Economic Management of Recreational Scuba Diving and the Environment

Journal of Environmental Management (1996) 48, 229–248

Economic Management of Recreational Scuba Diving and the Environment Derrin Davis and Clem Tisdell Centre for Coastal Management, Southern Cross University, Lismore 2480, NSW, Australia and Department of Economics, The University of Queensland, Brisbane 4072, QLD, Australia Received 3 August 1995; accepted 24 October 1995

Increasing use of marine protected areas for pursuits such as recreational scuba diving may lead to biological damage and reduced amenity values in popular locations. The relationships between biological and amenity values are discussed and the work of Dixon et al. (1993, Meeting ecological and economic goals: marine parks in the Caribbean. Ambio 22, 117–125) on allocating divers between sites is extended. It is concluded that the carrying capacity concept and the critical thresholds approach are constrained by a number of limitations on their effectiveness as resource management tools. The optimal allocation of users between recreational dive sites is, subsequently, examined and the potential application of economic instruments to achieve such an allocation is assessed. It is concluded that a judicious blend of regulation and the use of economic instruments will be required to overcome open access and boundary problems, and will provide for better overall management of popular marine recreational sites than is presently the case. Education will also have a significant role to play by increasing environmental awareness and reducing the damaging impacts caused by users of those popular sites.  1996 Academic Press Limited

Keywords: scuba, marine protected area, management, thresholds, carrying capacity, economics, recreation.

1. Introduction Recreational scuba diving is an important and growing component of the international tourism market, and is heavily reliant upon natural marine areas. Tabata (1992) and Dignam (1990) identified scuba diving as one of the world’s fastest growing sports, with dive travel being the fastest growing aspect of the sport. The growth in the industry is revealed by an examination of various data on diver training. The Professional Association of Dive Instructors (PADI), the world’s largest dive training organisation, began operations in 1967 and in February 1994 issued its five millionth certification. In 1993, PADI issued 565 000 certifications worldwide and, in the 10 years 1984–1993, experienced an average annual growth in certifications of 229 0301–4797/96/110229+20 $18.00/0

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13%, with the United States, Australia and Japan being the largest markets for scuba certifications. The National Association of Underwater Instructors (NAUI), another significant training body, issues around 130 000 certifications annually. Japan has become an important source of dive travellers in the Pacific Basin, with tours to Chuuk (formerly Truk), Palau, the Philippines and Australia’s Great Barrier Reef being popular. U.S. divers also travel widely, with Florida destinations, the Bahamas and Caribbean locations dominant. Hawaii, Australia and Micronesia are also popular destinations for U.S. divers (Tabata, 1992). While the reasons for the growth in the dive industry are doubtless wide-ranging, the development of more reliable diving equipment which is relatively cheap is an important factor. Kenchington (1990) noted that reliable diving equipment and underwater cameras brought a new perspective to attractive underwater environments. Kenchington referred to such areas as “striking examples of beautiful and fascinating natural environments vulnerable to misuse and abuse by humans” (p. 119). Kenchington’s comments serve to drive home the heavy reliance of divers on attractive natural resources in the marine setting. This includes marine protected areas (MPAs) in many parts of the world—locations such as Bonaire Marine Park in the Netherlands Antilles and Australia’s Great Barrier Reef Marine Park represent very attractive dive destinations. The special features and values of such areas—the reasons they were declared in the first place—are also the reasons that such areas attract divers. The granting of protected area status may also make these areas better known and easier to promote, again leading to heavier recreational use by groups such as scuba divers. Increasingly heavy use of popular dive sites, whether in MPAs or not, seems inevitable. The environmental impacts of recreational scuba diving have been studied only to a quite limited extent (e.g. Hawkins and Roberts, 1992a,b, 1993; Dixon et al., 1993), while the management of diving to prevent such impacts has also received scant attention. Dixon et al. (1993) focused on such management aspects in a study of Bonaire Marine Park in the Netherlands Antilles, using damage functions, carrying capacities and threshold stress levels as important components of their analysis. These are, however, not well defined, mainly because it is difficult to do so, particularly in a marine setting. This paper extends Dixon et al.’s proposals on managing recreational diving in areas where diver numbers are growing rapidly.

