendix Table A.4). The transplantations were made by experienced and non-experienced divers alike (Table 5.1). Surface preparation of the AR included re-drilling the holes to eliminate fouling. The corals

were transferred, submerged in sea water, via a boat, from the nursery to the AR location 1 km to its north. The corals were trimmed to fit firmly on the substrate allowing for at least one point of live tissue to touch the substrate, then a small two-part epoxy putty (Aquamend®, Polymeric Systems Inc.) was added on the back of the irrigation pipe, and the transplant was inserted in the appropriate pre-drilled hole (see movie at http://www.youtube.com/watch?v=grIeDIe9xKI).

5.2.5 Coral attachment

To examine the ability of new transplants to attach onto the AR I conducted three observation dives on 15 December..2009, 24 September 2010, and 07 October 2010, where all transplanted corals on the AR were checked for existence of tissue growth on the AR substrate. The number and species of coral were recorded.

5.2.6 Natural settlement of new recruits

To follow the natural recruitment of new coral recruits I used a coral fluorescence count technique (Piniak et al., 2005, Baird at al., 2006). I used a BlueStar blue LED torch (NightSea Inc.) and filter that allow for the fluorescence of the corals to be detected. Fifteen 0.5 m2 quadrates were chosen, representing all faces and heights of the AR. These exact quadrates were revisited every three months, at night, and coral recruits were counted. Since the exact location of the new recruits was not mapped and death of coral recruits is known to be prevalent (Smith, 1992), I counted the cumulative number of recruits and coral spats, that developed from earlier recruitments, on the sampling areas.

5.2.7 Manipulation experiment to deter divers

To calculate the effectiveness of avoiding direct contact with the AR corals, I manipulated seven flat surfaces of the AR by securing 30 cm long pipes (herein called pegs) protruding from the surface. Three to five pegs were added to each face depending on hole availability (Fig. 5.2c). Four other bare surfaces located adjacent to the manipulated surfaces, were used for control. Two weeks after peg placement, and after checking for their endurance to divers’ contact, a monitoring of the number of live corals and dead corals was done on the peg- filled plots. Sampling was conducted every 2–3 weeks (except for a missed sample on 06 May 2009).

5.2.8 Data analysis

To decipher the effect of the coral transplantation regime on coral type, ANCOVA was used controlling for time (days) as covariate with type of coral as fixed variable, and live, dead, or missing corals as the dependent variables. Since the number of corals used in each coral configuration transplantation treatment was different, data were converted first to percentages and later arcsine transformed. Coral recruitment was first multiplied by four to adjust data to 1 m2 and later tested with regression ANOVA (while I present the averages in the results; Fig. 5.3). Water current measurements were tested with Student’s t-test. Lastly, deterrence of divers from dislodging corals was analyzed with chi-square test.

5.3 Results

5.3.1 Comparison of coral transplantations

I conducted three major transplantations. In the first two events I transplanted only two branching species and in the last experiment I transplanted 10 species, six of which were massive corals (Table 5.1). The first transplantation exhibited similar rates of survivorship in the two corals species examined (ANCOVA, F(1,123)=1.027, p=0.313; Fig. 5.3a). In the second transplantation (Fig. 5.3b) a decreased survivorship (62%, n=120) was observed in the initial days after transplantation that leveled off after the first measurement date (70 days after transplantation). The rate of survival was different between P. verrucosa and (ANCOVA,

F(1,197)=21.647, p<0.001) probably as a result of high death after transplantation. Rate of survival in the third transplantation of the branching corals was similar between the four coral species

(ANCOVA, F(3,20)=2.185, p=0.121; Fig. 5.3c) and survival in the corals S. pistillata and P. verrucosa was similar as well (Sidak Pairwise comparison, p=0.692).

110 25 1 S. pistillata a d 100 P. verrucosa 20 Acropora sp. 90 M. dichotoma 15

80 10

70 5

60 0 0 20 40 60 80 100 120 140 0 20 40 60 80 100 120 140

110 50 2 100 b e 40 90

80 30 70

60 20

No. of corals Survival (%) 50 Dead 10 40 Missing

30 0 0 50 100 150 200 250 300 350 0 100 200 300

110 50 3 100 c f 40 90 30 80

70 20 60 10 50

40 0 0 50 100 150 200 250 0 50 100 150 200 250

Day after transplantation Figure 5.3 Survivorship (a-c) and number of dead and missing corals (e-g) of branching corals in transplantation events 1, 2, and 3 conducted on the flat surfaces of the AR. Increase in survival occurred as a result of casual divers, particularly dive instructors, reattaching corals (personal communication with various instructors in the study site) Note change in the scales. Trend lines in e-g are either linear (dashed) or logarithmic (solid).

Table 5.1 Details of the three transplantations including percent survival of coral at end of monitoring period.

Date of Initial End Missing Missing Colony Coral Survival Transplanter’s trans- End date Orientation Species corals corals corals corals morphology size (%) experience plantation (n) (n) (n) (%)

21 Sep. 31 Jan. 2008 Vertical Stylophora pistillata Branching Large 80 53 66.3 14.0 17.5 Inexperienced 2007 131 days Pocillopora verrucosa Branching Large 40 26 65.0 5.0 12.5 Total 120 79 65.8 19.0 15.8 01 Jan. 04 Jun. 2008 Vertical Stylophora pistillata Branching Large 80 69 86.3 8.0 10.0 semi- 2008 After 133 days Pocillopora verrucosa Branching Large 60 28 46.7 4.0 6.7 experienced Total 140 97 69.3 12.0 8.6

04 Nov. 2008 Stylophora pistillata Branching Large 80 45 43.8 20.0 25.0 After 286 days Pocillopora verrucosa Branching Large 60 24 60.0 20.0 33.3 Total 140 69 50.7 40.0 28.6 05 Mar. 27 Jul. 2010 Vertical and Stylophora pistillata Branching Medium 80 69 86.3 8 10.0 very 2010 After 144 days horizontal Pocillopora verrucosa Branching Medium 19 14 73.7 4 21.1 experienced Acropora spp. Branching Medium 16 10 62.5 6 37.5 Millepora dichotoma Branching Medium 15 13 86.7 2 13.3 Porites lutea Massive Medium 14 14 100.0 0 0.0 Platygyra daedalea Massive Medium 10 8 80.0 0 0.0 Psamocora Massive Medium 1 1 100.0 0 0.0 Favites pentagona Massive Medium 3 3 100.0 0 0.0 Favia favus Massive Medium 1 1 100.0 0 0.0 Goniastrea spp. Massive Medium 8 8 100.0 0 0.0 Total 167 141 84.4 20 12.0

24 Sep. 2010 Stylophora pistillata Branching Medium 80 56 70.0 22 27.5 After 203 days Pocillopora verrucosa Branching Medium 19 14 73.7 5 26.3 Acropora spp. Branching Medium 16 8 50.0 8 50.0

Millepora dichotoma Branching Medium 15 12 80.0 3 20.0 Porites lutea Massive Medium 14 12 85.7 2 14.3 Platygyra daedalea Massive Medium 10 7 70.0 1 10.0 Psamocora Massive Medium 1 1 100.0 0 0.0 Favites pentagona Massive Medium 3 3 100.0 0 0.0 Favia favus Massive Medium 1 1 100.0 0 0.0 Goniastrea spp. Massive Medium 8 7 87.5 1 12.5 Total 167 121 72.5 42 25.1

Survivorship (Table 5.1) of all corals combined was similar between the first two transplantations at days 131–133 (65.8% and 69.3% for the two transplantations, respectively). S. pistillata exhibited higher survivorship than P. verrucosa (86.3% and 46.7%, respectively, in the second transplantation; Table 5.1), Survivorship between days 133 and 286 in the second transplantation resulted in only 18.6% reduction in overall corals. Overall survivorship in the third transplantation event was even higher (84.4% at day 144) and coincided with higher level of transplanter’s experience and smaller coral size. There is a negative correlation between time and coral survivorship; thus the longer the time of the monitoring the fewer corals survived. In all transplantations, the rate of coral death was high after transplantation and later slowed. On the other hand, the rate of corals missing from the experimental transplantations was positively linear (Fig. 5.3 e, f, g). In the third transplantation the rate of corals missing was very high compared to the rate of coral death rate. There is a relationship of transplanter’s level of experience to coral size, but not to rate of missing corals. In the third transplantation I observed that the accumulated massive corals had a slower mortality rate than branching corals (ANCOVA, F(1,10)=72.627, p<0.001; Fig. 5.4a) and the rate of corals missing from the experiment was lower for the massive corals as well (F(1,10)=72.627, p<0.001; Fig. 5.4b). 140 a 120 100 80 Branched Massive 60 40 20 0 40 b

No. of corals 30

0 0 50 100 150 200 250

Days after transplantation Figure 5.4 Comparison of (a) dead and (b) missing corals in branched and massive corals in the third transplantation.

5.3.2 Coral self-attachment

Coral attachment to the AR substrate was dependent on species (Table. 5.2). The fastest to attach to the concrete surface was Millepora dichotoma (I consider this Hydrocoral as coral for the purpose of this study). Of the top five most attached corals, two were branching corals and three were massive ones (although the coral Cyphastrea chalcidicum should be regarded with caution due to very few repetitions). Of 13 coral species transplanted only three species did not show any signs of attachment; all three were massive corals. The most common species found in Eilat (Stylophora pistillata, Pocillopora verrucosa, and Acropora spp.) that were also most transplanted on the AR all observed attachment rates were higher than 28%. These data include both the first and second transplantation and spans a period of three years. Attachment of corals between flat and curved surfaces (curved surfaces were more sheltered from diver’s reach) showed no difference (χ2=3.29, df=10, p=0.97).

5.3.3 Natural settlement of new recruits

Accumulated coral recruits displayed a stepwise increase (Fig. 5.5), which coincides with the seasonal change in the water temperature. The increase appeared during the late summer with no recruits joining during the winter, outside the coral spawning season. Overall increase was linear (regression ANOVA, r2=0.23, regression coefficient=0.04, p<0.001). Further examination revealed higher recruitment rate at the leeward surfaces (ANCOVA, F(1,55)=11.519, p=0.001), the upper surfaces were marginally significant higher recruitment as well (ANCOVA, F(1,218)=3.323, p=0.07). Similarly, currents were higher on the upper location of the AR (4.5 m above sea bottom) than the lower location (2.5 m above sea bottom; t-test for unequal variances, t=4.073, p<0.001). Recruitment rate was similar between North and South surfaces (ANCOVA, F(1,

101)=0.403, p>0.5).

100 35

80 Temperature (C) 25

-2

60 20

15 40

# m of recruits 10

20 5

0 0 1/1/2007 1/7/2007 1/1/2008 1/7/2008 1/1/2009 1/7/2009 1/1/2010 1/7/2010 1/1/2011 1/7/2011 1/1/2012

Figure 5.5 Density accumulation of coral recruits (±SE) onto the AR in 1 m2 using fluorescent light methodology (see text) in correspondence to monthly sea surface temperature in Eilat (±SD).

5.3.4 Manipulation experiment to deter divers Using pegs for deterring divers from dislodging corals did not reduce the overall amount of corals missing from the experimental plots over 186 days surveyed (χ2=3.33, df=2, p=0.35). But comparing the number of missing massive corals from all faces monitored either with or without pegs resulted in less massive corals missing from the experimental faces (χ2≤10.7, df=2, p≤0.05, pair wise comparison with Bonferroni correction).

Table 5.2 Attached corals on the AR at flat and curved surfaces. Only corals that had more than two individuals are presented. Chi-square test was used to evaluate the difference between the flat surfaces over time and between the overall averages of the flat surfaces to the curved ones . Difference on curve surfaces is the difference between the average percent of the flat surfaces to percent curved surfaces.

Flat surfaces Curved surfaces 15 Dec. 2009 24 Sep. 2010 07 Oct. 2010 Colony Avg. Species morphology Attached n % Attached n % Difference % Attached n % Difference Stylophora pistillata Branching 30 101 29.7 19 56 33.9 4.2 31.8 25 53 47.2 13.2 Acropora spp. Branching 14 35 40 7 10 70 30 55 14 38 36.8 -33.2 Pocillopora verrucosa Branching 9 31 29 4 14 28.6 -0.5 28.8 5 13 38.5 9.9 Porites lutea Massive 7 28 25 5 11 45.5 20.5 35.2 0 5 0 -45.5 Millepora dichotoma Branching 11 13 84.6 11 15 73.3 -11.3 79 - - - - Platygyra daedalea Massive 4 11 36.4 5 9 55.6 19.2 46 - - - - Goniastrea spp. Massive 2 10 20 0 4 0 -20 10 - - - - Favia favus Massive 2 7 28.6 - - - 28.6 - 4 6 66.7 38.1 Favites spp. Massive 0 5 0 0 3 0 0 0 - - - - Cyphastrea chalcdicum Massive 1 2 50 1 2 50 0 50 - - - -

Chi Square test χ2=2.421 df=10 p=0.992 χ2=3.286 df=10 p=0.974

5.4 Discussion

Divers had a great impact on the survival of corals. The use of branching corals has resulted in high death and coral dislodgement rates and moderate attachment rate. Massive corals are found to be easily attached to the AR (at least some of the species), dislodge less rapidly from the reef, and have higher survivorship rate. Guest et al. (2011) found that massive corals attach at a slower rate than branching corals but have better survivorship. My hypothesis that fast growing corals will attach faster was dismissed, as a similar rate of attachment to substrate by some massive coral species was documented. It was additionally observed elsewhere that massive corals are important builders of natural reefs (Shaked et al., 2005) and thus should be integrated into coral transplantations on ARs (Ortiz-Prosper et al., 2001). Edwards and Clark (1998) have claimed that after five years of monitoring in highly disturbed shallow reefs, the use of branching transplants was less beneficial for restoration purposes, and they argued that massive corals should be transplanted as they have better survivorship and less dislodgement. I concur with their conclusions and recommend that in heavily dived sites, the use of massive corals, despite their relatively low growth rate, should be encouraged (but see Chapter 3 for the effects of coral structure on fish biodiversity).

5.4.1 Transplantation and survival

I observed high coral mortality and dislodgement from the AR. While mortality rate decreased over time, coral dislodgement was continuous, making it a potential detrimental factor in coral survivorship. Death of coral transplants is common and is estimated to vary at a range of 50-100% (Rinkevich, 1995). Transplants death was recorded to peak after transplantation and to decrease thereafter (Yap et al, 1992). These figures and pattern agree with the death rate observed in this study. In particular, I observed a high death rate (>50%) in the branching coral Pocillopora verrucosa. A high mortality rate of coral transplants on ARs was also found in Sulawesi, Indonesia (Ferse, 2008), where cumulative survival of P. verrucosa was <20% after a year, partially a result of storms.

Dislodgement of corals is usually related to high wave action (Edwards and Clark, 1998; Abelson, 2006; Ferse, 2008; Shaish et al., 2010; Guest et al., 2011). Coral dislodgement on coral nurseries was documented to have >20% of the corals in intensive nurseries (Shafir et al., 2006). In the current study coral dislodgement reached >50% after one year and the trend of dislodgement appeared linear, thus dislodgement was expected to continue.

Moreover, coral dislodgement equaled and exceeded natural coral mortality, thus making divers’ physical impact on transplanted corals the primary cause of death. Coral self- attachment was similar between flat and curved surfaces (areas where divers were more reluctant to enter). Guest et al., (2011) found that after 221 days of monitoring attachment ranged between 60-100% depending on the coral species. The current study showed much lower percentages of self-attachments (0-73.3%). In both studies Acropora sp. was the fastest to attach. Unlike Guest et al., (2011), the current study showed no difference in coral self- attachment between coral growth forms (i.e. branching and massive) .It was observed that the P. verrucosa attachment rate was relatively low (Table 5.2), as observed in the Philippines (Guest et al., 2011), supporting the idea that P. verrucosa is prone to easy dislodgement due to the long time needed to attach to substrate. It appears that there are different survivorship, attachment, and dislodgement rates between corals species (Ferse, 2008; Shaish et al., 2008). Therefore, if coral transplantation is to be conducted, one should choose the appropriate corals for the specific location. Interestingly, the use of experienced labor for the transplantation did not show reduction in the rate of coral dislodgement, but only in coral mortality rate, which could be incidental. Therefore, the use of volunteers and non-trained divers in transplantation events is possible and could also reduce transplantation costs.

5.4.2 Recruitment

Natural recruitment on the AR was seasonal and composed of mainly Pocilloporidae (soft corals were disregarded from the analysis) and few massive coral recruits. This coincides with other studies conducted in Eilat (Glassom et al., 2004; Perkol-Finkel and Benayahu, 2007). Increased water current measurements on the top part of the AR coincided with increased natural coral recruitment in contrast to a study conducted nearby (Perkol-Finkel et al., 2006). Recruitment rates were similar to another study in Eilat that measure recruitment on settlement plates and the location of the AR in close proximity to natural reef may explain the that rate of recruitment (Glassom et al., 2004). All other location on the AR highly varied and were consistent with the high variability found by Glassom et al., (2004). Difference in natural recruitment into various locations on the AR suggests that in designing ARs, such factors should be considered before AR immersion and particularly if transplantation of corals is planned. Preliminary recruitment measurements were suggested for harvesting the full potential for coral recruitment on ARs and to help locate optimal spot for AR deployment

(Shemla, 2002). It appears that the dislodgement rate of naturally recruited corals is lower (O.P. personal observation). Therefore rendering the coral transplants less durable than natural recruits. Since ARs that are intended for attracting divers they need to be attractive and hence containing large numbers of versatile corals (see Chapter 6). Therefore the use of fast growing and fast recruiting coral transplants should be limited, and time for recruitment of natural corals should be allowed.

5.4.3 Impact of divers on transplantation

Divers’ behaviors may reduce coral damage. It was already demonstrated that dive briefings and presence of dive leaders can reduce direct coral contact (Medio et al., 1996; Priskin, 2003). Transplantation of corals in proximity to untrained divers is expected to cause damage to the transplants. In the AR location divers inflicted more damage than natural storms. In January 2008 I counted only four corals that were dislodged from the AR after a severe southern storm. On the contrary, in one weekend, where a photography competition took place, 13 corals fell off the AR. The impact divers have on coral dislodgement could be mitigated by a diver management plan. I attached plastic tubes to deter divers from contacting the corals and dislodging them, but my results showed that divers were still able to knock off corals similarly to exposed faces on the AR. This may be attributed to the quantity, distance, and unsuccessful design of the pegs for deterring divers. It is possible that another design may achieve that goal. Another possible approach is to eliminate divers’ presence in the initial period until the corals will self-attach or fence the transplanted area.

5.5 Conclusions

Corals transplanted on the AR were found to attach slowly, dislodge quickly by divers’ contact and exhibited normal survival rates similar to natural reefs and to other locations (Yap et al., 1992; Rinkevich, 1995). The use of massive corals reduced transplant mortality and dislodgement but showed similar rates of self-attachment as branching corals. In contrast, natural recruitment was high and corals naturally recruiting appeared to be more resilient to mortality and dislodgement. It is evident that in the case of deployment of an AR at heavily dived locations, as presented in this study, coral transplantations should favor massive corals and continual restocking of branching corals until natural recruitment will take over. 11

Coral transplantation is time consuming and relatively expensive (Rinkevich, 2005). Likewise, the value of coral reefs is high (Cesar and van Beukering, 2004), even without considering the total economic value of these organisms and the reefs they reside in. In certain locations ARs could fail to recruit corals from nature and may exhibit low coral growth and high coral mortality (Miller et al., 2000; Glassom et al., 2004). Therefore, remediation by creating new artificial reefs with coral transplants should be meticulously observed to decide if it is worthwhile. Additionally, local stressors, such as wave action or divers’ activity, are potential factors reducing the success of growth and self-attachment of corals. It is suggested here that if the stressed environment persists, coral transplantations should be used for longer periods, entailing higher costs and longer time to reach coral cover targets, until natural recruitment matches the transplanted corals or until transplanted corals self-attach. As massive corals have a higher survival rate it is suggested to transplant massive corals over branching ones. In situations where transplantation of corals is necessary to achieve quick coral cover it is recommend to use several large branching corals to achieve AR attractiveness, or alternatively to use diver management measures to protect the corals during their self-attachment period to the AR. Additional studies are needed to define these diver management tools and further knowledge needs to be acquired to ascertain the role of coral species and location in coral transplantation.