2. Crowding, amenity values and biological damage The popularity of MPAs and other natural attractions arises from the amenity tourists realise from visiting such areas. However, the benefits to be appropriated from the use of these natural assets are affected by the extent of competitiveness in the use of the resource in question (Davis and Tisdell, 1995). This idea is often discussed by economists in terms of rivalry and excludability (Dixon and Sherman, 1990). Non-rivalry occurs when one person’s consumption does not affect the total amount available to anyone else. That is, the total amount of a good available can be enjoyed by anyone without diminishing the supply of that good. However, Dixon and Sherman made the point that some goods, and particularly recreational goods, are in fact “congestible”. These goods are non-rival up to a certain level of usage; beyond that level rivalry will set in. Hundloe (1979) stressed this point in the following way: “It is surely not possible for the ever increasing number of nature seekers to venture into the remaining wilderness or near natural areas without changing

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the nature of the satisfaction sought. Disregarding ecological impacts (which, presumably, could be controlled by regulating visitor numbers), there is a very real loss of satisfaction to some visitors if they are but some of a multitude that have visited the area” (p. 175). Hundloe’s suggestion about regulating visitor numbers relates to excludability. A nonexcludable good, according to Dixon and Sherman (1990), is one where the cost of excluding consumers is greater than the benefit received. A substantial degree of nonexcludability exists in most marine areas with certain exceptions such as in the Scientific Preservation Zones of the Great Barrier Reef Marine Park. Regulation of commercial operators in MPAs is, however, possible through such means as permit systems. Biological damage is also a potential concern as diver numbers increase at popular sites. Davis and Tisdell (1995) reviewed the available evidence concerning diver-induced biological damage from recreational scuba diving in MPAs, and concluded that, presently, there is only circumstantial evidence that severe biological degradation or loss of biological diversity results from these activities. However, no long-term monitoring studies have been undertaken of diver-induced impacts, even in very heavily used areas, despite concerns about critical threshold levels of use of particular dive sites (Dixon et al., 1993; Scura and van’t Hof, 1993). In other words, while there is limited evidence on biological degradation and reduced biological diversity, this may be the case because critical thresholds have not been reached in those few areas so far studied. If, however, diver numbers escalate substantially, then such thresholds may be reached and exceeded. For example, Hawkins and Roberts (1992b) examined the use of certain Egyptian reef sites and concluded that the present level of usage of 30 000 to 50 000 dives per year was approaching the maximum sustainable biological capacity of those reefs. However, they also concluded that a projected ten-fold increase in diver numbers would result in serious environmental damage. The biological threshold is, presumably, somewhere between the present level and the projected level of use. These findings drive home the difficulty of estimating the carrying capacities of dive sites, and raise doubts about whether critical thresholds exist and, if they do, where they lie. Nevertheless, damage caused by divers is still important, even if biological impacts are minimal. A significant proportion of the amenity values associated with scuba diving relate to the wilderness experience realised and the aesthetic quality of a particular dive site. Heavy use of a site is very likely to lead to significant amenity value losses from diver-induced damage (reducing aesthetic quality) and overcrowding (reducing wilderness and other experiences) (Liddle and Kay, 1987, 1989; Phillips, 1992; Dixon et al., 1993; Hawkins and Roberts, 1992a,b, 1993; Davis and Tisdell, 1995).