6. Economic value of biological attributes of artificial coral reefs

6.1 Introduction

Coral reefs provide humans with a range of services and goods. These include direct use values such as fishing tourism or other forms of recreation, indirect use values such as coastal protection or nursery grounds for fisheries, and the benefit gained from knowledge of the existence and importance of such diverse ecosystems (existence values) (Cesar, 2000; Brander et al., 2007). When focusing on tourism, coral reefs are well-known tourist magnets, with reef enjoyment due to its overall appearance as experienced while diving, snorkeling, boating (including glass bottom boats), recreationally fishing, and more. However, what about the reef do visitors actually value? It is widely accepted that people, especially tourists, value and appreciate natural scenes and habitats (Tietenberg, 2006; Curtin, 2010). However, what part do they most value? The landscape? The general scenery? What value, if any, does the visitor assign the biological component? Several studies that examined the value of changes in reef "biodiversity" failed to observe uniform results for biological attributes (Spash, 2000; Wielgus et al., 2003; Cesar and Van Beukering, 2004; Leujak and Ormond, 2007; Uyarra et al., 2009). The difficulty in isolating the effects of different biological attributes may generate confusion in understanding the contribution each attribute makes to the overall scene. Additional perplexity arises because the term "biodiversity" itself is rarely clear to the general population, and in most cases it is simplistically used as a surrogate to reef species diversity (Spash and Hanley, 1995; Tietenberg, 2006). Indeed, the term biodiversity appears to have a number of definitions and encompasses a broad range of social and natural goods and services that, as benefits, have both instrumental and intrinsic values. Four scales of biodiversity, from genetic up to ecosystem level, are recognized (Turner at al., 1999; Nunes and van den Bergh, 2001). In the context of recreation, the use of biodiversity is usually attributed to the species level (Nunes and van den Bergh, 2001). Biodiversity descriptors may provide some knowledge about the complexity, and therefore the probable resilience, of the community in question (Steneck et al., 2002; Folke at al., 2004). Though well-defined to field researchers, the differences between the variable descriptors and their meanings are rarely clear to the general public.

While the sustainability of biodiversity is costly (Mann and Plummer, 1993), its benefits are usually considered public goods. Since these are free for general use, there is no market price attributed to them, and they are commonly maintained by governments through subsidies or taxation. However, in certain circumstances individuals are willing to subsidize a public good often, though not always, if they have direct use of its resources (Tietenberg, 2006). Hence, the total economic value of a public good includes its direct use value along with (often higher) indirect or external values. These non-use indirect values entail the factors that motivate one to invest in public goods, including existence, bequest, and option (the value that people place on having the option to enjoy something in the future, although they may not currently use it) values (Turpie et al., 2003). A common method of valuating external values is the Contingent Valuation Method (CVM; Tietenberg, 2006), which uses hypothetical scenarios to investigate people’s Willingness To Pay (WTP) to evaluate their use of an amenity, frequently used in environmental studies. An advantage of this approach is that the values obtained supposedly represent both direct and indirect values for the described scenario. During CVM questionnaire design, it is important that each question and the different scenarios will be clear to the respondent and will be set in his/her own conceptual world. An additional consideration is that the different scenarios presented will be comparable. While the CVM has been criticized and reviewed (see Bateman and Willis, 1999), it is still considered an appropriate approach to valuate non-use services and hypothetical situations (Hanley, 2000). Artificial reefs (ARs) are traditionally defined as submerged, man-made structures that affect the local biological community (Seaman and Jensen, 2000; Svane and Petersen, 2001). In recent years the sinking of ships and immersion of dedicated structures for the benefit of the diving tourism has gained popularity worldwide despite the scarcity of meticulous research on the social and economic benefits of these deployments (Stolk et al., 2007) and the complicated process entailed in their economic valuation (Sutton and Bushnell, 2007). However, as artificially designed structures, ARs provide researchers with an ideal setting to perform designed and controlled studies. In my case, an AR was an appropriate setting for a CVM study as it was pre-designed for coral planting and similar manipulations (Polak and Shashar, 2012). Use of ARs can be useful for conservation purposes as they may serve as surrogates for the attractions of natural reefs (Leeworthy et al, 2006; Polak and Shashar, 2012). It was therefore hypothesized that biodiversity of the coral reef community (e.g., or

biological scenario) will influence divers willingness to pay for AR, and thus the non-market value of ARs to the community. Making these ARs more attractive by adding preferred biological attributes such as fish and corals may increase their effectiveness as natural reef surrogates. My study addressed the value of changes in biological parameters on the species diversity level at a single habitat scale. I manipulated and isolated different levels of diversity descriptors, such as richness, abundance, and biodiversity of corals and fish. To differentiate between these diversity descriptors and determine people’s motivation for preferring to pay for the proposed factors, I used the contingent valuation method (CVM). I predicted that visitors were able to monetarily differentiate between different biological scenarios and they were be able to detect varying degree of conservation levels. My goal was to better understand which biological attributes are more appreciated by divers and to evaluate the relative importance of each.

6.2 Methods

6.2.1 Preliminary surveys

To examine people’s appreciation of biological parameters in a reef environment, a CVM survey examining willingness to pay (WTP) for improvement of environmental conditions was design. Before final presentation the survey was pre-tested with two informal focus groups: one comprising seven marine biology students, and another comprising 24 biology and natural science students. The pre-tests allowed us to evaluate and improve survey procedures, the quality and clarity of the survey and the pictures presented therein, and the abilities to note differences between the pictures and adjust bidding levels using open-ended questions, among other indexes. For the final test, care was taken not to approach any students or academics of marine sciences.

6.2.2 Survey design

The survey was divided into four sections (Appendix supplementary material A5). The first section described a hypothetical situation in which the decline of coral reefs was presented and the option to use ARs for replenishment (not replacement) was offered. The background included the rationale to immerse and maintain these ARs and referred to a cost

that these processes may have. The respondents were ultimately presented with the question, “If an NGO will be established, whose only goal will be to maintain and nurture surrogate ARs, and all the money collected will be dedicated to this use only, what is the maximum value, in New Israeli Shekels (NIS), that you will be willing to pay, as a contribution on a yearly basis, for restoring the surrogate ARs to the following proposed state?” If the answer to this question was 0 (not willing to contribute), then the responder was instructed to proceed directly to the fourth section while skipping sections 2 and 3. The second section aimed to find the WTP for conserving specific biological features on these ARs and included the presentation of variations in seven different contributors to biological scenarios in typical AR settings: coral size, coral diversity, fish abundance, coral abundance, a combination of numbers of fish and corals, and fish and coral biodiversity (Fig. 6.2). Each situation was described by hypothetical images representing varying degrees of conservation efforts (Table 6.1), including no effort, small effort, medium effort, or high effort. For example, in the fish abundance scenario there was low, medium, and high numbers of fish in each image (Fig. 6.1).

a b

c d

Figure 6.1 Representation of scenario B (coral and fish abundance) at (a) no conservation effort and at (b) low, (c) medium, and (d) high conservation efforts.

The images were manipulated using standard imaging software (Photoshop™ CS3), which allowed for the control of the specific tested parameters while keeping all other aspects constant (Table 6.1). The background landscapes for the images were of an existing artificial reef (Tamar Reef) that was known to attract divers even without any coral coverage or fish recruitment (Shani et al., 2012; Polak and Shashar, 2012). The images presented how these landscapes should be, while maintaining the images in as realistic a state as possible and while keeping them in line with the concept of normative landscape scenarios (Pitt and Nassauer, 1992; Nassauer and Corry, 2004). The third section comprised questions explaining the motivation for the choices made. In this section the participants were presented with eight reasons to explain the basis of their decision, and they were asked to rate the choices from most to least relevant. The fourth section comprised a short, anonymous, socio-economic background survey. The monetary unit used in the survey was New Israeli Shekels (NIS), whose value at the time of the survey was 3.70 NIS per $US.

a e

b f

c g

d Figure 6.2 Representation of the different biological scenarios (a-g; Biodiversity, numbers of fish and corals, coral species richness, fish species richness, coral size, coral abundance, fish abundance; see Table 1) at their highest conservation level effort (except scenario e, which displays the medium level of conservation for ease of identification).

Table 6.1 Description of the different scenarios (A-G) presented to the participants. The calculated biodiversity descriptors were obtained from the pictures presented. Scenario Description Conservation Coral Coral Fish Fish species Shannon- Pielou Simpson Fisher’s level abundance species abundance richness Weaver (J) (D) α richness a Total Low 32 1 58 2 0.51 0.73 0.33 0.4 Biodiversity Medium 30 8 50 5 1.79 0.69 0.75 4.4

High 29 10 54 13 2.61 0.84 0.89 10.53

b Numbers of fish Low 11 1 18 1 0.66 0.96 0.49 0.49 and corals Medium 20 1 45 1 0.62 0.89 0.43 0.39

High 35 1 160 1 0.47 0.68 0.29 0.31

Coral species Low 33 1 0 0 0 0 0 0.19 c richness Medium 33 8 0 0 1.25 0.6 0.55 3.36

High 33 11 0 0 2.03 0.85 0.83 5.78 d Fish species Low 0 0 58 2 0.51 0.74 0.33 0.4 richness Medium 0 0 58 5 0.99 0.62 0.48 1.31

High 0 0 58 13 1.99 0.77 0.8 5.21

Coral size Low 35 1 0 0 0 0 0 0 e Medium 35 1 0 0 0 0 0 0

High 35 1 0 0 0 0 0 0

f Coral abundance Low 11 1 0 0 0 0 0 0 Medium 20 1 0 0 0 0 0 0

High 35 1 0 0 0 0 0 0 g Fish abundance Low 0 0 18 1 0 0 0 0

Medium 0 0 45 1 0 0 0 0

High 0 0 160 1 0 0 0 0

6.2.3 Data collection and analysis

The sampling event was conducted during a local holiday vacation between 29 March and 8 April, 2010, on the shores of the Gulf of Eilat, Israel. Eilat is a beach resort town at the northern tip of the Red Sea with a thriving diving industry and a flourishing local coral reef. In total, 305 participants answered the questionnaire. The survey was presented at or near diving centers (to insure that responders had realistic knowledge about the local reefs and their inhabitants). Therefore, all responders were either divers or people somehow connected to the diving industry. Hence, I refer to them as divers and note that they do not represent the general population. The order of the scenarios was presented three different ways. Preliminary analysis (n = 21, 36, 39 for each of the three presentation orders) showed no difference in the order in which the scenarios were presented (nested ANOVA, F(2,1922) = 0.49, p > 0.6). Data of protest responses (those that chose only one answer in section three to be either response E or H, see supplementary material, n = 11), outliers whose answers were over four times the average values, and participants replying with repeated, illogical answers were eliminated. Data was presented using the median statistic to use conservative estimates of the WTP. Potential differences between the different scenarios were examined using nested ANOVA where the conservation levels were nested within the different scenarios. Data was log transformed to meet the test’s assumption.

6.3 Results

6.3.1 Descriptive statistics

The survey yielded 305 interviews, of which 261 (85.6%) were suitable for further analysis (Table 2). The data included 97.7% Israelis, most of who were not locals (locals were defined as Eilat and Arava municipality residents- 81.3%). It contained mostly single (67.3%) men (76.7%) with no children (75.1%) between the ages of 25 and 35, inclusive (54.1%). Diving level comprised mostly beginners to advanced-open-water level divers (71.5%) with less than 6 years experience in the sport (54.6%) who had conducted fewer than 20 dives per year (58.1%).

Participants were typically full-time workers (39.9%) or students (26.5%), with or in the progress of completing an academic degree (62.3%), and with lower than average income (49.8%).

Table 6.2 Descriptive statistics of the interviewed participants. Mean, standard deviation and median WTP (in NIS) are in respect to the highest level of conservation in the highest biodiversity scenario (Scenario A –Total biodiversity). Variable Category N % Mean WTP SD Median No. of years certified 0 − 2 85 28.4 99.5 86.6 80 as a diver 3 − 4 45 15 104.2 85.0 100 5 − 6 34 11.4 118.8 99.6 97.5 7 − 8 23 7.7 84.1 61.1 65 9 − 10 32 10.7 91.6 91.6 57.5 10 − 12 19 6.4 83.2 66.5 80 12 − 14 18 6 63.8 59.4 50 14 − 16 13 4.4 128.8 103.5 100 > 16 31 10.4 85.7 80.8 57.5

Certification level Open water 8 2.6 60.6 42.5 65 Advanced 208 68.9 93.99 83.0 70 Divemaster 45 14.8 104.5 90.9 75 Assistant Instructor or higher 40 13.8 101.3 104.1 70 Number of dives/year Up to 20 (including 0) 173 58.1 78.5 73.4 50 20 − 40 44 14.8 117.4 101.1 100 40 − 60 36 12.1 129.7 91.4 100 60 − 100 21 7 125.2 75.9 100 > 100 24 8.1 93.7 91.0 52.5 Residence Local (Eilat and its region) 50 16.4 93 101.7 50 Domestic (Israelis) 248 81.3 95.8 81.2 77.5 Out of state 7 2.3 125.7 73.9 100 Gender Male 234 76.7 92.7 84.0 70 Female 71 23.3 107.1 86.5 92.5 Age 20 − 30 117 39.8 99.4 85.2 80 31 − 40 107 36.4 71.3 66.3 50 41 − 50 50 17 110.1 80.6 100 > 51 20 6.8 165 182.4 82.5 Marital status Single 202 67.3 94.4 78.7 80 Married 85 27.3 104.1 98.2 70 10

Other 16 5.3 84.7 90.86 37.5 No. of children 0 226 75.1 93.3 77.9 75

1 12 4 127.5 141.7 60 2 30 10 111.3 106.8 70 3 25 8.3 87.6 84.4 65 > 4 8 2.7 96.9 82.0 50

Occupation Student (marine science 79 26.5 89.1 70.2 77.5 students were asked not to participate) Soldier 37 12.4 87.2 81.2 50 Part time 17 5.7 90.0 115.3 30 Full time 119 39.9 97.9 80.1 70

Independent 32 10.7 128.1 120.0 100 Pensioner 3 1.0 96.7 132.9 25 Unemployed 1 3.7 86.8 70.7 100 Education Elementary 31 10.6 131.0 103.5 110

High school 52 17.8 85.7 75.3 70 Technical 27 9.2 108.5 83.7 100 Academic 182 62.3 92.0 83.0 70 Income Low 146 49.8 91.3 78.9 70 (by own perception) Average 48 16.4 95.6 90.8 70

High 99 33.8 103.1 90.0 70 Affiliation to a Belongs 38 12.8 111.4 86.5 90 "green" organization Does not belong 258 87.2 93.7 84.3 70

Visits per year 0 − 2 123 48.4 84.4 74.5 84.4 3 − 6 93 36.6 96.1 76.2 85 7 − 10 8 3.1 137.5 81.2 140 > 10 30 11.8 116.9 106.4 95

6.3.2 Willingness to pay

Divers consistently differentiated between the conservation levels within each scenario

(nested ANOVA F(21, 5741) = 34.7, p < 0.001, Fig. 6.3). Conservation level corresponded roughly with an increase in ecological indexes, e.g., numbers and diversity of species, number of individuals, etc. Median WTP for the bare AR (no additional conservation efforts) was 0 NIS, while those for higher levels of conservation increased progressively with each level: 10 − 35 11

NIS for low effort, 15 − 50 for medium effort, and 25 − 70 NIS for the highest effort. Nested

ANOVA also revealed that this overall response was similar in all scenarios (F(6, 5741) = 0.72, p > 0.64). Post-hoc analysis revealed that scenario A (total biodiversity), which has the highest biodiversity index (Table 6.1), also had the highest WTP (Tukey HSD, p < 0.001). Scenario G (fish abundance) had a lower WTP (Tukey HSD, p < 0.001) than all other scenarios. The remainder of the scenarios differed insignificantly from each other (Tukey HSD, p > 0.05). Analysis of the WTP at the different biological (conservation) levels revealed that divers were able to differentiate between the different biological levels (Tukey HSD, p < 0.001 between all levels). 80 a High b Medium Low None 60 bcd cd de bc 40

Median WTP (NIS) 20

0 A B C D E F G

Scenario

Figure 6.3 Median willingness to pay for different biological scenarios (A-G, see Table 1) at no (◊), low (■),medium () , and high (●) conservation efforts. Lower case letters represent post-hoc Tukey HS, D grouping; p < 0.05.

63.3 Motivation

A total of 716 answers were collected from 295 participants in response to the motivation questions. Most divers reported a motivation to pay for the public goods as presented (60.1%, n = 295), while the rest stated that their reasons for donating money were for their own benefit. The former group (60.1%) could be divided according to divers who wanted to ensure the continued existence of the reef (31.3%), divers who felt that the reef’s inheritance by future generations must be insured (26%), and divers who cited the option to visit the reef in the future 11

(2.8%). An examination of the top-rated motivation indicated that most divers were willing to pay for the existence of conserved ARs (56.3%), with similar importance attributed to maintaining use/options and to inheritance values (22.7% and 20%, n = 295).

6.4 Discussion

This study’s aim and uniqueness was in discriminating between the biological attributes of an artificial reef and its physical structures in people’s perceptions. While the structure itself was already familiar to the divers, both from their prior visits to the reef (Shani et al., 2012; Polak and Shashar, 2012) and from photographs, they were asked to attach a monetary value to its fish and coral inhabitants. Questionnaire responders were asked to report their willingness to contribute to adding different biological sections without altering the overall structure. Therefore, I was able to examine their appreciation of different aspects of the reef’s fauna.

In comparing people’s responses to fish vs. to coral alone, an interesting trend appeared: While Uyarra et al. (2009) documented that divers were able to differentiate between attractiveness of marine biological attributes but that they failed to rate the relative importance of each, in my case divers chose to pay more for fish species richness over abundance, but the same did not hold for corals. Additionally, when looking at abundance alone, divers rated corals with a higher value than fish. This incompatibility can be explained by the appearance of fish of the same species as more uniform than corals of the same species, as the latter may present a more diverse appearance than the former. Underwater structural complexity was shown to contribute to AR attractiveness in the divers’ perceptions (Shani et al., 2012). Preferences of corals over fish by reef visitors may be related to the belief that the investment in corals is of higher importance than that in fish (Shafer and Inglis, 2000). However, the opposite trend, of a preference for fish, has also been documented (Williams and Polunin, 2000; Leujak and Ormod, 2007). These differences in preferences may arise due not only to differences in local relative abundance, but also to demographic differences such as nationality and gender (Leujak and Ormond, 2007; Uyarra et al., 2009). When examining the motivation for the diver willingness to pay, 39.9% stated their own benefit as the primary reason for their willingness to donate money. The rest explained their choices out of a desire to contribute to the general public good. The investment of divers in ARs 16

that carry a conservation connotation demonstrates that divers are willing to pay significant amounts for public goods that only partially benefit them directly. Similarly, several papers demonstrated that people are willing to pay for public goods by assigning a high WTP for entrance into Marine Protected Areas (MPA) with a stated overall goal of marine conservation (Dixon et al., 1993; Seeprachawong, 2003; Peters and Hawkins, 2008; but see Ahmed et al., 2007). It may well be that the diving industry should be considered a form of mass ecotourism (Shani et al., 2012). As such, its capacity to actively promote conservation efforts of marine habitats should be further developed. Non-specialist visitors were able to differentiate between biological descriptors and subsequently assigned different values to them. Divers preferred scenarios with higher biological biodiversity. Important to everyday life, biodiversity and investments made to support it have implications for human health (Newman et al, 2008), the discovery of new medicines (Roopesh et al., 2008), the quality of the world future generations inherit (Abramovitz, 1991), and the maintenance of ecosystem services (Costanza et al., 1997) and resilience (Peterson et al., 1998). Therefore, recognizing the presence of ecosystems is deemed highly important. The ability to accurately recognize and measure biodiversity is usually attributed to professionals such as ecologists. Therefore, it is encouraging to find that the differences between the biological descriptors used in this study were intuitively clear to the non-professional visitors, who identified these differences with precision. Divers were willing to pay the highest sums for conservation efforts that protected higher biodiversity. This entails that they understood, without being told beforehand, that sustaining higher biodiversity levels may have significant ramifications for the reef ecosystem and that it incurs high conservation costs. This ability of non-specialists to recognize, and their willingness to invest in, biodiversity – an ability probably grounded in anthropomorphic and anthropocentric points of view – reflects the importance ascribed to biodiversity by the general public (Martín-López et al., 2007). This study was based on scenarios in which efforts are made to increase (or improve) different biological parameters on an AR. Although it is clear that such scenarios are realistic on the scale of the single AR, their effect on large scales seems improbable. ARs introduce new habitats, which can be better suited to specific organisms due to AR physical parameters, such as geometry (hiding places) and substrate type from which they are made (e.g., settlement of specific fauna) (Pickering and Whitmarsh, 1997), and may result in a specific community evolving around it (Arena et al., 2007). Increase of fouling on ARs causes increase in the 11

associated organisms on ARs (Svane and Petersen, 2001), which though are part of the local species, may be different in composition and relative abundance from the natural reefs (Perkol- Finkel and Benayahu, 2004, 2007). Toward coral reef management, it was suggested that the transplantation of corals may increase local biodiversity and catalyze natural succession (reviewed in Rinkevich, 2008). Therefore, with proper planning and design, it is reasonable to expect ARs to increase local biodiversity, as well as coral and fish abundance. As the trend of immersing artificial reefs increases, the appeal of ARs is critical for their ability to attract visitors (Shani et al., 2012) and equally to divert visitor pressure from natural reefs (Polak and Shashar, 2012). In designing such reefs, as in marine conservation projects aimed at attracting the public, it is important to know the biological characteristics that are most valued by the potential visitors. Though focusing on key species is common and often justified, it is shown here that diversity in itself is perceived to be an important attribute. The use of transplanted corals in natural coral reefs and ARs and the design of ARs to increase fish abundance may augment reef attractiveness, thereby increasing overall reef aesthetics. My results present an important step in pre designing ecotourism focal points that are aimed to divert attention of the public away from more conservation-sensitive areas.

7. Can a small artificial reef reduce diving pressure from a natural coral reef?

7.1 Introduction

Coral reefs are slowly declining around the globe (Wilkinson, 2004). Over 19% of reefs worldwide have disappeared in the last 60 years, and it is presumed that 20% of the world’s remaining reefs, which are currently considered threatened, may be lost by the year 2050 (Wilkinson, 2004). Reef decline can be attributed to a range of causes (reviewed in Dubinsky and Stambler, 2011), such as outbreaks of coral predators, like the crown of thorns (Acanthaster planci) (Bellwood et al., 2004) or the corallivorus gastropod (Drupella cornus) (Shafir et al., 2008), or due to coral diseases (Santavy et al., 2005), severe storms, and tsunamis (Wilkinson et al., 2006), or to indirect stressors, such as climate change (Hoegh-Guldberg et al., 2007). On the local scale, direct interventions by man are often major factors affecting the survival of coral reefs. These include the introduction of freshwater and sewage effluents, increased sediment from mainland construction, ship fuel spills, and coral breakage by SCUBA-divers (Richmond, 1993; Wilkinson, 2004). Coastal tourism is a rapidly growing business comprising a vast and sudden increase in water-related activities (Davenport and Davenport, 2006). SCUBA-diving has been found to be one of the fastest growing coastal tourism activities, and today, there are nine times more certified divers than there were in 1980. (PADI, 2010). Eilat, which is located between the desert and the Red Sea, is witnessing a constant growth in tourism, particularly beach tourism, in parallel with the growing development and subsequent increased tourism in the Sinai Peninsula (Hasler and Ott, 2008). An increase of over 50% in beds occupied by foreign visitors and of 4 % occupied by Israeli tourists has been observed in the last six years (Hayun and Alfia, 2004-2008). Diving tourism accounts for approximately 10% of the tourism in the area (Wilhelmsson et al., 1998), and consequently, an increase in diving tourism is also noticeable. SCUBA-diving is often considered to be a form of marine ecotourism (e.g., Cater and Cater, 2001), and the impact of diving was thought to be mild and localized. Previously, diving activity was considered a low impact activity, in terms of its effect on coral reefs, compared to

storms and waves (Tilmant and Schmal, 1981). The dramatic increase in diving worldwide in the last 30 years (PADI, 2010) has changed this attitude (Harriot et al., 1997). Damage caused by recreational divers has been documented as a significant cause for coral reef deterioration in specific, highly visited locations (Hawkins et al., 1999; Schleyer and Tomalin; 2000, Tratalos and Austin, 2001). Indeed, SCUBA-diving is known to have a negative impact on coral reef cover (Davis and Tisdell, 1995; Harriot et al., 1997; Hawkins and Roberts, 1997; Zakai and Chadwick- Furman, 2002; Hasler and Ott, 2008). Zakai and Chadwick-Furman (2002) mention that hard coral breakage may increase from the 15.5% observed in sites with low diving pressure to 80% in highly dived sites. Divers cause damage to the reefs through fin impact, sediment stirring, reef trampling, kneeling on benthic organisms, touching coral, and hitting coral with loose equipment (Zakai and Chadwick-Furman, 2002; Hasler and Ott, 2008, Luna et al., 2009). Although the damage is presumed to be unintentional (Uyarra and Côté, 2007; Hasler and Ott, 2008), it inevitably harms corals, especially the more fragile, branching corals (Rouphael and Inglis, 1997; Zakai and Chadwick-Furman, 2002; Uyarra and Côté, 2007, Hasler and Ott, 2008). This is subsequently followed by a decrease in the numbers of reef-associated organisms (Jones et al., 2004). Instructional dives such as introductory, course, or refresh dives, are known to cause substantial damage to the natural reef (Zakai and Chadwick-Furman, 2002). The Nature and Parks Authority in Eilat, therefore, invests efforts in diverting divers from the natural reef reserve to alternate areas that are less ecologically sensitive. When addressing the negative effects divers have on coral reefs, one must bear in mind that diving is a form of soft ecotourism. As such, it provides local communities with incentives to preserve their reefs and to reject other, more destructive use of the area (Shani et al., 2012). For example, in Eilat, Israel, a decision was made, driven in large part by the efforts of local diving organizations, to eliminate nearby profitable fish farms to reduce the potential risk to the coral reefs. (Rinat, 2008). Additionally, local diving organizations may participate in reef awareness projects, the cleaning of reefs, and other environmental projects. The reduction of divers’ negative impacts on the natural reefs is mostly related to diving management strategies. It was established that diving pressure should not exceed 5000-6000 dives per site, per year (Hawkins and Robert, 1997; Schleyer and Tomlin, 2000). These values depend on several factors, like the biological nature of the dive site, the activities pursued, and 233

the level of environmental awareness (Luna et al., 2009). Since diving pressures on the reefs of Eilat greatly exceed this value (Table 7.1; Zakai and Chadwick-Furman, 2002), diving regulations were suggested as a viable way to reduce diving pressures (Zakai and Chadwick- Furman, 2002; Hasler and Ott, 2008). Other preventive measures included improving diver education, especially for diving guides and instructors (Hasler and Ott, 2008), as well as diverting training dives to areas with sandy bottoms and diverting all divers to new artificial reef (AR) sites (Rouphael and Inglis, 1997; Zakai and Chadwick-Furman, 2002; Hasler and Ott, 2008). Dive leaders were found to play an important role in reducing the impact on corals (Barker and Roberts, 2004; Hasler and Ott, 2008; Luna et al., 2009). SCUBA instructors in Eilat usually adhere to rules and regulations as mandated by Israeli diving law, and they are monitored by inspectors from the Israeli diving authority and by rangers from the national parks authority. However, their dive plans in instructor-led dives (except guided dives) are not bound by the rules applied to diving operations; therefore, each instructor is free to choose their own unique dive plan (personal communication with dive instructors). Another way to reduce divers’ impact could be through the zoning of marine protected areas (MPAs) (Riegl and Riegl, 1996) to redirect divers from more sensitive sites, especially at beach diving locations where entry sites situated relatively far from protected areas entail less time spent in those areas. The ability to divert divers from a natural to an artificial reef was examined by Leeworthy et al. (2006) after the immersion of the USS Spiegel Grove in the Florida Keys. Based on information gathered from divers’ logbooks, the authors concluded that the sinking of the Spiegel Grove did indeed decrease the diving pressure on the nearby natural reef. In this study I observed the changes in the diving behavior of SCUBA divers following the insertion of a small, pre-planned, AR with transplanted coral. Specifically, I examined the effects of the introduced AR and of coral transplanting on the diving times in and around a MPA (i.e., marine nature reserve) and visitation time and frequency at adjacent natural coral outcrops. This new data has important implications for the future construction and design of ARs for use by divers, for AR site selection in the proximity of natural reefs and MPAs, and for the management of diving.

232

7.2 Methods

7.2.1 Study Site

An AR was deployed in October 2006 at the northern tip of the Red Sea in the Gulf of Eilat, Israel (29°32’85’’N, 34°57’47’E) on a flat sandy bottom at a depth of 7 m approximately 100 m offshore (Fig. 7.1). It was positioned in the area with the highest density of diving in Israel and where most instructional dives occur (Israeli Diving Federation, personal communication; Fig. 7.1; Table 7.1). The AR was positioned 10 m outside a local MPA and between coral outcrops A and B. Artificial reef deployment and coral transplantation were performed under special permit and with the cooperation of the Israeli Nature and Parks Authority.

Figure 7.1 Map of study site. House icons represent locations of area diving centers. Dotted line represents imaginary boundary of MPA.

231

Table 7.1 Descriptions of dive types and their frequencies in the study area.

Dive Type Instructor Aim Divers’ No. of divers No. of dives presence experience dive-1 (n) in study area (%) Introductory Yes Fun/ None 2 (66) 36720 (16.7) Instruction Refresh Yes Instruction Medium 3.6±1.5 (29) 22054 (10.0) Course Yes Instruction None-low 5.1±1.9 (49) 81458 (37.0) Guided Yes Fun Medium- High 5.6±3.47 (12) 21912 (9.95) Independent No Fun Medium- High 2.4±1 (9) 57972 (26.3)

7.2.2 Artificial Reef design

The AR comprises six concrete units, each weighing 3.5 metric tons (4.2 tons outside the water). The units were lowered to the bottom and connected using 2.5-inch (6.4 cm)-diameter rods, washers, and bolts. Overall AR dimensions are 4×4×4 meters with a 24-m2 surface area and a complexity having a mean rugosity index of 1.19 (SD 0.18) (n = 4 measurements, see McCormick, 1994; Fig. 7.2; Table 7.2). Design of the AR was chosen to allow a high complexity with ample void space and shaded areas, high relief, and free current flow and to minimize sand accumulation on the AR to allow the growth and recruitment of corals and fish. In each module, holes were drilled prior to immersion to serve as anchoring spots for the transplantation of corals. Concurrently to the placement of the AR, stony (Stylophora pisillata, Acropora spp, Pocillopora verrucosa and others) and soft corals (Dendronephthya hemprichi and others) were reared in specially designed coral nurseries (Shafir et al., 2006) located nearby. In April 2007, five months after deployment, 230 stony and 21 soft corals were transplanted onto the AR.

7.2.3 Data collection

All monitored dives were beach dives, in which the divers or dive leaders could choose to enter the water at a location either very close to the MPA (2 m from the MPA fence) or about 40 m north of the MPA to visit the sandy area with scattered coral heads outside the reserve. Dive groups, in both instructor-led and independent dives, were randomly monitored in all coral outcrops and reefs in the area (Fig. 7.1). The groups were followed, from August 2006 to November 2010, by a single snorkeler, and the times spent at distances of 3 m or less from large

230

size (> 21 m2) coral outcrops (i.e., coral-knoll, bommie) within the study site (containing 16 coral outcrops, the fringing reef, and the AR) were recorded.

Figure 7.2 Design of the artificial reef (AR). Openings in the AR were sealed in the interest of safety to prevent divers from entering.

I monitored the dive groups before and after deployment of the AR, immediately following coral transplantation, and two years after AR deployment (to assess diver habituation to the AR). The types of dives monitored were instructor-led dives, such as introductory, refresh, course, or guided dives, or non-instructional (independent) dives (Table 7.1). Instructor-led dives were often restricted to the study area since they are limited by diver air consumption and maximum permitted depth, and they may have been predetermined to take place in sandy, less ecologically sensitive areas. Additionally, group size and entry point to the water were recorded. An imaginary line was drawn perpendicular to the beach and as a continuation of the northern border fence of the MPA marking the boundary between the MPA and the area outside the reserve (Fig. 7.1). Overall times spent in and out of the MPA were also recorded. To assess visitation rate to the AR, a comparison of the number of divers (per half hour) was performed on the similarly sized, natural coral outcrop A. Before the AR was deployed, of the 16 coral outcrops monitored in the diving area, most dive groups visited coral outcrop A (21.4%, n = 41 dives) followed by outcrop B (14.3%), while the combined visitation rates to other outcrops in the area did not exceed 12.2% of the dives. Consequently, I focused my research on outcrops A and B (10 m south and north, respectively, of the AR) and on the AR itself, all three of which have similar properties (Table 7.2). The number of visiting divers was recorded at the AR and at the similarly sized, natural coral outcrop A by counting the number of divers per half hour.

231

Table 7.2 Physical and biological properties of the most visited structures in the study area. No. of Dimensions a Coral cover adult fish Outcrop Type (L×W×H) Volume Rugosity (%)b (Species (m’) c richness) A Natural 5.5×4.3×3.0 71.0 1.63±0.25 52.9±2.7 840 (49) B Natural 5.2×4.2×2.2 48.0 1.29±0.06 36.2±4.4 255 (38) AR Artificial 4.0×4.0×4.0 64.0 1.19±0.18 5.1±0.5 861 (44) a average of four measurements according to McCormick (1994) b average of four line transects (see methods) c unpublished data. Counted in January 2010 when juvenile groups are infrequent.

Coral outcrops A and B and the AR were assessed for coral cover using four 10-m line transects (following Loya, 1972) and four rugosity measurements (following McCormick, 1994). Reef fish abundance and richness were measured monthly on the AR and every three months on outcrops A and B. Fish censuses were 30-minutes long, of which the first two minutes were allocated to counting transient and moving fish. Species and abundance of fish larger than 3 cm were recorded. A sample taken on January 2010, outside of the juvenile recruitment season, was used to compare the structures. Unknown fish were photographed and later identified in the laboratory.

7.2.4 Data Analysis

Dive types, lengths, and rates of visitation were compared using analysis of variance (ANOVA). Detailed analyses were only executed for instructed dives visiting the coral outcrops or entering the MPA. Data that did not conform to the ANOVA assumptions even after transformation were analyzed using the non-parametric tests. A comparison of visitation times inside the MPA was analyzed with a chi-square one-sample test. Analysis was performed on the raw data and not on the percentiles, even if the latter are presented here for the purpose of clarity. A 50% cutoff between short and long entry times to the MPA was made to differentiate between divers who were diving in the vicinity of the AR, thus rendering its presence relevant to their dive, or far from it, entailing a different destination to more distant locations. A comparison of visitation rates on the AR and outcrop A was done using a t-test on data before coral transplantation. When calculating the relationship between the dive entry points and the time spent in the MPA, the data was classified as categorical, since each dive center adopts its own 231

entry point. As such, the regression analysis was categorical and a one-way ANOVA was conducted to observe the differences between the distance categories. Divers that did not reach the MPA at all were excluded from this analysis as the goal of their dive could not be ascertained.

7.3 Results

7.3.1 Reduction of time in MPA

The percent of dive time spent inside the MPA (nature reserve), according to different types of dives, was examined before AR deployment. Unlike many dive locations around the world, independent divers constitute only 26.3% (Table 7.1) of the divers visiting this area compared to in-training divers. Independent divers spent most of their time (89.8±4%; n = 5 dives) inside the nature reserve. These dive groups started their dives directly by entering the MPA and stated that their goal was to visit the inner parts of the nature reserve (as declared by the dive groups before the dive commenced). Data collected from interviews with dive center managers indicated that most dives (63.7%, n = 220116) conducted at my study area were in- training dives comprising introductory and instructed course dives (Table 7.1). Observations of these novice instructor-led dives, like course (n = 13), introductory (n = 14), and refresh dives (n = 6), revealed that the divers spent similar portions of their dives, ranging from 21.8± 21% (Mean± SD) to 38.5%± 30.9 of their total diving times, inside the nature reserve. My examination of all diver classes revealed that overall time spent inside the MPA did not change following AR deployment (Mann-Whitney U test, z = −1.61, p > 0.1). Indeed, the introduction of the AR did not affect the proportion of time spent inside the MPA by experienced independent and guided divers (Mann-Whitney U test, Z = −0.77, p > 0.05 and Z = −1.86, p > 0.05, respectively), but it did cause a shift in the proportion of diving time spent inside the MPA by in-training dives. For in-training dives, there was a decrease in the frequency (percent of time) of short visitations to the MPA (up to 50% of dive time spent in the MPA) from 20-40% of the dive time to a mere 10% (Fig. 7.3; χ2 = 7.82, df = 3, p < 0.05). Dives with longer durations in the MPA did not change (Fig. 7.3; χ2 = 6.63, df = 5, p > 0.05). There was no difference in the percentage of time spent inside the MPA between the period after AR deployment and that after coral transplantation (χ2 = 5.76, df = 3, p > 0.1). 236

50 Before After

(%) Frequency 10

0 10 20 30 40 50 60 70 80 90 100 Diving time in MPA (%)

Figure 7.3 Percent frequency of diving visitations inside the marine protected area (MPA) before (n = 35) and after (n = 30) AR deployment.

The percent of time spent inside the MPA was negatively correlated with the distance from the dive entry point (adjusted r2 = 0.257, β = −0.507, F = 3.63, p < 0.001, n = 108). There was a significant difference in percent of time spent in the MPA between divers entering at the different points (one way ANOVA, log transformed, F = 11.69, p < 0.001). While dive groups entering at 0-10 m from the MPA exhibited the highest visitation time inside the nature reserve (45.5± 30.3%, Mean±SD, n = 48), they were statistically similar to dive groups entering 15-25 m from the MPA (30.5±27.2%, n = 31, Tukey HSD, p < 0.07). On the other hand, both sets of dive groups exhibited significantly higher visitation times in the MPA than divers entering 90-100 m away from the MPA (16.7% ± 13.1, n = 21, Tukey HSD, p < 0.027). Additionally, the number of dives conducted inside the MPA was higher among dive groups entering closer to the MPA (χ2 = 7.82, df = 2, p < 0.025), where 87% (n = 48) of the dive groups entering the MPA entered 0-5 m from the border of the MPA, followed by 77% (n = 31) for the intermediate distance (15- 25 m) and 62% (n = 21) for the longest distance (90-100 m).

231

7. 3.2 Comparison to Adjacent Natural Reefs

A comparison of the AR to the adjacent natural outcrops A and B revealed no difference in the diving times spent around those reefs following AR deployment (Fig. 7.4). A significant difference was found between outcrops A and B (Tukey HSD, p < 0.01), while both were similar to the AR (Tukey HSD, p > 1.0). Outcrop A had a higher mean duration time around it than B, and it differed significantly from outcrop B and the AR (two way ANOVA, df = 2, F = 6.412, p < 0.002) but was independent of the period of examination (two way ANOVA, df = 3, F = 0.91, p > 0.4). This difference between the outcrops was significant (two way ANOVA, df = 2, F = 6.412, p < 0.002) but independent of the period of examination (two way ANOVA, df = 3, F = 0.91, p > 0.4).

600 AR A 500 B

400

300

200

100

Diving time around outcrops (sec) 0 Before After After After deployment deployment transplantation 2 years Period of exmination Figure 7.4 Time spent in the vicinity of the artificial reef (AR) and adjacent outcrops A and B (Mean ± SD) (ANOVA p > 0.05).

Outcrops A and B and the AR are similar in size, but outcrop A has a higher coral cover (mean = 52.9 ±2.7%, 36.2± 4.4%, and 5.1±0.5%, respectively). The richness of the reef fish was similar on all outcrops (mean = 40.6%± 4, 37.2%± 4.1, 38.2% ±6.1 of species, respectively, examined in the second year of monitoring), but fish abundance was higher in the AR than in the others (2570± 1360, 1117 ±691, 986±927, respectively). Visitation rates to the AR and to outcrop A showed no difference between them (t-test, n = 127 dive groups, p > 0.7). Note that the types of dive groups (intro, guided, independent, etc.) varied

231

in their numbers of divers. This is attributed to the different levels of activity per dive type (some dive types visit outcrop A and AR more than others) and to the number of divers typically composing each dive type. (Fig. 7.5; Table 7.1).

16 A 14 AR

-1 12

6 hr of divers No. 4

0 course guided independent introductory

Dive Type

Figure 7.5 Comparison between the numbers of divers hr-1 per dive type on outcrop A and on the AR.

An examination of dive destinations following AR deployment (n = 98 tracked dives) revealed that a similar proportion of the dives went to the AR and to outcrop B (9.2 and 7.2%, respectively), while outcrop A alone drew 20.4% of the dives. But visitation to any two of them or to all three objects drew 21.3% and 24.5%, respectively. Notably, 17.4% of all dive groups did not reach any of the three.