3. Carrying capacities in recreational diving The application of the carrying capacity concept to recreation areas became popular with the development of a methodology for assessing carrying capacity in 1973 (Urban Research and Development Corporation, 1980). This methodology was developed because of concerns about “resource overuse and user overcrowding” of recreation resources. The idea is, therefore, relevant in the context of recreational scuba diving and the associated concerns about biological damage and loss of amenity at heavily used dive sites. Phillips (1992) in a study of Julian Rocks Aquatic Reserve (a small MPA on the Australian east coast described in Section 6 below) concluded that, at least during peak

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periods, the carrying capacity of the Reserve’s dive sites was being exceeded. The carrying capacity concept used by Phillips implied both biological and human aspects, based as it was on a definition of carrying capacity which was focused on the “level of use . . . which a natural resource can sustain without an unacceptable degree of deterioration of the character and quality of the resource or the use of that resource” (Prosser, 1986, as quoted by Phillips, 1992, p. 67). Phillips referred also to divers being discouraged from visiting the area due to the sense of overcrowding or congestion which may be experienced. Scura and van’t Hof (1993), in a study in Bonaire Marine Park in the Netherlands Antilles, concluded that diver impacts at certain sites had exceeded an acceptable level, and stated that “visitation at these sites has exceeded the carrying capacity” (p. 25). This finding is based principally on biological considerations, although the issue of overcrowding is also addressed. The conclusions on carrying capacity reported above relate directly to biological and human perceptions of the carrying capacity concept. There are, additionally, other dimensions to the concept, including economic dimensions. These relate particularly to the consumers’ surplus realised by individuals from the availability and use of an asset such as a recreational diving site. Consumer’s surplus depends upon such factors as congestion at and the number of trips to a recreation site (McConnell, 1980), along with the physical state of the site. The carrying capacity idea, while conceptually appealing, is constrained by a number of limitations on its effectiveness as a resource management tool, with the concept being far from specific (Tisdell, 1988). The Resource Assessment Commission (1993) noted that translation of the concept into practice is “fraught with difficulty for those resources where the relationships between use levels and impacts are neither simple nor uniform” (p. 8). There is a danger also in mistaking carrying capacity estimates as finite limits or thresholds (Schneider et al., in Resource Assessment Commission, 1993). Consideration of these issues raises a series of questions. Chief amongst these is whether critical thresholds and carrying capacity estimates are useful in managing assets such as dive sites and, if they are, will it be biological values or amenity values which imply lower thresholds and carrying capacities. Second, what is the role of economics in managing environmental assets like dive sites? Consideration of economic aspects includes questions of substitutability between sites and the optimal allocation of divers between them. It involves also consideration of the use of economic instruments in managing the impacts of recreational activities (Clarke et al., 1995). Finally, and related to the above questions, is the issue of what management strategies—including the use of economic instruments—might be put in place to manage recreational dive sites. 4. Critical thresholds: biology and amenity Dixon et al. (1993) and Scura and van’t Hof (1993) made important contributions to the understanding of the ecological impacts of diving and of diver carrying capacities in MPAs. In the study of Bonaire Marine Park they concluded that there is a critical level of use above which reef degradation sets in. This critical level was estimated as 5000 dives per year per site, with this figure based on both qualitative data from diver surveys and quantitative data derived from a photoanalysis of coral cover and species diversity. By spreading diving over a range of alternative sites, Scura and van’t Hof suggested that the carrying capacity of the marine park is approximately 200 000 dives per year.

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D

Stress level

Threshold stress levels B A

O

D1 F E

M Number of dives

C

Figure 1. Threshold stress levels of diver numbers (after Dixon et al., 1993).