231

7.4 Discussion

This study focused on the effects of SCUBA divers on coral reefs. However, one must bear in mind that these effects tend to be localized to diving spots. Therefore, their impact is limited in comparison to large scale coral reef damaging factors such as coral bleaching or ocean acidification (Bellwood et al., 2004; Hoegh-Guldberg et al., 2007). Similarly, artificial reefs (ARs) and their use for habitat enhancement and conservation are inherently localized. With the rapid deployment of ARs worldwide, however, their interactions with the natural reefs should be considered. The introduction of an AR in close proximity to an MPA and in an area with a high rate of dives did not have a significant effect on the overall number of dives inside the MPA, but it changed the distribution of divers within it. For in-training (instructor-led) dives, AR deployment had a small yet significant effect in reducing the durations of dives inside the boundaries of the MPA (Fig. 7.3). The relatively low magnitude of this change may be ascribed either to the small size of the AR, and hence, its limited attractiveness to the divers, particularly the diving instructors, or to my arbitrary choice of drawing the imaginary boundary line, which is unknown to the divers, of the MPA. Therefore, immersing an AR in such close proximity to the MPA may be successful in diverting divers if the AR is attractive and large enough. On the other hand, locating an attractive AR further away may elicit a change in the diving direction taken by divers and consequently further reduce the time they spend inside the MPA. Qualified divers that spent most of the time inside the MPA before AR deployment did not reallocate their diving times to include sometime around the AR. Potential reasons for their lack of interest in the structure could be the unattractiveness of the AR or its relatively small size, both of which will reduce the time experienced divers will spend at it. In contrast, Leeworthy et al. (2006) found that divers reduced their visitations to natural reef sites as a result of the introduction of a large decommissioned ship, the Spiegel Grove, off the Florida coast. Such ARs are large enough to be suitable dive sites for entire dives, and thus, they serve as viable alternatives to proximate natural reef dives. In general, when given a choice, divers prefer natural coral reefs over artificial ones (Ditton and Baker, 1999; Johns et al., 2001). Indeed, the natural coral reefs of the Red Sea are rated as among the top diving locations in the world (Garrod and Gössling, 2008). It may be that such highly rated natural reefs outdo the nearby AR for the

223

attention of divers. Therefore, whether a diver prefers to explore an AR instead of a nearby natural reef may be related not solely to the attributes of the AR, but also to those of the surrounding natural reefs. An examination of the time spent around the AR compared to the adjacent natural reef outcrops A and B revealed no differences over time (Fig. 7.4). While the AR and outcrop B exhibited similar diver visitation rates, that at outcrop A was higher in all sampled periods. This is possibly due to the higher coral cover on outcrop A (Table 7.2). Notably, the transplantation of corals onto the AR did not increase the visitation time of divers. Additionally, visitation frequencies of divers to the different outcrops next to the AR revealed a preference for outcrop A with a tendency to visit two or three outcrops during the same dive. The preferences of divers for high coral cover and fish richness and abundance were described for natural settings in the Red Sea (Wielgus et al., 2003; Leujak and Ormond, 2007). In the current study, which was conducted on an AR, no such pattern was observed. Moreover, fish richness was similar to all outcrops, but abundance was highest in the AR. I suggest that in this case, fish richness and abundance on the AR are secondary to coral coverage and the overall aesthetics of the structure. Diving instructors are group leaders, and the diving pattern and path of a dive group depends mostly on them. The similar numbers of divers at outcrop A and at the AR and the high numbers of visits to all three outcrops or the avoidance of all of them may indicate that they were all part of the same diving “route.” Conversing with diving instructors (10 open interviews) and observing the routes of their dives (n = 162 followed dives) indicate that each diving operation has a regular “main route” that is usually unique to it and that is used regularly by its instructors. As mentioned by an experienced instructor in one of the leading dive centers in Eilat, “although there is no ‘official’ route that we are mandated to follow, we usually go by the same route” (Y. M. personal communication, translated from Hebrew). Note that I monitored over 25 different dive instructors belonging to eight different dive operations and using different dive plans. The important role of diving guides in reducing diver impact on natural reefs was recorded in St. Lucia in the Caribbean (Barker and Roberts, 2004), in Australia (Davies and Tisdell, 1997), and in the Red Sea (Hasler and Ott, 2008). I suggest that when the location for AR deployment is being chosen that local instructors be consulted to help choose a site that they will subsequently use as a potential destination. (In the current case, the location was chosen by the Israeli Nature and Parks Authority.) Diving instructor education will also improve the use of a distinct diving site and help promote the use of newly introduced ARs. During the study I received several 222

unsolicited letters, in Hebrew, from dive center managers. Their opinions about AR deployment are presented in Table 7.3. In general, the diving community sees the addition of the AR to the local diving environment as a positive development, not only for training and education, but also for conservation and, in short, as an integral part of the diving industry.

Table 7.3 Statements expressed in four unsolicited letters sent by leading dive center managers (A-D) regarding the planned AR. Original letters are in Hebrew.

Topic/ dive club A B C D

AR location + NA NA +

Ability to serve as a − NA − − stand-alone destination

Reduces diving on NR + NA + + Supports local diving + NA NA + industry Influence on + +/− + NA environment Use of AR as an educational/conservation NA + + NA tool Type of most common NA NA Introductory Introductory dive at AR site +/− = positive or negative effect or feature. NA = topic not addressed.

7.5 Conclusions

Managers of natural reefs use a range of tools in their efforts to protect and preserve the environment. Limiting access to the reef may be an effective tool in the short term, but in the long run, it is counterproductive. ARs help redirect divers from natural reefs without reducing their enjoyment of the environment. However, the insertion of an AR into any marine environment with regular diving activity should be examined thoroughly prior to implementation, taking into account divers’ preferences and behavior. In this study, I found that (a) the introduction of a small-scale AR is sufficient to reduce, at least in part, the diving pressure on natural reefs. (b) The AR was effective in changing the behavior of in-training, novice divers but not of advanced divers (independent and guided dives). (c) The transplantation of corals unexpectedly had no effect on diver behavior (although it naturally affected local biodiversity). Notably, the small scale of the AR and the use of a single AR are likely to be 221

among the reasons for this limited effect. Differences in location, measuring methods, scale of natural area, depth of AR, and AR accessibility and scale are important factors influencing diver behavior. To increase the attractiveness of ARs, it is necessary to understand the features that will attract divers to them. Factors like the aesthetic appeal of the structure, its immersion depth, its distance from MPAs and natural reefs, its biotic attributes such as coral and fish biodiversity, and its size should all be taken into consideration in future conservation projects.

220

8. Discussion

Deployment of artificial reefs (ARs) is often done haphazardly and sometimes unintentionally (as in the case of wrecks). But the introduction of new habitats to the marine environment has great implications for the ecology of both fish and corals (and other marine invertebrates) and for its use by man, particularly fisherman and SCUBA divers. The introduction of ARs may benefit each party separately or together and also pose some conflicts between man and environment. Deploying ARs is a multipurpose project covering several disciplines of ecology, social behavior, and economic aspects. Therefore, providing sound guidelines that encompass all disciplines is challenging. Ecological knowledge that can increase the proper use of pre-planned ARs is needed (Miller, 2002), especially as natural reefs are being degraded and the need for successful conservation tools that incorporate man’s interaction (e.g., eco-tourism), is in demand. Creation of new habitat results in an increase in biodiversity as demonstrated by the species-area concept (Belmaker et al., 2007a). Therefore the simple introduction of ARs, even without any manipulation or pre-designed consideration, is expected to increase local biodiversity (Baine, 2001), and is in agreement with results presented in this work. Introduction of ARs to the natural environment will also increase the number of available dive sites appealing to divers. Deployment of new dive sites may either elicit increased diving to the area (Adams et al., 2009) or divert divers from natural to artificial sites (Leeworthy et al., 2006; chapter 7). However, simply dumping objects into the water is ineffective and should be better tailored to increase attractiveness to fish, corals, and divers by pre-planning and designing the AR deployment. Several factors, such as structure type, location, coral transplantation, and the dynamics of time on ARs can enhance AR attractiveness, making it an improved conservation tool and enhancing fish and other fauna for commercial use as well. The structure of an AR has great influence on the fish and coral community. The AR structure creates new microhabitats that can support increased abundance and richness of fish (Bohnsack, 1989, Burt et al., 2009). The difference in micro-habitats between natural and artificial reefs creates distinct fish communities and may support specific species (Clark and Edwards, 1994, Abelson, 2006; Edelist and Spanier, 2009). ARs may have unique features that are lacking in the natural reef (NR), like void space (Sherman et al., 2002), substrate type

221

(Harriot and Fisk, 1987), shaded areas (Kojis, 1957; Bohnsack and Sutherland, 1985), and vertical relief (Rilov and Benayahu, 1998; Perkol-Finkel and Benayahu, 2004). I found that fish composition was different from that on natural reefs, as seen elsewhere (Arena et al., 2007; Burt et al., 2009), and the main contributors to this difference were shade-loving nocturnal fish as was found in the Mediterranean Sea (Spanier et al., 1990). This result conformed to the first hypothesis in chapter 4 stating that time since deployment will maintain a difference in fish communities between the AR and the natural reef. AR structure serves as a new niche for coral recruitment as well. The use of artificial substrate is expected to limit recruits to species that are more adapted to AR substrates, such as Pocilloporids (Perkol-Finkel and Benayahu, 2007) and soft coral Nephthyiids (Benayahu and Loya, 1984, 1987; Perkol-Finkel and Benayahu, 2007). ARs may serve as good substrate for algae growth and, in turn, as observed in this study, attract herbivorous fish and other herbivores (sea urchins). The resultant differences in structure between ARs and NRs dictate differences in the biological attributes of the ARs, but it does not necessarily apply to recreational divers that appear to like ARs and NRs similarly (Spieler et al., 2001; Shani et al., 2012; but see Ramos et al., 2006 and Sutton and Bushnell, 2007 for divers that prefer NRs). Instructor-led divers visit the AR deployed in the study site similarly as they visit the natural reef (Chapter 7) therefore dismissing the prediction that ARs placed in close proximity to a natural reef will reduce the dive time on the natural reef. However, it appears that certified divers preferred larger objects where they can spend a significant time of their dive. Divers were interviewed in a parallel study (Shani et al., 2012) and were found to prefer large vessel ships and submerged airplanes as most favourite. Divers, in that study, claimed to visit AR dive sites similarly to natural sites and showed a preference for themed structures that have resemblance to known objects rather than abstract objects. It is not surprising that the deployed AR comprising small sized trigonometric figure with no added themed features, did not elicit change in the diving behaviour of experienced divers. Complexity is another attribute of structure that is known to increase fish and coral biodiversity (Spanier et al., 1990; Abelson and Shlesinger, 2002; Gratwicke and Speight, 2005; Sherman et al., 2002; Okamoto et al., 2008) and can increase the attractiveness for divers. Increased complexity increased the number of microhabitats and ultimately increased fish and coral biodiversity. Indeed, increase in coral transplantation caused an increase in complexity and an increase in fish biodiversity. Additionally, the introduction of holes of several sizes created 221

shelter niches for species-specific fish (Spanier et al., 1990) and other invertebrates (Spanier, 1994). The micro-complexity of the AR substrate is shown to enable coral recruitment. Increased complexity also affects divers’ preference of the dive site. I believe that because the experimental ARs’ complexity attributes were designed especially for fish and corals rather than for divers, it rendered the structure too abstract and reduced its human attractiveness. Location of deployment can have considerable consequences on the AR attractiveness. ARs serve as new islands in the sea and thus their fish community is subject to different ecological processes dictated by the proximity of the AR to the NR (as predicted by the Island Biogeography model; MacArthur and Wilson, 1967), the isolation of the object (Ault and Johnson, 1998; Nanami and Nishihara, 2003; Belmaker et al., 2011), the contrast between the object and the environment (Tews et al., 2004), as well as internal processes such as competition, predation, recruitment, and immigration (Hixon and Menge, 1991; Mora et al., 2003). I observed that relocation of ARs from a location close to an NR to an area with sand and sea grass beds increased richness and abundance and maintained differences between AR and natural reef fish community composition. These results conform to the second hypothesis suggested, that predicted that fish richness, abundance, biodiversity and community structure will be higher in more isolated reef patches. Location can influence the availability of coral recruits and coral survivorship (Abelson, 2006). Shemla (2002) has found that small ARs, even in close distances, have different recruitment and coral growth rates, hence the success of ARs as coral attractant is questionable. Location of AR deployment can influence the visitation of divers as well. Divers prefer easily accessible dive sites (“not too close and not too far”; Shani et al., 2012). The deployment location should take into account the depth, the distance from other attractions (natural reefs or other ARs), and the regular diving routes of divers. The AR deployed in my study site was located in the route of instructor-led dives, and therefore was not distinct enough to serve as independent dive site and thereby reduce its effectiveness to divert divers from nearby natural reefs. Hence, the hypothesis declaring that ARs in heavily dived area will reduce the diving pressure off natural reefs, was not accepted. Transplantation of corals increases complexity and habitat availability and therefore potentially increases fish and invertebrate biodiversity. My study showed that coral transplantation increased the abundance and richness of fish but did not cause changes in the fish community structure, therefore partially agreeing with hypothesis that coral transplantation on the ARs will influence the richness and abundance of reef fishes on the ARs. Coral 226

transplantations on the AR enhanced fish richness mainly of coral-associated fish. Additionally, the configuration of the transplantation can have other important effects. Coral-associated fish were found to increase their activity levels in more complex habitats and also in corals less than 50 cm apart (Chapter 3). These results agree with the hypotheses of the study that declared fish movement to be negatively correlated with the distance of adjacent corals and that habitat complexity between coral patches influences the number of times the fish move between coral patches. It was suggested that this behavioral change is caused by predation risk avoidance and increase in shelter availability. Therefore the orientation of the corals transplanted can affect the distribution and behavior of coral-associated fish. As corals are being degraded worldwide (Feary et al.,2007), this effect may hinder the development of the fish inhabiting them and those in neighbouring corals. Transplantation of corals was found to be dependent on corals species, where coral morphology, death and self-attachment were correlated with varied species. This agrees with the initial hypothesis that species of the coral transplants will influence its overall survival and the coral’s ability to self-attach to the AR. The selection different coral species, based both on the coral morphology and its survival ability is critical for the successful transplantation of corals on ARs, especially if they are located in heavily dived areas. Transplantation of corals may also change the aesthetic appearance of ARs and make them more attractive to divers. Using an economical approach I tested different situations where divers rated images of highly diverse habitats, including high coral and fish ARs, as more attractive (Chapter 6). This result conforms with the study hypothesis that biodiversity of the coral reef community (e.g., or biological scenario) will influence divers willingness to pay for AR. As corals may increase fish abundance it is important to try and transplant corals on ARs to achieve enhancement in AR attractiveness to both fish and divers. That said, in my experimental AR, transplantation of corals did not elicit a change in the behavior of divers and failed to divert them from natural to artificial settings. This is in contrast to the initial prediction that coral transplants on ARs will decrease the dive time off nearby natural reefs. Perhaps the transplantations were not sufficient to cause an aesthetic difference (coral cover was still very low compared to natural reefs) and the transplantation of very few species was not diverse enough. Alternatively, it is possible that the relatively small size of the AR had more influence than the transplantations and the whole AR was not big enough to create a change in diving behavior regardless of coral transplantation. In addition, transplantation in heavily dived (and with abundant novice divers) area caused dislodgement of transplanted corals resulting in their demise. Addition of the 221

protective measures did not result in increase coral survivorship contrary to the prediction that protected areas will reduce coral dislodgement. It was noted that divers, and particularly underwater photographers, were able to penetrate between the protective divers deterring pegs. It is known that photographers cause extensive damage to corals (Rouphael and Inglis, 2001). This conflict between adding corals and the destruction of these corals by divers needs to be resolved through improved transplantation techniques and maybe in combination with diving management measures and education. Transplantation of corals can be a successful tool for increasing AR attractiveness, but should be used cautiously. In areas where natural coral recruitment is prevalent this investment in transplantation may span just a few years until naturally recruited corals will establish on the AR. However, in areas where low natural recruitment occurs, the investment in the attachment of transplanted corals is longer and its cost effectiveness should be taken into account. Recurring coral transplantations and protective measures of transplanted corals may be costly, although coral nurseries can significantly reduce these costs. Future advancement in coral transplantation techniques will improve this conservation tool. Colonization of fish to ARs is time dependent as well as associated with distance to the nearby natural reefs and local ecological processes such as competition and predation. By intensively (monthly) monitoring the fish composition on the AR over a prolonged period of time, I found that (1) fish composition remained distinct between AR and natural reef outcrops, (2) composition on the AR became more similar as time progressed, and (3) the major determinant for the change in composition was the first major fish recruitment event (Chapter 4). My results correlated with the difference in fish communities between the AR and the natural reef will be maintained over time, since the specific features of the AR influences the fish composition of the AR and coral transplantation on the AR will influence the richness and abundance of reef fishes on the AR. Interestingly, I found that community maturation followed a stepwise pattern rather than a continuous model. I did not find any similar documentation elsewhere although recruitment-dependent colonization have been documented as structuring fish community composition (Doherty, 1983). Coral transplantation can reduce the amount of time for corals to colonize and pass over the initial phases of recruitment. Seasonality of coral reproduction will determine the availability of new coral recruits to ARs, and timing of deployment should consider this for reducing the amount of time for the development of corals. Indeed, coral recruitment was found to follow a seasonality pattern on the experimental AR. Transplanted corals are also prone to seasonality, as faster growth rate is correlated with 221

seasonality and water temperature (Klein and Loya, 1991). Time affected transplanted corals negatively, as a result of SCUBA divers’ damage. As transplanted corals’ time for attachment is species dependent (Chapter 5), it is suggested that enclosing some areas of transplanted corals from divers’ direct touch could increase their rate of attachment and survival. Divers association with ARs could also depend on time of the AR in the water. I did not notice a change in the diving behavior of divers over time, but divers have responded in interviews (Chapter 6) that ARs with increased natural attributes such as corals and fish, are more preferable, thus making them more attractive. It is assumed that an increase in the fouling fauna will ultimately increase divers’ behavior on the ARs. The immersion of ARs for recreational purposes is a complicated effort that has ecological and socio-economic implications. Most AR deployments usually aim at just one of these aspects and disregarded the other. ARs for recreational purpose put emphasis on the socio- economic aspects, and submersion of ARs for conservation and reef restoration unusually give weight to ecological aspects. But since the two aspects are not mutually exclusive it is imperative that a sensible and sustainable approach be undertaken in both cases. In this work I found some complementary aspects between ecologic and socio-economic approaches. The addition of corals onto ARs increased fish biodiversity and increased AR attractiveness. Also, AR structure provides new habitats that can be used both for coral and fish and visited by divers. To the contrary, some aspects turned out to be conflicting. Divers had a negative impact on coral transplants. Coral transplantation was effective for increasing fish biodiversity but these corals were easily dislodged by divers and died. The main goals of this work were to produce knowledge on how to introduce an AR to divert divers off nearby natural reefs and to create a novel habitat suitable for fish and corals. I used coral transplantation to enhance both goals. Results showed that the submerged AR only partly diverted the divers and that the transplantation of corals did not induce change in diver’s behavior. Additionally. the AR did not provide a safe location for increased coral survival but was very successful for fish aggregation and production. Other work related to other ARs nearby suggested the potential for using ARs with corals transplants exists. It is, therefore predicted that the incompatibility between man and nature can be reduced substantially through habitat design, diving management and educational measures, and better transplantation techniques. At the current state of declining reefs, active measures should be taken to relieve pressure off natural reefs. These should include introducing new habitats that can serve both man and nature, thus reaching a win-win situation that will 221

support sustainability of nature without reducing the benefits nature can offer mankind (Rosenzweig, 2003).

213

Appendix

Table A.1 The location of an artificial reef possibly has an effect on the fish community structure that was formerly related to distance from the natural reefs, isolation, and the heterogeneity of the local habitat. In my study I monitored the relocation of four artificial reefs from a location next to natural reef to a new location in the proximity of sand flats and sea grass beds. Table A.1 lists the fish species with their corresponding locations in the study and their biological traits. Section 2.3.1 elaborates on this experiment.