Dixon et al. (1993) developed a simple model of the relationship between diver density (number of individual dives) and threshold stress level (Figure 1). This model relates a perceived stress threshold on the marine ecosystem to the intensity of diver use. Level A represents the threshold stress level where degradation to the reef becomes noticeable. Dixon et al. suggested that impacts below this level are minimal or even non-existent. But above the threshold there is “a loss of coral cover, reduction in species diversity, decreased visibility and other impacts” (p. 124). Line OD in Figure 1 is a “damage function” which represents the impact of diver use on the marine park, with use measured by the number of individual dives per year. Point E is the number of dives (M) where the threshold is reached. According to Dixon et al.’s analysis, M would be 5000 dives per site per year in Bonaire Marine Park. Dixon et al. argued that this might be increased through better park management, raising the threshold from, say, A to B. Better park management might, they suggested, arise from strategies such as rotation of dive sites, spacing out divers and regulating underwater photography. Improved diver education might also increase the number of divers who can use the park without causing degradation. In Figure 1 this would be shown as a rotation of the damage function from OED to OFD1. Diver education would be expected to lessen the impacts of individual divers by encouraging improved buoyancy control and reef etiquette. The combination of improved management and diver education increases the carrying capacity to OC dives. In the case of Bonaire Marine Park, Dixon et al. concluded that these strategies would allow for a doubling of individual dives to 400 000 per year. These results are based on ecological impacts as observed by divers and according to photoanalysis of reef sites. Yet the results are very subjective in the absence of long term monitoring. This is emphasised by comparing the results reported above with those of Hawkins and Roberts (1992b) who concluded that 30 000–50 000 dives per year at certain Egyptian coral reef sites was approaching the maximum biological capacity of those reefs. It is, therefore, unclear whether the use of critical thresholds and/or carrying capacity concepts which are based on likely ecological impacts are useful in framing management responses to perceived pressure from recreational diving. Long-term monitoring studies are needed to identify critical thresholds, while further

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information is required on questions such as the recovery (regrowth) rates of different species of coral. If the critical thresholds approach is of limited usefulness (Tisdell, 1988), are there other guides to the capacity of dive sites and, consequently, the management of those sites? One approach may be to focus still more on the perceptions of divers, particularly as these perceptions relate to amenity values and overcrowding. The evidence available here is limited and confusing. On the one hand, it seems that high quality, apparently undisturbed underwater environments serve to attract divers. The “wilderness experience” gained by divers is also thought to be an important determinant of the demand for diving and dive sites (Hundloe, 1979; McKinnon et al., 1989). In this regard, Kenchington (1993) related wilderness to a very low level of “consequential impact” upon the environment. The wilderness experience may also be impacted by crowding when too many divers are using a site. On the other hand, Scura and van’t Hof (1993) noted that “some reef degradation” does not seem to have resulted in loss of attraction to divers. Tabata (1990, 1992) also commented that divers do not require pristine conditions in all cases. In Julian Rocks Aquatic Reserve a degree of apparent (though unproven) degradation had not discouraged divers, although crowding was a growing issue (Phillips, 1992; Austin, 1993). Pursuit of a wilderness experience, based on Kenchington’s (1993) framework, implies a number of dives less than M in Figure 1. If only a very low level of “consequential impact” is acceptable then the number may be considerably below M. Conversely, if there is no “loss of attraction” (Scura and van’t Hof, 1993) from some level of degradation, then the number of dives may be at, or even above M for some period of time. Again, there is uncertainty relating to diver perceptions of amenity values, wilderness experience, crowding, and the “acceptable level” of degradation of a dive site. Resolution of these questions requires studies in which the perceptions of divers are examined and efforts are made to establish when amenity values begin to decline.

5. Biology, amenity and economics Both biological and amenity losses are, ultimately, economic losses through reduced value of the dive sites being considered. Tisdell (1991) summarised this in a more general way in noting that tourism can destroy tourism in two ways. First, overcrowding may reduce the total benefits received from tourism and deter some tourists from a particular area. Second, tourist pressures and facilities designed to service tourists may degrade, partially or completely, the attraction which drew tourists in the first place. These generalised comments can be applied to special interest tourism such as scuba diving, notwithstanding that environmental assets such as coral reefs are living systems which can recover from damage whether caused by natural or human impacts. Marine protected areas are declared principally because of a concern to maintain ecological values. However, most MPAs are also managed as multiple-use areas with scuba diving being one of those (increasing) uses. Following Scura and van’t Hof (1993), the question arises as to whether MPAs can be managed to produce both ecological benefits as protected areas and economic benefits from activities such as dive tourism. The issue of diver numbers and possible threshold impact levels is a key aspect of this question and, subsequently, of the approach to managing dive tourism in MPAs.