Species Locations Activity Feeding Life-history Level Guild stage (Adlt./Juv.) Acanthopagrus bifasciatus OAR,OTAR T Zoobenthos A Acanthurus nigrofuscus NAR T Herbivore J Amblyeleotris steinitzi NR C Zoobenthos A Amblyglyphidodon flavilatus NAR S Planktivore J Amblygobius albimaculatus SG S Zoobenthos A Amphiprion bicinctus NAR,NR,OTAR S Planktivore A Anampses lineatus OTAR S Zoobenthos A Anthias taeniatus OAR,NAR S Planktivore A Apogon annularis OAR C Planktivore A Apogon aureus OTAR C Zoobenthos A Apogon cyanosoma OAR,NAR,OTAR C Planktivore A, J Apogon pseudotaeniatus NAR,OTAR S Planktivore A Apogon sp. NAR S Planktivore A Apolemichthys xanthoides OAR,NAR T Zoobenthos A Arothron stellatus NAR T Zoobenthos A Balistes fuscus OTAR T Zoobenthos A Bodianus anthiodes OAR,NAR T Zoobenthos A, J Canthigaster coronata OAR,NR S Zoobenthos A Canthigaster margaritatta NAR S Zoobenthos A Cephalopholis hemistictus NR T Piscivore A Chaetodon fasciatus OAR T Corallivore A Chaetodon paucifasciatus OAR,NAR,NR T Corallivore A Cheiliditerus quinquelinetus NAR S Zoobenthos A Cheilinus mentalis NAR T Zoobenthos A Cheilodipterus macrodon NAR,OTAR,OAR S Piscivore A Cheilodipterus quinquelineatus OTAR,OAR C Zoobenthos A Cirihlabrus rubriventralis NAR,OTAR S Planktivore A Corythoichthys schultzi NR,OTAR C Zoobenthos A Cyclichthys spilostilus OAR,NAR T Zoobenthos A Dascyllus marginatus OAR,NAR,NR S Planktivore A Dascyllus trimaculatus OAR,NAR,OTAR S Planktivore A 212

Dendrochirus brachypterus OTAR S Zoobenthos A Diagrama pictum OAR T Piscivore A Ecsenius frontalis OTAR S Herbivore A Ecsenius gravieri OAR,NAR C Herbivore A Epinephelus aeroltus NAR,OTAR T Piscivore A Epinephelus fasciatus OAR,NAR,NR,OTAR T Piscivore A Epinephelus fasciatus NAR T Piscivore A, J Fistularia commersonni OTAR T Piscivore A Gnatholepis anjerensis OTAR C Zoobenthos A Gymnothorax griseus OTAR C Piscivore A Gymnothorax nudivomer OAR C Piscivore A Heniochius difreutes NAR,OTAR T Planktivore A Heniochus intermedius OAR,NAR,OTAR T Planktivore A Istigobius decoratus OAR,NR,OTAR C Zoobenthos A Labrid sp. NAR,SG S Zoobenthos A Labroides dimidiatus OAR S Zoobenthos A Labroides dimidiatus NAR,OTAR,NR S Zoobenthos A Lethrinus borbonicus NAR T Piscivore A Meiacanthus nigrolineatus OAR,NAR,OTAR S Zoobenthos A Mulloidichthis flavolineatus OAR,NAR T Zoobenthos A Myripristis murdjan OAR,NAR C Planktivore A Neopomacentrus miriyae OAR,NAR,OTAR T Planktivore A Ostracion cubicus OAR,NAR,NR,OTAR T Zoobenthos A, J Ostracion cyanurus NAR T Zoobenthos A Oxycheilinus mentalis NR,OTAR T Zoobenthos A Paracheilinus octotaenia NAR,NR S Planktivore A Parapriacanthus ransonneti OAR S Planktivore J Parupeneus forsskali NAR,OTAR T Zoobenthos A Parupeneus macronema OAR,NAR,NR T Zoobenthos A Parupeneus rubescens NAR T Zoobenthos A Pastinachus sephen OTAR T Zoobenthos A Plagiotremus townsendi NAR S Zoobenthos A Platycephalus NAR T Zoobenthos A Pomacentrus trichorus OAR,NAR,NR,OTAR S Planktivore A Psedocheilinus hexataenia OAR,NAR,OTAR S Zoobenthos A Pseudoanthias squamipinnis OAR,NAR,OTAR S Planktivore A Pseudocheilinus evanidus NR S Zoobenthos A Pseudochromis flavivertex OAR,NR S Zoobenthos A Pseudochromis fridmani NAR S Zoobenthos A Pseudochromis springeri NAR,NR S Zoobenthos A Pterocaesio chrysozona OTAR T Planktivore A Pterois miles OAR,NAR,OTAR T Piscivore A Sargocentron diadema OAR,NAR,OTAR C Zoobenthos A

211

Saurida gracilis NAR C Piscivore A Scarus sp. NAR,OTAR T Herbivore A Scarus sp. 2 NAR,OTAR T Herbivore A Scarus sp. 3 NAR T Herbivore A Scolopsis ghanam OAR,NAR,OTAR S Zoobenthos A Scorpaenopsus diabolus OAR C Piscivore A Siganus argentus NAR,OTAR T Herbivore A Siganus rivulatus NAR T Herbivore A Stethojulis albovittata OTAR S Zoobenthos J Sufflamen albicaudata OAR,NAR T Zoobenthos A Synodus variegatus OAR,NAR,OTAR S Piscivore A Thalassoma rueppellii NAR S Zoobenthos A Thallassoma lunare OTAR S Zoobenthos A Variola louti NR T Piscivore A Locations: OAR – Old Artificial Reef, NRA – New Artificial reef, NR – Natural Reef, OTAR – Other Artificial Reefs, SG – Sea Grass. Activity levels: T–Transient, S–Sedentary, C–Cryptic.

210

Table A.2 Home range in Dascyllus marginatus was previously thought to be constrained to a few meters (<3 m) from its obligatory coral. But observations in the field have shown that the number of fish in a single coral is constantly changing. Therefore I investigated the mobility of marked D. marginatus in four coral plots. Table A.2 presents the home range attributes of all marked D. marginatus fish. Further information can be found in the main text section 3.3.3.

Plot Fish Fish Size Sighting Distance to Max. distance Change in Percent ID (cm) incidents closest coral travelled (cm) location Change (cm)

1 A 5.5 32 18 0 0.0 1 B 5.6 32 18 0 0.0 1 C 5.5 32 18 10 1 3.2 1 D 5.7 31 18 1 3.2 1 E 5.7 30 300 1 3.2 1 F 5.2 28 140 3 9.7 1 G 6.7 29 18 3 9.7 1 H 4.5 29 18 3 9.7 1 I 4.6 29 18 3 9.7 1 J 5.6 25 18 3 9.7 1 K 5.4 27 18 3 9.7 1 L 5.7 25 18 5 16.1 1 M 4.7 25 81 120 5 16.1 1 N 5.5 27 81 270 6 19.4 1 O 5.5 29 18 572 6 19.4 1 P 4.4 26 85 140 10 32.3

2 Q 5.6 29 75 1 3.2 2 R 5.2 30 175 2 6.5 2 S 5.6 30 10 125 6 19.4 2 T 26 10 70 6 19.4 2 U 5.9 31 10 58 6 19.4 2 V 5 31 35 35 7 22.6 2 W 5.5 25 35 125 8 25.8 2 X 5.5 31 35 275 9 29.0

3 Y 5.9 31 23 0 0.0 3 Z 5.5 30 52 1 3.2 3 AB 4.5 28 35 3 9.7 3 AC 5.9 27 35 4 12.9 3 AD 4.5 25 35 5 16.1 3 AE 4.6 26 23 111 9 29.0

4 AF 4.7 31 53 0 0.0 4 AG 5.1 32 53 0 0.0 4 AH 5.1 32 53 0 0.0 4 AI 5.2 32 32 0 0.0 4 AJ 5.3 32 32 0 0.0 4 AK 5 31 50 1 3.2 211

4 AL 4 31 32 1 3.2 4 AM 6.3 30 50 2 6.5 4 AN 5 32 50 50 3 9.7 4 AO NA 28 32 57 7 22.6

211

Table A.3 Artificial reefs interact with the environment in which they are found. I investigated the effect of the deployment of a planned artificial reef on the local environment. I inspected the resemblance of the AR to similar sized coral outcrops and also how the fish community resemblance changed over time. Table A.3 presents the results of ANOASIM and SIMPER analysis of the most common fish species on the AR and natural reef outcrops A and B. See section 4.3 in the main text for more details.

Species Avg. abundance Avg. similarity % contribution

A. AR vs. NR comparison

Natural Reef

Average similarity 60.56 Pseudoanthias squamipinnis 3.92 6.51 10.75 Dascyllus marginatus 2.58 4.43 7.31 Chromis viridis 2.38 3.83 6.32 Neopomacentrus miryae 2.64 3.00 4.95 Pseudochromis fridmani 1.61 2.75 6.62

Artificial Reef

Average similarity 53.33 Pseudoanthias squamipinnis 4.00 6.33 11.88 Neopomacentrus miryae 3.55 4.50 8.44 Thalassoma kluzingeri 2.05 3.10 5.82 Acanthurus nigrofuscus 1.5 2.37 4.45 Dascyllus marginatus 1.94 2.25 4.21

B. AR comparison over time

Before Transplantation

Average similarity 49.11 Pseudoanthias squamipinnis 3.28 15.26 31.08 Bodianus anthioides 1.06 1.39 10.56 Acantharus nigrofuscus 1.22 1.78 10.04 Parupeneus forskali 1.10 1.25 7.26 Zebrasoma xanthurum 0.86 1.19 6.29

< 17 months

Average similarity 58.89 Pseudoanthias squamipinnis 4.02 7.58 12.87 Neopomacentrus miryae 2.71 4.47 7.59 216

Thalassoma kluzingeri 1.97 3.64 6.18 Dascyllus marginatus 2.03 3.13 5.31 Acanthurus nigrofuscus 1.41 2.74 4.66

> 17 months

Average similarity 63.72 Pseudoanthias squamipinnis 3.99 5.31 8.34

Apogon cyanosoma 3.87 4.38 6.87 Neopomacentrus miryae 3.92 4.18 6.55 Dascyllus marginatus 2.36 3.07 4.83 Archamia fucata 2.82 2.76 4.33

211

Table A.4 Coral transplantation on artificial reefs is an uncommon practice and the procedure and the benefits of such restoration process are still not known. In my main study site in Eilat’s reefs, I transplanted corals on an experimental artificial reef. While the main results of the experiments are described in section 5.3, in Table A.4 I list the number of corals transplanted on the artificial reef at flat (vertical and horizontal) and curved surfaces that were not monitored regularly during the study.

1st trans- 2nd Trans- 3rd Transplantation plantation plantation Area on artificial reef Flat Flat Curved Flat Total Acropora spp. 39 40 45 10 55 Pocilopora verrucosa 21 17 15 14 29 Stylophora pistillata 16 22 69 70 139 Favia favus 3 6 5 3 8 Dendronephthya hemprichii 21 13 9 0 9 Acabaria spp. 1 0 0 0 0 Porites lutea 4 13 7 13 20 Millepora dichotoma 18 19 1 12 13 Cyphastrea serallia 1 1 1 1 2 Cyphastrea chalcidicum 2 1 1 1 2 Platygyra daedalea 1 2 1 14 15 Plerogyra sinuosa 1 1 1 1 2 Lithophyton arboreum 0 1 1 1 2 Goniopora spp. 2 2 0 0 0 Goniastrea spp. 4 4 Total 130 138 156 144 300

211

Supplementary material A.5

The preference for biological attributes of natural reefs was formerly studied in social and economic papers. Nonetheless, very few studies have investigated the relationship between marine biodiversity and preferences of SCUBA divers. In Chapter 7 I examined this relationship via a contingent valuation method for the willingness to pay for an artificial reef at different hypothetical biological states. Supplementary material 5 presents the survey delivered to 306 participants on the beaches of Eilat.

Survey (English translation; Survey was distributed in Hebrew)

Coral reefs possess a large biodiversity of fish and corals. Coral reef conditions across the world and in Eilat in particular, are in decline. The coral reef in Eilat, a one-of-a-kind natural feature of Israel that contains unique local coral and fish species, is frequently damaged as a result of a variety of different factors. Consequently, the reef is expected to continue to decline, and some researchers anticipate that it may become extinct within 20–30 years.

One way to conserve the state of the reef is to deploy artificial reefs. These artificial reefs will be deployed in the vicinity of the natural reef, and corals will be transplanted and maintained on them. These corals will serve as habitats for fish and invertebrates and will be specially grown by professionals and researchers. Artificial reef deployment is expected to provide both new habitats for fish and corals and new dive sites designed for divers.

All activities connected with deployment of alternate reefs, growing of the corals, and subsequent maintenance of these reefs incurs heavy expenditures. Therefore, donations will be needed to cover these expenses. The money collected by an NGO will be earmarked specifically for the care and maintenance of the artificial reefs.

The Israel Nature and Parks Authority, the Israel Ministry of Tourism, and the Israel Government Tourist Corporation are interested in learning whether and to what extent the public benefits from the installation of such alternative artificial reefs and whether the project is worth the investment.

In this questionnaire you will be presented with a number of choices of artificial reefs with different biological attributes and you will be asked to give the monetary value of each of these choices. In every scenario you will be presented with the reef alternatives comprising four biological states (for a total of 7 scenarios). For each picture presented, you will be asked to declare how much you would be willing to donate to visit the reef depicted in a particular biological state.

When making your choices, remember the financial costs that will be incurred by yourself and anybody who accompanies you in spending your vacation and diving in Eilat. These costs include accommodation,

211

food, drinks, other recreational pastimes, and also the cost of renting diving/snorkelling gear. In addition, remember that there are other environmental and social causes that are equally as important.

In each scenario you will be presented with four pictures of reef alternatives differentiated by their biological states:

A. A “naked” reef, with no restoration. B. A reef after low conservation efforts. C. A reef after medium conservation efforts. D. A reef after high conservation efforts. If an NGO will be established whose only goal will be to maintain and nurture surrogate ARs, and all the money collected will be dedicated to this use only, what is the maximum value, in NIS, that you will be willing to pay as a contribution on a yearly basis, for restoring the surrogate ARs to the following proposed state?

A. To maintain the current state/ reef without restoration, I am willing to pay on a yearly basis: 0 NIS 5 NIS 10 NIS 15 NIS 20 NIS 25 NIS 30 NIS 35 NIS 40 NIS 45 NIS 50NIS 55 NIS 65 NIS 70 NIS 75 NIS 80 NIS 85 NIS 90 NIS 95 NIS 100NIS 105NIS 110NIS 115NIS 120NIS 125NIS 130NIS 135NIS 140NIS 145NIS 150NIS Other______I don’t have enough information to decide______

B. To improve the current state to an initial biological state, I am willing to pay on a yearly basis: 0 NIS 5 NIS 10 NIS 15 NIS 20 NIS 25 NIS 30 NIS 35 NIS 40 NIS 45 NIS 50NIS 55 NIS 65 NIS 70 NIS 75 NIS 80 NIS 85 NIS 90 NIS 95 NIS 100NIS 105NIS 110NIS 115NIS 120NIS 125NIS 130NIS 135NIS 140NIS 145NIS 150NIS Other______I don’t have enough information to decide______

C. To improve the current state to a medium biological state, I am willing to pay on a yearly basis: 0 NIS 5 NIS 10 NIS 15 NIS 20 NIS 25 NIS 30 NIS 35 NIS 40 NIS 45 NIS 50NIS 55 NIS 65 NIS 70 NIS 75 NIS 80 NIS 85 NIS 90 NIS 95 NIS 100NIS 105NIS 110NIS 115NIS 120NIS 125NIS 130NIS 135NIS 140NIS 145NIS 150NIS Other______I don’t have enough information to decide______

D. To improve the current state to a high biological state, I am willing to pay on a yearly basis: 0 NIS 5 NIS 10 NIS 15 NIS 20 NIS 25 NIS 30 NIS 35 NIS 40 NIS 45 NIS 50NIS 55 NIS 65 NIS 70 NIS 75 NIS 80 NIS 85 NIS 90 NIS 95 NIS 100NIS 105NIS 110NIS 115NIS 120NIS 125NIS 130NIS 135NIS 140NIS 145NIS 150NIS Other______I don’t have enough information to decide______

203

Now we ask you to choose between reasons that best explain your choice of the values you rated. You can mark one reason or more. If you choose more than one reason then circle it and write next to the reason its order of importance, where 1 is “most important,” etc.

A. I empathize with the goals of conservation, and it is important for me to ensure the existence of the fish and corals in Eilat. B. Reef conservation is not important enough to me to spend money on it. C. I am willing to pay for reef conservation to ensure that my children and their children have the opportunity to view fish and corals in the future. D. The values I placed are appropriate for this goal. E. It is not my duty to finance conservation efforts directed at the reefs in Eilat. F. I would like to keep to myself the option to view the reefs in the future although I do not do this currently. G. I regularly visit the reef of Eilat, and I want to maintain the coral reef and its biodiversity. H. I prefer visiting natural reefs. In conclusion, a few questions about you (for statistical purposes only):

1. How many years have you been a certified diver? ______2. What is your level of certification? A. Open Water B. Advanced Open Water C. Divemaster D. Assistant Instructor or higher

3. How many dives have you done in the last year?______4. Place of residence: 1. Local 2. Domestic 3. Outside Israel 5. Gender? 1. Male 2. Female 6. Age ______7. Marital status: 1. Single 2. Married 3. Other 8. Number of children:______9. Occupation: Student/Soldier/Part time/Full time/Independent/Pensioner/Unemployed 10. Education: Elementary/High school/Technical/Academic 11. Is your income lower than/equal to/ higher than 8,000 NIS gross per month (the average income)? 12. Do you belong to an environmental organization? If so, which one______13. Frequency of visitation to Eilat ______per month/year. 14. Is there anything else you would like to share with us?______

202

Bibliography

Abelson, A. 2006. Artificial reefs vs. coral transplantation as restoration tools for mitigating coral reef deterioration: Benefits, concerns, and proposed guidelines. Bulletin of Marine Science, 78(1): 151-159.

Abelson, A., Shlesinger, Y. 2002. Comparison of the development of coral and fish communities on rock-aggregated artificial reefs in Eilat, Red Sea. Journal of Marine Science, 59: 122-126.

Abramovitz, J. 1991. Biodiversity: Inheritance from the past, investment in the future. Environmental Science and Technology, 25: 1817−1818.

Adams, C., Lindberg, B., Stevely, J. The Economic benefits associated with Florida’s artificial reefs. A publication of the Food and Resource Economics Department, Florida Cooperative Extension Service, Institute of Food and Agricultural Sciences, University of Florida, http://edis.ifas.ufl.edu, Accessed 04.06.2013

Addicott, J.F., Aho, J.M., Antolin M.F., Padilla, D.K., Richardson, J.S., Soluk, D.A. 1987. Ecological neighborhoods: scaling environmental patterns. Oikos, 49: 340-346.

Ahmed, M., Umali, G.M., Chong, C.K., Rull, M.F., and Garcia, M.C. 2007. Valuing recreational and conservation benefits of coral reefs – The case of Bolinao, Philippines. Ocean and Coastal Management, 50: 103−118.

Alevizon, W.S, Gorham, J.C. 1989. Effects of artificial reef deployment on nearby resident fishes. Bulletin of Marine Science, 44: 646-661.

Almany, G.R. 2004a. Does increased habitat complexity reduce predation and competition in coral reef fish assemblages? Oikos, 106(2): 275-284.

Almany, G.R. 2004b. Differential effects of habitat complexity, predators and competitors on the abundance of juvenile and adult coral reef fishes. Oecologia, 141: 105- 113.

201

Almany, G.R., Webster, M.S. 2006. The predation gauntlet: early post-settlement mortality in reef fishes. Coral Reef, 25: 19-22.

Ambrose, R.F., Sawbrick, S.L. 1989. Comparison of fish assemblages on artificial and natural reefs off the coast of southern California. Bulletin of Marine Science, 44(2): 718- 733. Angel, D.L., Spanier, E., 2002. An application of artificial reefs to reduce organic enrichment caused by net-cage fish farming: preliminary results. Journal of Marine Science, 59: 324-329.

Arena, P.T., Jordan, L.K.B, Spieler, R.E. 2007. Fish assemblages on sunken vessels and natural reefs in southeast Florida, USA. Hydrobiologia, 580: 157−171.

Auberson, B. 1982. Coral transplantation: an approach to the reestablishment of damaged reefs. Kalikasan, Philippines Journal of Biology, 11(1): 158-172.

Ault, T.R., Johnson, C.R. 1998. Spatially and temporally predictable fish communities on coral reefs. Ecological Monographs, 68(1): 25-50.

Baine, M. 2001. Artificial reefs: a review of their design, application, management and performance. Ocean and Coastal Management, 44: 241-251.

Baird, A.H., Hughes, T.P. 2000. Competitive dominance by tabular corals: an experimental analysis of recruitment and survival of understorey assemblages. Journal of Experimental Marine Biology and Ecology, 251: 117–132.

Baird, A.H., Salih, A., Trevor-Jones, A. 2006. Fluorescence census techniques for the early detection of coral recruits. Coral Reefs, 25(1): 73-76.

Barker, N.H.L., Roberts, C.M. 2004. SCUBA diver behaviour and the management of diving impacts on coral reefs. Biological Conservation, 120: 481-189.

Bateman, I.J, Willis, K.J. 1999. Valuing Environmental Preferences: theory and practice of the contingent valuation method in the US, EU, and developing countries. Oxford University Press, Oxford. 645 pp.

200

Bell, J.D. Galzin, R. 1984. The influence of live coral cover on coral-reef fish communities. Marine Ecology Progress Series, 15: 266-274.

Bell, A.M., Stamps, J.A. 2004. Development of behavioral differences between individuals and populations of sticklebacks, Gasterosteus aculeatus. Animal Behaviour, 68: 1339-1348.

Bellwood, D.R., Hughes, T.P., Folke, C., Nystrom, M. 2004. Confronting the coral reef crisis. Nature, 429: 827-833.

Belmaker, J., Shashar, N., Ziv, Y. 2005. Effects of small-scale isolation and predation on fish diversity on experimental reefs. Marine Ecology Progress Series, 289: 273–283.

Belmaker, J., Ben-Moshe, N., Ziv, Y., Shashar, N. 2007a. Determinants of the steep species–area relationship of coral reef fishes. Coral Reefs, 26: 103-112.

Belmaker, J., Polak, O., Shashar, S., Ziv, Y. 2007b. Geographic divergence in the relationship between Paragobiodon echinocephalus and its obligate coral host. Journal of Fish Biology, 71: 1555-1561.

Belmaker, J. 2009. Species richness of resident and transient coral-dwelling fish responds differentially to regional diversity. Global Ecology and Biogeography, 18(4): 426- 436.

Belmaker, J., Ziv, Y., Shashar, S. 2009. Habitat patchiness and predation modify the distribution of a coral-dwelling fish. Marine Biology, 156: 447-454.

Belmaker, J., Ziv, Y., Shashar, N. 2011. The influence of connectivity on richness and temporal variation in reef fishes. Landscape Ecology, 26: 587-597.

Benayahu, Y., Loya, Y. 1984. Substratum preferences and planulae settling of two Red Sea alcyonaceans: Xenia macrospiculata Gohar and Parerythropodium fulvum fulvum (Forskål). Journal of Experimental Marine Biology and Ecology, 83: 249–261.

201

Benayahu, Y., Loya, Y. 1987. Long-term recruitment of soft-corals (Octocorallia: Alcyonacea) on artificial substrata at Eilat (Red Sea). Marine Ecology Progress Series, 38: 161–167.

Ben-Moshe, N. 2007. Habitat selection of the coral dwelling fish Dascyllus marginatus – effects of corals and “roommates”. M.Sc. thesis, Faculty of Natural Sciences, Ben Gurion University of the Negev. 65 pp.

Ben-Tuvia, A., Diamant, A., Baranes, A., Golani, D. 1983. Analysis of a coral reef fish community in shallow waters of Nuweiba, Gulf of Aqaba, Red Sea Bulletin of the Institute of Oceanography and Fisheries, 9: 193-205.

Ben-Tzvi, O., Abelson, A., Polak, O., Kiflawi, M. 2008. Habitat selection and the colonization of new territories by Cromis viridis. Journal of Fish Biology, 73: 1005-1018.

Beukers, J.S., Jones, G.P. 1997. Habitat complexity modifies the impact of piscivores on a coral reef fish population. Oecologia, 114: 50-59.

Bohnsack, J.A. 1983. Species turnover and the order versus chaos controversy concerning reef fish community structure. Coral Reefs, 1: 223-228.

Bohnsack, J.A. 1989. Are high densities of fish at artificial reefs the result of habitat limitation or behavioral preference? Bulletin of Marine Science, 44 (2): 631-645.

Bohnsack, J.A., Sutherland, D.L. 1985. Artificial reef research: A review with recommendations for future priorities. Bulletin of Marine Science. 37(1): 11-39.

Bohnsack, J.A., Harper, D.E., McClellan, D.B., Hulsbeck, M. 1995. Effects of Reef Size on Colonization and Assemblage Structure of Fishes at Artificial Reefs Off Southeastern Florida, U.S.A. Bulletin of Marine Science, 55(2-3): 796-823.

Bortone, S.A., 2006. A perspective of artificial reef research: The past, present, and future. Bulletin of Marine Science, 78(1): 1-8.

Bowden-Kerby, A. 2001. Low-tech coral reef restoration methods modeled after natural transplantation processes. Bulletin of Marine Science, 69: 915-931. 201

Brander, L.M., Van Beukering, P., Cesar, H.S.J. 2007. The recreational value of coral reefs: A meta-analysis. Ecological Economics, 63: 209−218.

Brickhill, M.J., Lee, S.Y., Connolly, R.M. 2005. Fishes associated with artificial reefs: attributing changes to attraction or production using novel approaches. Journal of Fish Biology, 67: 52-71

Brock, V.E. 1954. A preliminary report on a method for assessing reef fish populations. Journal of Wildlife Management, 18:297–308.

Brokovich, E. 2001. The community structure and biodiversity of the reef-fishes of the northern Gulf of Eilat and their relation to their habitat. M.Sc. thesis. Faculty of Life Sciences. Tel-Aviv University Tel-Aviv.

Brokovich, E., Baranes, A., Goren, M. 2006. Habitat structure determines coral reef fish assemblages at the northern tip of the Red Sea. Ecological Indicators, 6: 494–507.

Bruckner, A.W., Bruckner, R.J. 2001. Condition of restored Acropora palmata fragments off Mona Island, Puerto Rico, 2 years after the Fortuna Reefer ship grounding. Coral Reefs, 20(3): 235-243.

Burt, J., Bartholomew, A., Usseglio, G., Bauman, A., Sale, P.F. 2009. Are artificial reefs surrogates of natural habitats for corals and fish in Dubai, United Arab Emirates? Coral Reefs, 28: 663-675.

Cabaitan, P.C., Gomez, E.D., Alino, P.M. 2008. Effects of coral transplantation and giant clam restocking on the structure of ﬁsh communities on degraded patch reefs. Journal of Experimental Marine Biology and Ecology, 357: 85–98.

Carr, M.H., Hixon, M.A. 1997. Artificial reefs: The importance of comparisons with natural reefs. Fisheries, 22(4): 24-33.

Cater, C., Cater, E. 2001. Marine environments, pp. 265-282. In Weaver D.B. (Ed.), The encyclopedia of ecotourism. CAB International, Wallingford, UK. 688 pp.

206

Cesar, H.S.J. 2000. Coral reefs: Their functions, threats and economic value, pp.14– 39. In Cesar H.S.J. (Ed.) Essays on the economics of coral reefs. Coral Reef Degradation in the Indian Ocean Program. Kalmar University, Kalmar, Sweden. 243 pp.

Cesar, H.S.J., Van Beukering, P.J. 2004. Economic Valuation of the Coral Reefs of Hawai‘i. Pacific Science, 58: 231–242.

Chabanet, P., Ralambondrainy, H., Amanieu, M., Faure, G., Galzin, R. 1997. Relationships between coral reef substrata and fish. Coral Reefs, 16: 93-102.

Chapman, M.R., Kramer, D.L. 2000. Movements of fishes within and among fringing coral reefs in Barbados. Environmental Biology of Fishes, 57: 11-24.

Chesson, P. 1998. Recruitment limitation: A theoretical perspective. Australian Journal of Ecology 23: 234-240.

Clark, S., Edwards, A.J. 1994. Use of artificial reef structures to rehabilitate reef flats degraded by coral mining in the Maldives. Bulletin of Marine Science, 55(2-3): 724-744.

Clark, S., Edwards, A.J. 1995. Coral transplantation as an aid to reef rehabilitation: Evaluation of a case study in the Maldive Islands. Coral Reefs, 14: 201–213.

Clarke, K.R. 1993. Non-parametric multivariate analyses of changes in community structure. Australian Journal of Ecology, 18: 117-143.

Cocheret de la Morinèire, E., Nagelkerken, I., Meij, H., Velde, G. 2004. What attracts juvenile coral reef fish to mangroves: habitat complexity or shade? Marine Biology, 114(1): 136-145. Colwell, R.K. 2006. EstimateS: Statistical estimation of species richness and shared species from samples. Version 8. Persistent URL

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

201

Croker, D.J., Graham, N.A.J., Pratchett, M.S. 2012. Interactive effects of live coral and structural complexity on the recruitment of reef fishes. Coral Reefs, 31(4): 919-927.

Crosby, M.P., Reese, E. S. 2005. Relationship of habitat stability and intra-specific population dynamics of an obligate corallivore butterflyfish, Aquatic Conservation: Marine and Freshwater Ecosystems, 15: 13–25.

Cummings, S.L. 1994. Colonization of a nearshore artificial reef at Boca Raton (Palm Beach County), Florida. Bulletin of Marine Science, 55: 1193-1215.

Curtin, S. 2010. Wildlife tourism: the intangible, psychological benefits of human– wildlife encounters. Current Issues in Tourism, 12: 451−474.

Davenport, J., Davenport, J.L., 2006. The impact of tourism and personal leisure transport on coastal environments: a review. Estuarine Coastal and Shelf Science, 67: 280– 292.

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

Ditton, R.B., Baker, T.L. 1999. Demographics, Attitudes, Management Preferences, and Economic Impacts of Sport Divers using Artificial Reefs in Offshore Texas Waters. Report prepared for the Texas Parks and Wildlife Department through a research contract with Texas A and M University.

Dixon, J.A., Scura, L.F., Vanthof V. 1993. Meeting ecological and economical goals- marine parks in the Caribbean. Ambio, 22: 117−125.

Doherty, P.J. 1983. Tropical territorial damselfish: is density limited by aggression of recruitment? Ecology, 61: 176-190.

Doherty, P., Fowler, T. 1994. An empirical test of recruitment limitation in a coral reef fish. Science, 263: 935-939.

Dubinsky, Z., Stambler, N. 2011. Coral reefs: an ecosystem in transition. Springer. New York. 552 pp. 201

Ebersole, J.P. 2001. Recovery of fish assemblages from ship groundings on coral reefs in the Florida Keys National Marine Sanctuary. Bulletin of Marine Science, 69(2): 655- 571. Edelist, D. Spanier, E. 2009. Influence of Levantine Artificial Reefs on the fish assemblage of the surrounding seabed. Mediterranean Marine Science, 10(1): 35-54.

Edwards, A.J., Clark, S. 1998. Coral transplantation: a useful management tool or misguided meddling? Marine Pollution Bulletin, 37: 8-12.

Edwards, A.J., Gomez, E. 2007. (Eds.) Reef Restoration Concepts and Guidelines: Making sensible management choices in the face of uncertainty. Coral Reef Targeted Research and Capacity Building for Management Program. St. Lucia, Australia, 38 pp.

Einbinder, S., Perelberg, A., Ben-Shaprut, O., Foucart, M.H., Shashar, N. 2006. Effects of artificial reefs on fish grazing in their vicinity: Evidence from algae presentation experiments. Marine Environmental Research, 61: 110-119.

Epstein, N., Bak, R.P.M., Rinkevich, B. 1999. Implementation of a small-scale ‘‘no- use zone’’ policy in a reef ecosystem: Eilat’s reef-lagoon six years later. Coral Reefs, 18: 327-332.

Epstein, N., Bak, R.P.M., Rinkevich, B. 2001.Strategies for gardening denuded coral reef areas: the applicability of using different types of coral material for reef restoration. Restoration Ecology, 9: 432-442.

Feary, DA. 2007. The influence of resource specialization on the response of reef fish to coral disturbance. Marine Biology, 153: 153–161.

Feary, D.A., Almany, G.R., McCormick, M.I., Jones, G.F. 2007. Habitat choice, recruitment and the response of coral reef Fishes to coral degradation. Oecologia, 153:727– 737.

Fernandes, L., Day, J., Lewis, A., Sledgers, S., Kerrigan, B., Breen, D., Cameron, D., Jago, B., Hall, J.,Lowe, D., Innes., J., Tanzer, J., Chadwick, V., Thompson, L., Gorman, K., Simmons, M., Barnett, B., Sampson, K., De’ath, G., Mapstone, B., Marsh, H., Possingham, 201

H., Ball, I., Ward., T., Dobbs, K., Aumend, J., Slater, D., Stapleton, K. 2005. Establishing Representative No-Take Areas in the Great Barrier Reef: Large-Scale Implementation of Theory on Marine Protected Areas. Conservation Biology, 19(6): 1733-1744.

Ferse, S.C.A. 2008. Artificial reef structures and coral transplantation: fish community responses and effect on coral recruitment in North Sulawesi/ Indonesia. PhD dissertation, Center for Tropical Marine Ecology, Bremen. 169 pp.

Fitzhardinge, R.C., Bailey-Brock, J.H. 1989. Colonization of artificial reef materials by corals and other sessile organisms. Bulletin of Marine Science, 44: 567-579.

Folke, C., Carpenter S., Walker, B., Scheffer, M., Elmqvist, T., Gunderson, L., Holling, C.S. 2004. Regime shifts, resilience and biodiversity in ecosystem management. Annual Review of Ecology Evolution and Systematics, 35: 557−581.

Folpp, H, Lowry, M., Gregson, M., Suthers, I.M. 2011.Colonization and community development of fish assemblages associated with estuarine artificial reefs. Brazilian Journal of Oceanography, 59(1): 55-67.

Fox, H.E., Pet, J.S., Dahuri, R. 2001. Enhanced Coral Reef Recovery After Destructive Fishing Practices: Initial Results in Komodo National Park. Indonesian Journal of Coastal and Marine Resources, 3: 36-44.

Freeman, M. 2003. The measurement of environmental and resource values: Theory and methods, 2nd ed. RFF press, Washington DC, USA. 491 pp.

Fricke, H.W. 1980. Control of different mating systems in a coral reef fish by one environment factor. Animal Behavior, 28: 561-569.

Frieler, K., Meinshausen, M., Golly, A., Mengel, M., Lebek, K., Donner, S.D., Hoegh-Guldberg, O. 2013. Limiting global warming to 2 °C is unlikely to save most coral reefs. Nature Climate Change, 3: 165-170.

213

Gabrie, C., Porcher, M., Masson, M. 1985. Dredging in French Polynesian coral reefs: towards a general policy of resource exploitation and site development. Proceedings of the 5th International Coral Reef Congress, 4: 271-277.

Garrod, B., Gössling, S. 2008. Introduction. pp. 8-29. In Garrod, B. and Gössling, S., (Eds.), New frontiers in Marine Tourism: Diving experiences, Sustainability, Management. Elsevier Press, Amsterdam, Netherlands, 226 pp.

Gladfelter, W.B., Ogden, J.C., Gladfelter, E.H. 1980. Similarity and diversity among coral reef fish communities: a comparison between tropical western Atlantic (Virgin Islands) and tropical central pacific (Marshall Islands) patch reefs. Ecology, 61: 1156-1168.

Glassom, D., Zakai, D., Chadwck-Furman, N.E. 2004. Coral recruitment: a spatio- temporal analysis along the coastline of Eilat, northern Red Sea. Marine Biology, 144: 641– 651.

Golani, D., Diamant, A. 1999. Fish colonization of an artificial reef in the Gulf of Eilat, northern Red Sea. Environmental Biology of Fishes, 54, 275-282.

Golani, D., Lerner, A. 2007. A long-term study of the sandy shore ichthyofauna in the Northern Red Sea (Gulf of Aqaba) with reference to adjacent mariculture activity. Raffles Bulletin of Zoology, 14: 255-264.

Gratwicke, B., Speight, M.R. 2005. Effects of habitat complexity on Caribbean marine fish assemblages. Marine Ecology Progress Series, 292: 301-310.

Greenfield, D.W., Johnson, R.K. 1990. Heterogeneity in habitat choice in cardinalfish community structure. Copeia, 1990: 1107–1114.

Grober-Dunsmore, R., Frazer, T.K., Beets, J.P., Lindberg, W.J., Zwick, P., Funicelli, N.A. 2008. Influence of landscape structure on reef fish assemblages. Landscape Ecology, 23: 37-53.

Guest, J.R., Dizon, R.M., Edwards, A.J., Franco, C., Gomez, E.D. 2011. How quickly do fragments of coral “self-attach” after transplantation? Restoration Ecology, 19(2): 234- 242. 212

Hanley, N. 2000. Contingent valuation as a means of valuing the conservation of coral reefs: an assessment of the method, pp. 241−246. In Integrated coastal zone management of coral reefs: Decision support modeling. Gustavson, K., Huber, R.M., Ruitenbeek, J. (Eds.). The World Bank, Washington DC, USA. 292 pp.

Harcourt, J.L., Sweatman, G., Johnstone, R.A., Manica, A. 2009. Personality counts: the effect of boldness on shoal choice in three-spined sticklebacks. Animal Behaviour, 77: 1501-1505.

Harriott, V.J., Fisk, D.A. 1987. A comparison of settlement plate types for experiments on the recruitment of scleractinian corals. Marine Ecology Progress Series, 37: 201-208.

Harriott, V.J., Fisk, D.A. 1988. Coral Transplantation as a reef management option. Proceedings of the 6th International Coral Reef Symposium 2: 375-379.

Harriott, V. J., Davis, D., Banks, S.A., 1997. Recreational diving and its impact in Marine Protected Areas in Eastern Australia. Ambio, 26: 173-179.

Hasler, H., Ott, J.A. 2008. Diving down the reefs? Intensive diving tourism threatens the reef of the northern Red Sea. Marine Pollution Bulletin, 26: 1788-1794.

Hawkins, J.P., Roberts, C.M. 1997. Estimating the carrying capacity of coral reefs for SCUBA diving. Proceedings of the 8th International Coral Reef Symposium, Panama City, Panama. 2: 1923–1926.

Hawkins, J.P., Roberts, C.M., Van’t Hof, T., Meyer, K.D., Tratalos, J., Aldam, C. 1999. Effects of Recreational Scuba Diving on Caribbean Coral and Fish Communities. Conservation Biology, 13(4): 888-897.

Hayun, M., Alfia, L. 2004-2008. Tourism to Israel – Statistical report. Ministry of Tourism, Israel.

211

Herrera, R., Espino, F., Garrido, M., Haroun, R.J. 2002. Observations on fish colonization and predation on two artificial reefs in the Canary Islands. Journal of Marine Science, 59: 69-73.

Hess, R., Rushworth, D., Hynes, M., Peters, J. 2001. Disposal Options for Ships. National Defense Research Institute RAND. 59-80.

Hilbertz, W.H., Fletcher, D., Krausse, C. 1977. Mineral accretion technology: applications for architecture and aquaculture. Industrial Forum, 8: 75-84.

Hixon, M.A. 1991. Predation as a process structuring coral reef fish communities, pp. 475-508. In Sale, P.F. (Ed.) The Ecology of Fishes on Coral Reefs. Academic Press, San Diego and London. 754 pp.

Hixon, M.A. 1998. Population dynamics of coral-reef fishes: Controversial concepts and hypotheses. Australian Journal of Ecology, 23: 192-201.

Hixon, M.A., Beets, J.P. 1993. Predation, prey refuges, and the structure of coral-reef fish assemblages. Ecological Monographs, 63: 77-101.

Hixon, M.A., Menge, B.A., 1991. Species diversity: prey refuges modify the interactive effects of predation and competition. Theoretical Population Biology, 39: 178- 200.

Horoszowski-Fridman, Y.B., Izhaki, I., Rinkevich, B. 2011. Engineering of coral reef larval supply through transplantation of nursery-farmed gravid colonies. Journal of Experimental Marine Biology and Ecology, 399: 162-166.

Hoegh-Guldberg, O. 1999. Climate change, coral bleaching and the future of the world’s coral reefs. Marine and Freshwater Research, 8: 839–866.

Hoegh-Guldberg, O., Mumby, P.J., Hooten, A.J., Steneck, R.S., Greenfield, P., Gomez, E., Harvell, C.D., Sale, P.F., Edwards, A.J., Caldeira, K., Knowlton, N., Eakin, C.M., Iglesias-Prieto, R., Muthiga, N., Bradbury, R.H., Dubi, A., Hatziolos, M.E. 2007. Coral reefs under rapid climate change and ocean acidification. Science, 318: 1737–1742.

210

Holbrook, S.J., Brooks, A.J., Schmitt, R.J. 2002. Predictability of fish assemblages on coral patch reefs. Marine and Freshwater Research, 53: 181-188.

Holbrook, S.J., Schmitt, R.J., Brooks, A.J. 2008. Resistance and resilience of a coral reef fish community to changes in coral cover. Marine Ecology Progress Series, 371: 263– 271.

Holland, K.N., C.G. Lowe, Wetherbee, B.M. 1996. Movements and dispersal patterns of blue trevally (Caranx melampygus) in a fisheries conservation zone. Fish Research, 25: 279–292.

Hudson, J.H., Goodwin, W.B. 1997. Restoration and growth rate of hurricane pillar coral (Dendrogyra cylindricus) in the Key Largo National Marine Sanctuary, Florida. Proceedings of the 8th International Coral Reef Symposium 1: 567-570.

Hudson, J.H., Goodwin, W.B. 2001. Assessment of vessel grounding injury to coral reef and seagrass habitats in the Florida Keys National Marine Sanctuary, Florida: protocols and methods. Bulletin of Marine Science, 69(2): 509-516.

Hughes, T.P., Baird, A.H., Bellwood, D.R., Card, M., Connolly, S.R., Folke, C., Grosberg, R., Hoegh-Guldberg, O., Jackson, J.B.C., Kleypas, J., Lough, J.M., Marshall, P., Nystroem, M., Palumbi, S.R., Pandolfi, W.J.M., Rosen, B., Roughgarden, J. 2003. Climate change, human impacts, and the resilience of coral reefs. Science, 301: 929–933.

Hunter, W.R., Sayer, M.D.J. 2009. The comparative effects of habitat complexity on faunal assemblages of northern temperate artificial and natural reefs. Journal of Marine Science, 66: 691–698.

Jaap, W.C. 2000. Coral reef restoration. Ecological Engineering, 15(3-4): 345-364.

Jensen, A.C. 1997. European Artificial Reef Research. (Ed.) Proceedings of the first EARRN conference. Ancona, Italy. Southampton Oceanography Centre, Southampton, UK. 90 pp.

211

Johns, G., Leeworthy, V.R., Bell, F., Bonn, M., 2001. Socioeconomic Study of Reefs in Southeast Florida. Final report. Florida Fish and Wildlife Conservation Commission and National Oceanic and Atmospheric Administration, Fort Lauderdale, Florida.

Jones, G.P., Syms, C. 1998. Disturbance, habitat structure and the ecology of fishes on coral reefs. Australian Journal of Ecology, 23: 287–297.

Jones, G.P., McCormick, M.I., Srinivasan, M., Eagle, J.V. 2004. Coral decline threatens fish biodiversity in marine reserves. Proceedings of the National Academy of Science, 101: 8257-8253.

Jordan, L.K.B., Gilliam, D.S., Spieler, R.E. 2005. Reef fish assemblage structure affected by small scale spacing and size variations of artificial reef patches. Journal of Experimental Biology and Ecology, 326: 170-186.

Kojima, S. 1957. Reaction of fish to a shade or floating substances. Bulletin of Japanese Society of Scientific Fisheries, 22: 730-735.

Kotler, B.P., Brown, J.S., Mitchell, W.A. 1994. The role of predation in shaping the behaviour, morphology and community organization of desert rodents. Australian Journal of Zoology, 42(4): 449-466.

Klein, R., Loya, Y. 1991. Skeletal growth and density patterns of two Porites corals from the Gulf of Eilat, Red Sea. Marine Ecology Progress Series, 77: 253-259.