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In drawing together the economics and ecology of scuba diving in MPAs a range of considerations arise. These include the following. (i) The quasi-public-good nature of most MPAs and the existence of market failure. (ii) Determinants of demand for particular dive sites. (iii) Threshold or critical levels of usage in terms of both ecological impact and diver perceptions of crowding and changed amenity values. (iv) Willingness to pay and consumer surplus related to specific sites. (v) Related to (i)–(iv) above, the question of how to manage scuba diving to achieve efficiency in the allocation of resource use (that is, maximising net social benefits from the use of certain sites in scuba diving). Thampapillai (1991) pointed out that most emvironmental goods are quasi-publicgoods, that is, they exhibit public-good properties up to a certain point, beyond which they become private goods. Thampapillai used the example of national parks, yet scuba diving sites could similarly provide an example. A certain number of divers can enjoy the amenities of a dive site without diminishing the enjoyment of each others’ use of these amenities (and the number, over a year, may be quite large). That is, up to a certain level of use the marginal cost of providing the amenity value of a dive site is zero. However, beyond some point congestion sets in, the marginal cost becomes positive and finally tends to infinity at high levels of congestion. The public good nature of the resource (the dive site) also creates market failure. Until congestion sets in the price for use of the good is zero. This is typical of many environmental goods and means that their “true” value (total economic value) has been underestimated or, commonly, ignored altogether (Turner et al., 1994). According to Turner et al., the fact that such goods have remained unpriced means that they have been “inefficiently exploited” (p. 26). This, in turn, leads to external impacts (damaged corals, impaired ecological functions, reduced amenity values) which are not compensated for by those who have benefited from the use of the site. Therefore, the danger for dive sites is that they may be subject to excessive use, overcrowding, and biological degradation. The foregoing discussion serves to highlight aspects of points (ii) and (iii) above, relating to determinants of diver demand and threshold levels, respectively. As discussed previously, the demand for particular dive sites is a function of many variables. Two of the more important variables are likely to be price and environmental quality, with “quality” including aspects such as aesthetic appeal, interesting marine life and visibility. The quality aspects, as discussed by Hawkins and Roberts (1992a,b) and others, will be critically affected by diver numbers. Furthermore, and as discussed by these authors, along with Dixon et al. (1993), there may be a critical threshold after which biological damage becomes severe and perhaps irreversible, although this is, as yet, unproven. Tisdell (1991) provided an analysis of the effects of crowding on tourism resources, and made the point that an individual’s demand to visit a tourist site (which could be a dive site) may be affected not only by price but also by the number of tourists (divers) present. This implies that individuals who are averse to crowding will exhibit lowered willingness-to-pay as the number of other users increases. The implication for managing resources such as dive sites is that aversion to crowding reduces consumer satisfaction and, therefore, the number of users should be restricted by some means. User permits or increased fees may, for example, be among the strategies to restrict diver numbers,

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as may industry self-regulation, as occurs in the Ribbon Reef area of Australia’s Great Barrier Reef Marine Park (Department of Environment and Heritage, 1992). Over-crowding at dive sites may lead also to excessive deterioration of those sites. Tisdell (1991) explained that the relationship between tourism development, the destruction of natural assets and the maintenance of tourism is a complex one, and made a number of important points in this regard. First, the initial demand curve—the curve existing when an area is in a pristine state—may be somewhat different to the aggregate demand curve for use of an area when account is taken of the deterioration of the area. Tisdell concluded that it is important to prevent an over-expansion in tourism facilities in natural areas in the first place as this is likely to lead to overcrowding and deterioration in the asset upon which tourism relies. The suggestion is that some limit on activities such as scuba diving may be socially optimal in terms of crowding and resource deterioration. If critical biological thresholds do indeed exist, as discussed by Dixon et al. (1993) and Hawkins and Roberts (1992a, 1993), then such limitations may have a further effect of conserving ecological values. 6. Diver demand with no substitutes The analysis based on the use of damage functions, along with the analyses of crowding provided by Tisdell (1991), is a special case in which only one natural asset (or site) is being considered. However, the availability of substitute resources will affect the use of an environmental asset and, consequently, the management of that asset. Consider first a case where there are few or no effective substitutes. An example is provided in the Julian Rocks Aquatic Reserve which is located approximately two kilometres off the coast, north-east of the township of Byron Bay, the most easterly point of the Australian mainland and a popular holiday destination. Wright (1990) described the Reserve as one of the best diving locations on the Australian east coast. The Reserve is in a tropical–temperate overlap zone and contains an abundant marine community including 33 species of coral and more than 460 species of fish. The Reserve also features interesting geological features such as tunnels, caves and sand gutters, and is known for the presence of marine turtles and, at certain times of the year, for the congregation of grey nurse sharks. However, diving in the Reserve is concentrated heavily in one or two sites which are adjacent to the rocky outcrop known as Julian Rocks. Phillips (1992) and Copeland and Phillips (1993) raised concerns about intensive use of Julian Rocks for scuba diving citing, amongst other impacts, anecdotal evidence of a reduction in hard coral cover of 50% in the past 10 years. Why might this happen? One obvious answer is that the price to use the resource is too low. For resources such as Julian Rocks it is, in fact, zero. This leads to a degree of overuse (the “inefficient exploitation” referred to by Turner et al., 1994) and, consequently, the likelihood of thresholds being exceeded and of irreversible biological damage occurring. The price of the asset may need to be increased substantially to overcome this effect. The interesting outcome, however, for Julian Rocks (and presumably other, similar sites) is that critical thresholds are said to be exceeded, biological degradation supposedly apparent and amenity values reduced (Phillips, 1992; Copeland and Phillips, 1993), yet it remains a very popular dive site. Three causes of this popularity may be suggested. First, the site is extremely well known to the Australian diving fraternity and has a substantial reputation as a fine dive site. Consequently, many divers and dive clubs have it on their “must visit” list, and there is a continual stream of divers to the area

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who have not previously dived at Julian Rocks. Again, uncertainty over whether biological degradation is significant, and whether thresholds have been exceeded, is apparent—neither of these supposedly undesirable results has been proven. Further, diver preferences for the use of the site seem to indicate that, in other ways, the carrying capacity has not been exceeded. Second, Julian Rocks is nearby to the township of Byron Bay which is a very popular holiday destination. Therefore, a number of people combine a beach holiday with some diving. Third, recreational diving is typified by a high drop-out rate. The Professional Association of Dive Instructors (Australia) estimates that, on average, divers are “active” for a period of only 21 months (Windsor, pers. comm.). At the same time there is a continual stream of learners and new divers entering the sport. These new divers may be less averse to dive site degradation where it does occur, possibly because they are unable to make comparisons with other sites. Taking these factors into account, it seems likely that there will be continued heavy pressure on “no substitute” sites such as Julian Rocks, and that these sites may become degraded. If degradation is proven, at least three management responses are apparent. The first is to restrict the number of divers either by regulation or by increasing the price to use the site. However, the open access nature of the resource renders this option difficult. A second possibility would be to create one or more substitute dive sites. This could be achieved, for example, by creating an artificial reef. Such structures have been widely used for diving and for both recreational and commercial fishing in many locations around the world (Kerr, 1992). The availability of a substitute resource may lessen the intensity of use of the existing natural asset, even if the two are not perfect substitutes. Third, it may be possible to alter the allocation of divers between existing sites. The potential of this strategy is limited at Julian Rocks because there are relatively few substitute dive sites. However, in many areas it will be a possible solution to localised pressures. But how can a reallocation of divers between sites be effected? It is on this question that the remaining focus is placed. 7. Distributing divers between sites The situation is likely to be somewhat different from Julian Rocks when a range of substitute dive sites exist as occurs, for example, in the Great Barrier Reef Marine Park. Scura and van’t Hof (1993) also raised this point in relation to Bonaire Marine Park, concluding that “a more even distribution of dives among the most popular sites is . . . required at present levels of visitation in order not to exceed carrying capacity” (p. 27). Therefore, consideration should be given to the allocation of users between different recreational sites. In most cases dive sites will not be perfect substitutes. For example, ease of access to a site is generally thought to be a critical determinant of the recreational use of that site (Kenchington, 1984). Demand for certain sites, consumer surplus and the utility (amenity) derived from the use of those sites are important, interrelated aspects of optimally distributing users among them. Utility in the case of recreational scuba diving may depend on factors such as ease of access (which also determines cost), conditions at the site (surface conditions, current, etc.), the “quality” of the diving (underwater visibility, marine life, etc.) and the availability of substitutes. Consider the case where only two substitutable dive sites are available but one (dive