Leeworthy, V.R., Maher, T., Stone, E.A. 2006. Can Artificial Reefs alter user pressure on adjacent natural reefs? Bulletin of Marine Science, 78(1): 29−37.

Leitão, F., Santos, M.N., Erzini, K., Monteiro, C.C. 2008a. The effect of predation on artificial reef juvenile demersal fish species. Marine Biology, 153: 1233-1244.

Leitão, F., Santos, M.N., Erzini, K., Monteiro, C.C. 2008b. Fish assemblages and rapid colonization after enlargement of an artificial reef off the Algarve coast (Southern Portugal). Marine Ecology, 29: 435-448.

211

Leujak, W., Ormond, R.F.G. 2007. Visitor Perceptions and the Shifting Social Carrying Capacity of South Sinai’s Coral Reefs. Environmental Management, 39: 472−489.

Lima, S.L., Dill, L.M. 1990. Behavioral decisions made under the risk of predation: a review and prospectus. Canadian Journal of Zoology, 68: 619-640.

Lipton, D.W., Wellman, K., Sheifer, I.C., Weiner R.F. 1995. Economic valuation of natural resources: A handbook for coastal resource policymakers. NOAA Coastal Ocean Program Decision Analysis Series No.5. NOAA Coastal Ocean Office, Silverspring MD. 131 pp.

Loya, Y. 1972. Community structure and species diversity of hermatypic corals at Eilat, Red Sea. Marine Biology, 13: 100-123.

Luckhurst, B.E., Luckhurst, K. 1978. Analysis of substrate variables on coral reef fish communities. Marine Biology, 49: 317-323.

Luna, B., Pérez, C.V., Sánchez-Lizaso J.L. 2009. Benthic impacts of recreational divers in a Mediterranean Marine Protected Area. Journal of Marine Science, 66: 517-523.

MacArthur, R.H., Pianka, E.R. 1966. On the optimal use of a patchy environment. American Naturalist, 100(916): 603-609.

MacArthur, R.H., Wilson, E.O. 1967. The theory of island biogeography. Princeton University Press, Princeton, New Jersey, USA. 203 pp.

Manatunge, J., Asaeda, T., Priadarshana, T. 2000. The influence of structural complexity on fish–zooplankton interactions: a study using artificial submerged macrophytes. Environmental Biology of Fishes, 58: 425–438,

Mann, C.C., Plummer, M.L. 1993. High cost of biodiversity. Science, 260: 1868−1871.

Martín-López, B., Montes, C., Benayas, J. 2007. Economical valuation of biodiversity conservation: the meaning of numbers. Conservation Biology, 22: 624−635.

216

McAfee, S.T., Morgan, S.G. 1996. Resource use by five sympatric parrotfishes in the San Blas Archipelago, Panama. Marine Biology, 125: 427–437.

McCormick, M.I. 1994. Comparison of field methods for measuring surface topography and their associations with a tropical reef fish assemblage. Marine Ecology Progress Series, 112: 87-96.

Medio, D., Ormond, R.F.G., Pearson, M. 1996. Effect of briefing on rates of damage to corals by SCUBA divers. Biological Conservation, 79: 91-95.

Mellin, C., Huchery, C., Caley, M.J., Meekan, M.G., Bradshaw, C.J.A. 2010. Reef size and isolation determine the temporal stability of coral reef fish populations. Ecology, 91(11): 3138-3145.

Miller, M.W., Barimo, J. 2001. Assessment of juvenile coral populations at two reef restoration sites in the Florida Keys National Marine Sanctuary: Indicators of success? Bulletin of Marine Science, 69: 395–405.

Miller, M.W, Weil, E., Szmant, A.M. 2000. Coral recruitment and juvenile mortality as structuring factor for reef benthic communities in Biscayne National Park, USA. Coral Reefs, 19: 115-123.

Miller, M.W. 2002. Using ecological processes to advance artificial reef goals. Journal of Marine Science, 59: 27-31.

Moberg, F., Folke, C. 1999. Ecological goods and services of coral reef ecosystems. Ecological Economics, 29: 215-233.

Moore, M., Erdmann, M. 2002. EcoReefs: a new tool for coral reef restoration. Conservation in Practice, 3: 41-44.

Mora, C., Chittaro, P.M., Sale, P.F., Kritzer, J.P., Ludsin, L.A. 2003. Patterns and processes in reef fish diversity. Nature, 421, 933-936.

211

Munday, P.L. 2000. Interactions between habitat use and patterns of abundance in coral dwelling fishes of the genus Gobiodon. Environmental Biology of Fishes, 58(4): 355- 369.

Munday, P.L., Wilson, S.K. 1997. Comparative efficacy of clove oil and other chemicals in anaesthetization of Pomacentrus amboinensis, a coral reef fish. Journal of Fish Biology, 51: 931-938.

Muñoz-Chagín, R.F. 1997. Coral transplantation program in the Paraiso coral reef, Cozumel Island, Mexico. Proceedings of the 8th International Coral Reef Symposium 2: 2075-2078.

Myhre, L.C., Forsgren, E., Amundsen, T. 2012. Effects of habitat complexity on mating behavior and mating success in a marine fish. Behavioral Ecology, 24(2): 553-563.

Nanami, A., Nishihira, A. 2001. Survival rates of juvenile coral reef fishes differ between patchy and continuous habitats. Bulletin of Marine Science, 69: 1209–1221.

Nanami, A., Nishihira, M. 2003. Population dynamics and spatial distribution of coral reef fishes: comparison between continuous and isolated habitats. Environmental biology of Fishes, 68: 101-112.

Nassauer, J.I., Corry, R.C. 2004. Using normative scenarios in landscape ecology. Landscape Ecology, 19: 343−356.

Newman, H., Chuan, C.S. 1994. Transplanting a reef: a Singapore community project. Coastal Management in Tropical Asia, 3: 11-14.

Newman, D.J., Kilama, J., Bernstein, A., Chivian, E. 2008. Medicines from nature, pp.117-162. In Sustaining Life: How Human Health Depends on Biodiversity. Chivian, E., Bernstein, A. (Eds.). Oxford University Press, New York. 542 pp.

Nomakuchi, S., Park, P.J., Bell, M.A. 2009. Correlation between exploration activity and use of social information in three-spined sticklebacks. Behavioural Ecology, 20(2): 340- 345.

211

Nunes, P.A.L.D., and van den Bergh, J.C.J.M. 2001. Economic valuation of biodiversity: Sense or nonsense? Ecological Economics, 39: 203−222.

Öhman, M.C., Rajasuriya, A., 1998. Relationships between habitat structure and fish communities on coral and sandstone reefs. Environmental Biology of Fishes, 53(1): 19-31.

Okamoto, M., Nojima, S., Fujiwara, S., Furushima, Y. 2008. Development of ceramic settlement devices for coral reef restoration using in situ sexual reproduction of corals. Fisheries Science, 74: 1245–1253.

Oren, U. Benayahu, Y. 1997. Transplantation of juvenile corals: a new approach for enhancing colonization of artificial reefs. Marine Biology, 127: 499-505.

Ortiz-Prosper, A.L., Bowden-Kerby, A., Ruiz, H., Tirado, O., Caban, A., Sanchez, G., Crespo, J.C. 2001. Planting small massive corals on small artificial concrete reefs or dead coral heads. Bulletin of Marine Science, 69: 1047-1051.

Overholtzer-McLeod, K.L. 2006. Consequences of patch reef spacing for density- dependent mortality of coral-reef fishes. Ecology, 87(4): 1017-1026.

PADI Professional Association of Diving Instructors. 2010. PADI Worldwide Certification History. Available from http://www.padi.com/padi/en/footerlinks/certhistorygraph.aspx

Pandolﬁ, J.M., Bradbury, R.H., Sala, E., Hughes, T.P., Bjorndal, K.A., Cooke, R.G., McArdle, D., McClenahan, L., Newman, M.J.H., Paredes, G., Warner, R.R., Jackson, J.B.C., 2003. Global Trajectories of the Long-Term Decline of Coral Reef Ecosystems. Science, 301: 955-958.

Pendleton, L.H. 2004. Creating underwater value: the economic value of artificial reefs for recreational diving. The San Diego Oceans Foundation.

Perkol-Finkel, S., Benayahu, Y. 2004. Community structure of stony and soft corals on vertical unplanned artificial reefs in Eilat (Red Sea): comparison to natural reefs. Coral Reefs, 23: 195−205.

211

Perkol-Finkel, S., Shashar, N., Benayahu, Y. 2006. Floating and fixed artificial habitats: effects of substratum motion on benthic communities in a coral reef environment. Marine Ecology Progress Series, 317: 9-20.

Perkol-Finkel, S., Benayahu, Y. 2007. Differential recruitment of corals onto artificial and natural reefs. Journal of Experimental Marine Biology and Ecology, 340: 25–39.

Peters, H., Hawkins, J.P. 2008. Access to marine parks: A comparative study in willingness to pay. Ocean and Coastal Management, 52: 219−228.

Peterson, G., Allen, C.R., Holling, C.S. 1998. Ecological resilience, biodiversity, and scale. Ecology, 1: 6−18.

Pickering, H., Whitmarsh, D. 1997. Artificial reefs and fisheries exploitation: a review of the ‘attraction versus production’ debate, the influence of design and its significance for policy. Fisheries Research, 31: 39−59.

Pickering, H., Whitmarsh, D., Jensen, A. 1998. Artificial reefs as a tool to aid rehabilitation of coastal ecosystems: Investigating the potential. Marine Pollution Bulletin, 37: 505-514.

Piniak, G.A., Fogarty, N.D., Addison, C.M., Kenworthy, W.J. 2005. Fluorescence census techniques for coral recruits. Coral Reefs, 24: 496-500.

Pitt, D.G., Nassauer, J.I. 1992. Virtual reality systems and research on perception, simulation and presentation of environmental change. Landscape and Urban Planning, 21: 269−271.

Plucer-Rosario, G.P., Randall, R.H. 1987. Preservation of rare coral species by transplantation: an examination of their recruitment and growth. Bulletin of Marine Science, 41: 585-593.

Polak, O., Shashar, N. 2012. Can a small artificial reef reduce diving pressure from a natural coral reef? Lessons learned from Eilat, Red Sea. Ocean and Coastal Management, 55: 94-100.

213

Priskin, J. 2003. Tourist perceptions of degradation caused by coastal Nature- based recreation. Environmental Management, 32(2): 189-204.

Ramos, J., Santos, M.G., Whitmarsh, D., Monteiro, C.C. 2006. The usefulness of the analytic hierarchy process for understanding reef diving choices: A case study. Bulletin of Marine Science, 78(1): 213-219.

Raymundo, L.J. 2001. Mediation of growth by conspecific neighbors and the effects of site in transplanted fragments of the coral Porites attenuata Nemenzo in the central Philippines. Coral Reefs, 20, 263-272.

Raymundo, L.J.H., Maypa, A.P., Luchavez, M.M. 1999. Coral seeding as a technology for recovering degraded coral reefs in the Philippines. Phuket Marine Biology Center Special Publication, 20: 81-92.

ReefBall Foundation. 2012. http://www.reefball.org (08/2012).

Richmond, R.H. 1993. Coral reefs: present problems and future concerns resulting from anthropogenic disturbance. American Zoologist, 33: 524–536.

Rickel, S., Genin, A. 2005. Twilight transitions in coral reef fish: the input of light- induced changes in foraging behavior. Animal Behaviour, 70: 133-144.

Riegl, B., Riegl, A. 1996. Studies on coral community structure and damage as a basis for zoning marine reserves. Biological Conservation, 77: 269-277.

Rilov, G., Benayahu, Y. 1998. Vertical artificial structures as an alternative habitat for coral reef fishes in disturbed environments. Marine Environmental Research, 45(4/5): 431-451.

Rilov, G., Benayahu, Y. 2002. Rehabilitation of coral reef-fish communities: The importance of artificial-reef relief to recruitment rates. Bulletin of Marine Science, 70(1): 185-197

212

Rilov, G., Figueria, W.F., Lyman, S.J., Crowder, L.B. 2007. Complex habitats may not always benefit prey: linking visual field with reef fish behaviour and distribution. Marine Ecology Progress Series, 329: 220-238.

Rinat, Z., 2008. Fish cages finally leaving Eilat. Haaretz-Front 11.06.2008. Available from: http://www.haaretz.com/print-edition/news/fish-cages-finally-leaving-eilat-1.247554. Accessed 22.07.2011.

Rinkevich, B. 1995. Restoration strategies for coral reefs damages by recreational activities: the use of sexual and asexual recruits. Restoration Ecology, 3: 241-251.

Rinkevich, B. 2005. Conservation of coral reefs through active restoration measures: Recent approaches and last decade progress. Environmental Science and Technology, 39: 4333-4324.

Rinkevich, B. 2008. Management of coral reefs: we have gone wrong when neglecting active reef restoration. Marine Pollution Bulletin, 56: 1821−1824.

Roberts, C.M., Ormond, R.FG. 1987. Habitat complexity and coral reef fish diversity and abundance on Red Sea fringing reefs. Marine Ecology Progress Series, 41: 1-8.

Rooker, J.R., Dokken, Q.R., Pattengill, C.V., Holt, G.J. 1997. Fish assemblages on artificial and natural reefs in the Flower Garden Banks national marine sanctuary. Coral Reefs, 16: 83-92.

Roopesh, J., Shailendra, S., Noopur M. 2008. Marine organisms: Potential Source for Drug Discovery. Current Science, 94: 292.

Rosenfeld, D., Woodley, W.L., Lerner, A., Kelman, G., Lindsey, D.T. 2008. Satellite detection of severe convective storms by their retrieved vertical profiles of cloud particle effective radius and thermodynamic phase. Journal of Geophysical Research - Atmospheres, 113: D04208, 22 pp.

Rosenzweig, M.L. 2003. Reconciliation ecology and the future of species diversity. Oryx, 37(2): 194-205.

211

Ross, P.M., Thrush, S.F., Montgomery, J.C., Walker, J.W., Parsons, D.M. 2007. Habitat complexity and predation risk determine juvenile snapper (Pagrus auratus) and goatfish (Upeneichthys lineatus) behaviour and distribution. Marine and Freshwater Research, 58: 1144–1151.

Rouphael, A.B., Inglis, G.J. 1997. Impacts of recreational SCUBA diving at sites with different reef topographies. Biological Conservation, 82: 329–336.

Rouphael, A.B., Inglis, G.J. 2001. “Take only photographs and leave only bubbles”? An experimental study of the impacts of underwater photographers on coral reef diver sites. Biological conservation, 100: 281-287.

Sale, P.F. 1971. Extremely limited home range in a coral reef fish Dascyllus aruanus (Pisces: Pomacentridae). Coepia, 1971: 324- 327.

Sale, P.F. 2004. Connectivity, Recruitment Variation, and the Structure of Reef Fish Communities. Integrative Comparative Biology, 44: 390-399.

Sandin, S.A., Pacala, S.W. 2005. Fish aggregation results in inversely density dependent predation on continuous coral reefs. Ecology, 86: 1520–1530.

Santavy, D.L., Summers, J.K., Engle, V.D., Harwell, L.C. 2005. The condition of coral reefs in South Florida (2000) using coral disease and bleaching as indicators. Environmental Monitoring and Assessment, 100: 129-159.

Savino, J.F., Stein, R.A. 1989. Behvioural interactions between fish and their prey: effects of plant diversity. Animal Behaviour, 37: 311-321.

Schleyer, M.H., Tomalin, B.J. 2000. Damage on South African coral reefs and an assessment of their sustainable diving capacity using a fisheries approach. Bulletin of Marine Science, 67(3): 1025-1042.

Schmahl, G.P., Deis, D., Shutler, S.K. 2006. Cooperative natural resource damage assessment and coral reef restoration at the container ship Houston grounding in the Florida

210

Keys National Marine Sanctuary, pp. 235-256. In Coral Reef Restoration Handbook. Precht W.F. (Ed.). CRC Press/Taylor and Francis, Boca Raton, Florida, USA. 384 pp.

Schmitt, R.J., Holbrook, S.J. 2000. Habitat-limited recruitment of coral reef Damselfish. Ecology, 81(12): 3479-3494.

Schrandt, M.N. Hardy K.M. Johnson, K.M., Lema, S.C. 2012. Physical habitat and social conditions across a coral reef shape spatial patterns of intraspeciﬁc behavioral variation in a demersal ﬁsh. Marine Ecology. 33: 149-164.

Schroeder, R.E. 1987. Effects of patch reef size and isolation on coral reef fish recruitment. Bulletin of Marine Science, 14 (2): 441-451.

Seaman, W. 2000. Artificial reefs evaluation with application to natural marine habitats, 1st ed. CRC Press LLC. Boca Raton, Florida. 264 pp.

Seaman, W., Jensen, A.C. 2000. Purposes and Practices of artificial reef evaluation, pp. 2-1. In Artificial reef evaluation with application to natural marine habitats. Seaman, W. (Ed.). CRC Press LLC, Boca Raton, Florida. 264 pp.

Seeprachawong, E. 2003. Economic valuation of coral reefs at Phi Phi Islands, Thailand. International Journal of Environmental Issues, 3: 104−114.

Shafer, C.S., 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.

Shafir, S., Gur, O., Rinkevich, B., 2008. A Drupella cornus outbreak in the northern Gulf of Eilat and changes in coral prey. Coral Reefs, 27: 379.

Shafir, S., van Rijn, J., Rinkevich, B. 2001. Nubbins of coral colonies: a novel approach for the development of inland broodstocks. Aquatic Science and Conservation, 3: 183-190.

Shafir, S., van Rijn, J., Rinkevich, B. 2006. Steps in the construction of underwater coral nursery, an essential component in reef restoration acts. Marine Biology, 149: 679-687. 211

Shaish, L., Levy, G., Gomez, E., Rinkevich, B. 2008. Fixed and suspended coral nurseries in the Philippines: Establishing the first step in the “gardening concept” of reef restoration. Journal of Experimental Marine Biology and Ecology, 358: 86-97.

Shaish, L., Levy, G., Katzir, G., Rinkevich, B. 2010. Coral Reef Restoration (Bolinao, Philippines) in the Face of Frequent Natural Catastrophes. Restoration Ecology, 18(3): 285-299.

Shaked, Y., Lazar, B., Marco, S., Stein, M., Tchernov, D., Agnon, A. 2005. Evolution of fringing reefs:space and time constraints from the Gulf of Aqaba. Coral Reefs, 24(1): 165- 172.

Shani, A., Polak, O., Shashar, N. 2012. Artificial reefs and mass marine ecotourism. Tourism Geographies, 14(3): 361-382.

Shemla, A. 2002. Environmental factors that are shaping artificial reef communities. M.S. Thesis, Tel Aviv University, Tel Aviv (in Hebrew, abstract in English). 95 pp.

Sherman, R.L., Gilliam, D.S., Spieler, R.E. 2002. Artificial Reef Design: void space, complexity, and attractants. Journal of Marine Science 59: 196-200.

Shinn, E.A. 1976. Coral reef recovery in Florida and the Persian Gulf. Environmental Geology, 1(4): 241-254.

Shpigel, M. 1980. Niche overlap among two species of coral dwelling fishes of the genus Dascyllus (Pomacentridae). Environmental Biology of Fishes, 7: 65-68.

Shuai, L., Song, YL. 2011. Foraging behaviour of the midday gerbil (Mariones meridianus): Combined effects of distance and microhabitat. Behavioural Processes, 86: 143- 148.

Shulman, M.J. 1985. Recruitment of coral reef fishes: effects of distribution of predators and shelter. Ecology, 66: 1056–1066.

211

Smith, S.R. 1992. Patterns of coral recruitment and post-settlement mortality on Bermuda’s reefs: Comparison to Caribbean and Pacific reefs. American Zoologist, 32: 663- 673.

Sokal, R.R., Rohlf, F.J., 1995. Biometry: the principles and practice of statistics in biological research, 3rd ed. WH Freeman and Company, New York. 887 pp.

Soong, K., Chen, T. 2003.Coral transplantation: regeneration and growth of Acropora fragments in a nursery. Restoration Ecology, 11: 62-71.

Spanier, E. 1994. What are the characteristics of a good artificial reef for lobsters? Crustaceana, 67(2): 173-186.

Spanier, E. 2000. Changes in the icthyofauna of an artificial reef in southeastern Mediterranean in one decade. Scientia Marina, 64(3): 279-284.

Spanier, E., Tom, M., Pisanty, S. 1985. Enhancement of fish recruitment by artificial enrichment of man-made reefs in the southeastern Mediterranean. Bulletin of Marine Science, 37(1): 356-363.

Spanier, E., Tom, M., Pisanty, S., Almog-Shtayer, G. 1990. Artificial Reefs in the Low Productive Marine Environments of the Southeastern Mediterranean. Marine Ecology, 11(1): 61-75.

Spash, C.L., Hanley, N. 1995. Preferences, information and biodiversity preservation. Ecological Economics, 12: 191–208.

Spash, C.L. 2000. Assessing the benefits of improving coral reef biodiversity: the contingent value method, pp. 40−54. In Collected essays on the economics of coral reefs. Coral Reef Degradation in the Indian Ocean Program. Cesar, H.S.J. (Ed.). Kalmar University, Kalmar, Sweden. 243 pp.

Spurgeon, J.P.G., Lindahl, U. 2000. Economics of Coral Reef Restoration, pp.211- 201. In Essays on the economics of coral reefs. Coral Reef Degradation in the Indian Ocean Program. Cesar H.S.J. (Ed.). Kalmar University, Kalmar, Sweden. 243 pp.

216

Steneck, R.S., Graham, M.H., Bourque, B.J., Corbett, B., Erlandson, J.M., Estes, J.A., Tegner, M.J. 2002. Kelp forest ecosystems: biodiversity, stability, resilience and future. Environmental Conservation, 29: 436−459.

Stolk, P., Markwell, K., Jenkins, J.M. 2007. Artificial Reefs as recreational scuba diving resources: a critical review of research. Journal of Sustainable Tourism, 15: 331−350.

Strelcheck, A.J., Cowan, J.H., Shah, A. 2005. Influence of reef location on artificial reef fish assemblages in the north-central Gulf of Mexico. Bulletin of Marine Science, 77(3): 425-440.

Sutton, S.G., Bushnell, S.L. 2007. Socio-economic aspects of Artificial Reefs: Considerations for the Great Barrier Reef Marine Park. Ocean and Costal Management, 50: 829-846.

Svane, I., Petersen, J.K. 2001. On the problems of epibiosis, fouling and artificial reefs, a review. Marine Ecology, 22: 169−188.

Sweatman, H., Robertson, D.R. 1994. Grazing halos and predation on juvenile Caribbean surgeon fishes. Marine Ecology Progress Series, 111: 1–6.

Syms, C., Jones, G.P. 2001. Disturbance, habitat structure and the dynamics of a coral-reef fish community. Ecology, 81(10): 2714-2729. Tews, J., Brose, U., Grimm, V., Tielbörger, K., Whichmann, M.C., Schwager, M., Jeltsch, F. 2004. Animal species diversity driven by habitat heterogeneity/diversity: the importance of keystone structures. Journal of Biogeography, 31: 79-92.

Tietenberg, T. 2006. Environmental economics and policy. Pearson education Inc: Boston, MA, USA. 538 pp.

Tilmant J.T., Schmal G.T. 1981. A comparative analysis of coral damage on recreationally used reefs within Biscane National Park. Florida. Proceedings of the 4th International Coral Reef Symposium, 1: 187-192.

211

Tratalos, J.A., Austin, T.J. 2001. Impacts of recreational SCUBA diving on coral communities of the Caribbean island of Grand Cayman. Biological Conservation, 102(1): 67- 75.

Turgeon, K., Robillard, A., Grégoire, J., Duclos, V., Kramer, D.L. 2010. Functional connectivity from a reef fish perspective: behavioural tactics for moving in a fragmented landscape. Ecology, 91(11): 3332-3342.

Turner, R.K., Button, K., Nijkamp, P. 1999. Ecosystems and nature: Economics, science and policy. Edward Elgar Press, Cheltenham. 520 pp.

Turpie, J.K., Heydenrych, B.J., Lamberth, S.J. 2003. Economic value of terrestrial and marine biodiversity in the Cape Floristic Region: implications for defining effective and socially optimal conservation strategies. Biological Conservation, 112: 223−251.

Uyarra, M.C., Côté, I.M. 2007. The quest for cryptic creatures: Impacts of species- focused recreational diving on corals. Biological Conservation, 136: 77-84.

Uyarra, M.C., Watkinson, A.R., Côté, I.M. 2009. Managing dive tourism for the sustainable use of coral reefs: Validating diver perceptions of attractive site features. Environmental management, 43: 1-16.

Walker, B.K., Jordan, L.K.B., Spieler, R.E. 2009. Relationship of reef fish assemblages and topographic complexity on Southeastern Florida coral reef habitats. Journal of Coastal Research, 53: 39-48.

Wall, M, Herler, J. 2009. Postsettlement movement patterns and homing in a coral- associated fish. Behavioral Ecology, 20(1): 87-95.

Walsh, W.J. 1985. Reef fish community dynamics on small artificial reefs: The influence of isolation, habitat structure, and biogeography. Bulletin of Marine Science, 36 (2): 357-376.

211

Warner, R.R. 1995. Large mating aggregations and daily long distance spawning migrations in the bluehead wrasse, Thalassoma bifasciatum. Environmental Biology of Fish. 44: 337–345.

Werner, E.E., Gilliam, J.F., Hall, D.J., Mittelbach, G.G. 1983. An experimental test of the effects of predation risk on habitat use in fish. Ecology, 64(6): 1540-1548.

Wielgus, J., Chadwick-Furman, N.E., Zeitouni, N., Shechter, M. 2003. Effects of coral reef attribute damage on recreational welfare. Marine Resource Economics, 18: 225−237.

Wilhelmsson, D., Öhman, M.C., Ståhl, H., Shlesinger, Y. 1998. Artificial reefs and dive tourism in Eilat, Israel. Ambio, 27: 764-766.

Wilkinson, C., 2004. Executive summary, pp. 7-8. In Status of Coral Reefs of the World. Volume 1. Wilkinson C. (Ed.), Australian Institute of Marine Science and Global Coral Reef Monitoring Network, Townsville, Australia, 301 pp.

Wilkinson, C.E. 2008. Status of coral reefs of the world: 2008. Global Coral Reef Monitoring Network and Reef and Rainforest Research Center, Townsville, Australia, p. 296.

Wilkinson, C., Souter, D., Goldberg, J. 2006. Executive summary, pp. 6-7. In Status of Coral Reefs in Tsunami Affected Countries: 2005. Wilkinson, C., Souter, D., and Goldberg , J. (Eds.). Australian Institute of Marine Science and Global Coral Reef Monitoring Network., Townsville, Australia, 154 pp.

Williams, I.D., Polunin, N.V.C. 2000. Differences between protected and unprotected reefs of the western Caribbean in attributes preferred by dive tourists. Environmental Conservation, 27: 382−391.

Williams, I.D., Polunin, N.V.C. 2001. Large-scale associations between macro algal cover and grazer biomass on mid-depth reefs in the Caribbean. Coral Reefs, 19: 358–366.

Wilson, A.D.M., Godin, J.-G.J. 2009. Boldness and behavioral syndromes in the bluegill sunfish, Lepomis macrochirus. Behavioral Ecology, 20(2): 231-237.

211

Wilson, A.D.M., Godin, J.-G.J. Ward, A.J.W. 2010. Boldness and reproductive fitness correlates in the eastern mosquitofish, Gambusia holbrooki. Ethology, 116: 96-104.

Yap, H.T. 2000. The case for restoration of tropical ecosystems. Ocean and Coastal Management, 43: 841-851.

Yap, H.T. 2004. Differential survival of coral explants on various substrates under elevated water temperatures. Marine Pollution Bulletin, 49: 306-312.

Yap, H.T. 2009. Local changes in community diversity after coral transplantation. Marine Ecology Progress Series, 374: 33-41.

Yap, H.T., Aliño, P.M., Gomez, E.D. 1992. Trends in growth and mortality of three coral species (Anthozoa: Scleractinia), including effects of transplantation. Mar. Ecol. Prog. Ser. 83, 91–101.

Zakai, D., Chadwick-Furman, N.E. 2002. Impacts of intensive recreational diving on reef corals at Eilat, northern Red Sea. Biological Conservation, 105: 179–187.

Work published or near publication

1. Chapter 6 entitled was published in a paper entitled “Economic value of biological attributes of artificial coral reefs” in Journal of Marine Science.

2. Chapter 7 was published in a paper entitled “Can a small artificial reef reduce diving pressure from a natural coral reef? Lessons learned from Eilat, Red Sea.” in Ocean and Coastal Management.

3. Additional manuscript entitled “Artificial reefs and mass marine ecotourism.” in which I took active part during my study was published in Tourism Geographies.

263

שונית מלאכותית קטנה )שש יחידות של 2×2×2 מ'( מבטון הושקעה באזור הצלילה העמוס ביותר באילת ובקרבתה של שונית טבעית משגשגת. אלמוגים שגודלו במיוחד למטרה זו נשתלו על השונית המלאכותית במטרה לזרז אכלוס של דגים ושפר את היות השונית מוקד משיכה עבור צוללים. השונית נוטרה עבור )1( אכלוס דגים ובחינת שינויים במבנה חברת הדגים )2( גיוס טבעי של אלמוגים ובחינת שרידות של אלמוגים שנשתלו )3( והיכולת לשנות התנהגות של צוללים ולהסיט צוללים משונית טבעית לשונית מלאכותית. דגים אכלסו את השונית במהירות והשכיחות והעושר שלהם השתווה ואף עלה על שוניות טבעיות סמוכות כעבור 360 ימים לאחר הכנסת השונית. נמצא כי הגורם העיקרי המשפיע על מבנה חברת הדגים הוא הזמן ונראה כי חברת הדגים התייצבה לאורך זמן אם כי בצורה הדרגתית ולא רציפה, כנראה עקב גלי גיוס של דגים. שתילת האלמוגים העלתה את שכיחות ועושר מיני הדגים אך לא הייתה משמעותית מספיק לגרום לשינוי במבנה החברה. האלמוגים השתולים נפגעו עקב עקירה על ידי צוללים אשר היו אחראיים ל n=60( 46.7%( מתמותת האלמוגים. תמותה טבעית הייתה גבוהה, בעיקר לאחר השתילה, אך בקנה מידה המקובל גם במאמצי שתילה אחרים. גיוס אלמוגים טבעי היה קבוע וגבוה )16.25±3.5 מתגייסים למ"ר לשנה(. באופן כולל, נראה כי צוללים לא שינוי את אופי הצלילה שלהם בעקבות הכנסת השונית או שתילת האלמוגים. רק בקטגורית הצוללים שבילו זמן קצר בלבד בשמורת האלמוגים )צללו על הגבול של השמורה( נמצא שזמן הצלילה שלהם בתוך השמורה ירד מ- 40-20% מזמן הצלילה ל- 11% לאחר הצבת השונית . לא הובחנה השפעה של שתילת אלמוגים על הצוללים. הערכה של הנכונות לשלם עבור רמות שימור שונות של דגים ואלמוגים, במתווה של שוניות מלאכותיות, נבחנה אף היא. נערך סקר בקרב צוללים בו הוצגו למרואיינים תמונות של שונית מלאכותית עם כמויות שונות של דגים ואלמוגים שהוספו באופן מלאכותי בתכנת עיבוד תמונות. המרואיינים התבקשו לציין את המחיר שהם היו מוכנים לשלם עבור ארבע רמות שימור שונות )שונית עירומה ללא דגים ואלמוגים, רמה נמוכה של אורגניזמים, בינונית וגבוהה( ובשבעה תרחישים שונים )לדוגמה, עליה בכמות האלמוגים בלבד, בדגים בלבד, עלייה בכמות הדגים והאלמוגים יחד וכיו"ב(. צוללים הצליחו להבחין ברמות השימור השונות אך הבדילו בין התרחישים השונים באופן חלקי בלבד. הם העדיפו רמות גבוהות יותר של מגוון מינים ודרגו את תרחיש דגים בלבד בעדיפות הנמוכה ביותר. מחקר זה בחן את הקשרים בגישות האקולוגיות והסוציו- אקונומיות של שוניות מלאכותיות. בעוד המגמה של הכנסת שוניות מלאכותיות נמשכת, חשוב ביותר להבין את הצרכים של שוניות מלאכותיות בהתאם למטרות ההשקעה של השונית. שילוב בר קיימה של שוניות למטרות שימור ושל שימושן על ידי האדם היא אידיאלית אך יתכנו מצבים בהם ייווצר קונפליקט בין השתיים. הגברת הידע, הבנת הצרכים של כל אחד מהגורמים ושימוש נכון ומושכל של שוניות מלאכותיות עם שתילה של אלמוגים, תאפשר לגשר על הפערים בין הגישות השונות.

מילות מפתח: שמירת טבע, ים סוף, אילת, שוניות מלאכותיות, שונית טבעית, שתילת אלמוגים, שיקום, מגוון מינים, מבנה חברה, מורכבות בית גידול, מרחב מחייה, סיכון, טריפה, מחסה, ביטחון, אכלוס, סוקצסיה, נזק, פגיעה, גיוס, התנהגות, נכונות לשלם, שיטת ההערכה המותנית, ממשק.

262

תקציר

שיקום ושימור של שוניות אלמוגים מהווה נושא למחקר תיאורטי ויישומי. כלי ייחודי לשיקום שוניות הוא שימוש בשוניות מלאכותיות. שוניות מלאכותיות מהוות בית גידול חדש ולפעמים כחלופה לבתי גידול קיימים בשוניות טבעיות ההולכות ונעלמות ומספקים בתי גידול לדגים אלמוגים וחסרי חוליות אחרים. אך להשקעה של שונית מלאכותית יש השלכות נוספות. שוניות מלאכותיות מספקות שירותים סוציו-אקונומיים באזורים חופיים ובמיוחד לתעשיית הצלילה. בעבודה זו, בחנתי במקביל את ההיבטים האקולוגיים וסוציו- אקונומיים, הקשורים לשוניות אלמוגים מלאכותיות, כמקשה אחת ואפיינתי דגמים המקשרים בין שני הדיסציפלינות השונות. הרציונאל העומד מאחורי העבודה הוא הצורך להעריך כיצד שונית מלאכותית יכולה להיות ברת קיימא ולשמש אטרקציה )בית גידול( עבור דגים, אלמוגים והאדם )צוללים ומשנרקלים( תוך כדי הסטה של צוללים משונית טבעית סמוכה. כדי לבחון את הקשר בין מבנה שונית מלאכותית והסביבה בו היא נמצאת, ניטרתי ארבעה שוניות מלאכותיות שהועתקו מאזור חולי הסמוך לשונית טבעית אל אזור חולי הסמוך לכרי עשב ים. אומדן חברת הדגים נעשתה על גבי השוניות המלאכותיות, האזורים החוליים, השונית הטבעית וכרי עשב הים, לפני ואחרי ההעתקה של השוניות המלאכותיות. נמצאה עליה בשכיחות, עושר ומגוון המינים. עושר המינים היה תלוי בשכיחות הדגים. חברת הדגים השתנתה בין שני המיקומים כאשר המיקום החדש הראה יותר מגוון בדגים הליליים. כמו כן נמצא שחברת הדגים על השונית המלאכותית שונה משונית טבעית. בשל היכולת לשנות את הרכב כיסוי האלמוגים על שונית מלאכותית )שתילת אלמוגים(, בחנתי את הקשר בין בהתנהגות דגים למבנה בית גידולם בסקלה קטנה. נערך ניסוי עם זוג אלמוגי שיחן שכיח )Stylophora pistillata( ושני דגי אלמוגית שוליים )Dascyllus marginatus( החיים בו, שניהם נפוצים מאוד באזור המחקר. האלמוגים, מנותקי המצע, מוקמו באזור חלוקי האבן במעבדה הימית באילת. המרחק בין האלמוגים שונה באופן אקראי ונמדד מספר המעברים בין האלמוגים של אלמוגיות השוליים. חזרה על הניסוי נערכה באותו מקום כאשר הוספו חפצים שונים להגדלת מורכבות הסביבה וכן בכתם חול במרכז שונית טבעית. התוצאות הראו ירידה לוגריתמית במעבר הדגים כאשר המרחק בין האלמוגים גדל. תוצאות דומות נראו גם בין צמדי אלמוגים בשונית הטבעית. בנוסף, מספר המעברים בבתי הגידול המורכבים יותר )השונית והמורכבות המלאכותית( היו רבים יותר אך דעכו מהר יותר ככל שהמרחק בין האלמוגים גדל. באופן כללי, אלמוגית שוליים נעה עד מרחק של 01-01 ס"מ וגיחות למרחקים רחוקים יותר הן נדירות. תצפיות שנערכו במשך חודש שלם על אלמוגיות שוליים שסומנו בתוך אלמוגים מבודדים בשונית הטבעית הראו כי רוב הדגים )n= 40 ,42%( עברו בין אלמוגים שונים רק ב 11-1% מהפעמים שנבדקו )n=31( והם מבלים את עיקר זמנם באלמוג יחיד.

261

לטל שלי

260

תודות

תודה לכל חברי המעבדה של נדב ששר ובמיוחד לד"ר עמית לרנר על הרעות והעזרה בכתיבת העבודה, לקרן לוי על שחיזקה לי את החשק לצלול, לנועם יוסף שהצמיד לי תמנון לישבן, לסטיב מקקוסקר שאני עדיין חייב לו בירה. ותודה מיוחדת לטל אידן שהפתיעה אותי ביכולותיה ועזרתה הנדיבה. תודה ליובל באר שעזרה לי רבות בהקלדת נתונים וניתוחי וידאו של דגים.

תודה מיוחדת לסגל המעבדה הימית באילת. אתם שימשתם לי משפחה אוהבת במשך 12 השנים שלי במעבדה והייתם הרבה יותר מעמיתים ואולי יותר חברים. הקסם שלכם הוא שהופך את המעבדה לחוויה שאין שני לה. אני לעולם אתגעגע. אני חייב לומר תודה מקרב לב לעודד בן שפרוט שלימד אותי איך לומר לא, למוטי אהביה על הידיים הנפלאות שלו ולמקסים בן לולו שתמיד עזר וחידד, ולחיבור המאגי בין שניהם. לאיציק לרר איש עקרונות שעזר לי רבות בתחילת המחקר וכמובן למלכה אפרת שהיא ללא ספק הרבה יותר מאם הבית. אני מצטער אם הצקתי לכם יותר מידי.

לסטודנטים הרבים שפגשתי שהטו לי אוזן קשבת, נתנו עצה והיו לי כאחים ושקצרה היריעה לפרטם ובעיקר לויקטור קינה, רועי וקנין ועפר בן צבי.

תודה לאין ספור מתנדבים וסטודנטים שצללו איתי גם בימים הקרים של החורף, על אף שאילצתי אותם להוריד חליפה לפני הכניסה למכונית ובעיקר לטובה הנדלמן הנפלאה.

לאנשי רשות הטבע והגנים שעזרו לי ובעיקר לאבי גדליה מנהל שמורת חוף אלמוג וגיא איילון שעזרו בהכנסת שונית תמר ותמכו ברעיון של שתילות אלמוגים ולד"ר דוד זכאי וד"ר אסף זבולוני.

לד"ר משה כיפלאווי שנתן לי חסות לשהות במעבדתו ולעבוד תחת סף דלתו ולשימוש בציוד המעבדה שלו. ולמספר שיחות נפש קצרות אך רבות משמעות עבורי.

לד"ר יוני בלמקר עמיתי וחברי שעזב אותי ללמוד בניכר ושחלק עימי הרבה שעות בתכנון ניסויים ובעיבוד הנתונים ועוד יותר זמן בתמיכה נפשית. אני אוהב אותך ושמח בהצלחתך.

תודה מיוחדת לד"ר יעל גוטר על התמיכה המקצועית והנפשית ועל עצות בזוגיות שתרמו להשלמת המחקר.

לד"ר נדב ששר שפרט להיותו מורה דרך הוא גם בן אדם נפלא. תודה ששמרת על אורך רוח ותמכת בשעות קשות. היה לי כבוד לקחת חלק בחייך ולראות איך סקרנות של חוקר רק גוברת עם הזמן.

ואחרונים. לטל אהובתי, יותם בני שהגיח פתאום לחיי, ולהוריי, שחיזקו אותי והאמינו בי.

261

הצהרת תלמיד המחקר עם הגשת עבודת הדוקטור לשיפוט

אני החתום מטה מצהיר/ה בזאת: )אנא סמן(:

 חיברתי את חיבורי בעצמי, להוציא עזרת ההדרכה שקיבלתי מאת מנחה/ים.

 החומר המדעי הנכלל בעבודה זו הינו פרי מחקרי מתקופת היותי תלמיד/ת מחקר.

 בעבודה נכלל חומר מחקרי שהוא פרי שיתוף עם אחרים, למעט עזרה טכנית הנהוגה בעבודה ניסיונית. לפי כך מצורפת בזאת הצהרה על תרומתי ותרומת שותפי למחקר, שאושרה על ידם ומוגשת בהסכמתם.

תאריך: 23.11.2012 שם התלמיד/ה: עומר פולק חתימה:

261

העבודה נעשתה בהדרכת

ד"ר נדב ששר

במחלקה למדעי החיים, התכנית לביולוגיה וביוטכנולוגיה ימית

בפקולטה למדעי הטבע

היבטים אקולוגיים וסוציו-אקונומיים של שוניות מלאכותיות

מחקר לשם מילוי חלקי של הדרישות לקבלת תואר "דוקטור לפילוסופיה"

מאת

עומר פולק

הוגש לסינאט אוניברסיטת בן גוריון בנגב

אישור המנחה, דר' נדב ששר

אישור דיקן בית הספר ללימודי מחקר מתקדמים ע"ש קרייטמן ______

ט' כסליו תשע"ג 32.11.3213

באר שבע קמפוס אילת, אילת