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Number of visits to site 2

x2

I1 I2

I3

d1

d I3

B2 B1

U1

I2 I1

A2 A1 c1 Number of visits to site 1

c

x1

Figure 2. Dive site preferences and optimal choice of usage.

site 2) is less accessible than the other (dive site 1). The indifference map for an individual diver might be like that shown in Figure 2, where the indifference curves are assumed to be “well-behaved” Hicksian type curves (Tisdell, 1972). The difference in the relative accessibility of the sites is represented in Figure 2 by the size of the slope of line dc which is drawn on the assumption that the individual diver allocates a given budget to diving. As with the travel cost method of evaluating the demand to use outdoor recreation sites, the cost of access (e.g. travel costs to the site, possibly including an economic allowance for the time involved also) is used as a proxy for the price of access. The theory therefore being applied is neo-classical demand theory as expounded by Hicks (1946). According to Hicks’ theory, this will result in this diver consuming a unique, optimal mix of dive sites of a level of A1 for site 1, and B1 for site 2. That is, based on the relative prices of the two sites, utility is maximised at Point U1 on indifference curve I1I1, implying heavy usage of site 1 and light use of site 2. At Point U1 utility is maximised because: M1/P1=M2/P2 where M1 and M2 is marginal utility from sites 1 and 2, respectively, and P is price. As noted, a diver confronted by the budget line dc will distribute his/her diving between the sites to a level of A1 dives at site 1 and a level of B1 dives at site 2. If a management agency becomes concerned about damage from the level of use of dive site 1—the more accessible (cheaper) site—then the agency might respond by changing the relative prices of the two sites. An increase in the cost to divers of using site 1, in combination with reducing the cost of site 2, would result in a rotation of the budget line to a new position such as d1c1, and a new combination of site visits of A2B2 (cet. par.). This assumes that the total cost of the use of the two sites has not changed. The relationship M1/P1=M2/P2 has, therefore, not changed. The diver achieves the same level of utility but changes his/her combination of the use of each site. A reduction

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in the cost of a dive site might be achieved by the installation of a boat ramp or moorings, while an increase might be effected by removing supporting infrastructure or, perhaps, by applying a cost (such as an entry fee) to use the site. The existence of boundary problems (Clarke et al., 1995) may render this latter option difficult in marine areas. This point is revisited in a later section. This analysis refers to the individual diver, and will determine that individual’s demand for diving at each site. If all divers react in a similar way to changed prices, then aggregate demand would be expected to change also, leading to greater usage of the now cheaper site, and reduced use of the site which is now more expensive to access. 8. Allocating divers amongst multiple sites Having considered the “no substitutes” and “only one substitute” cases, the situation where a number of different dive sites are available should also be assessed. As previously suggested, this is the case in most parts of the Great Barrier Reef Marine Park, as well as in Bonaire Marine Park. Additionally, the objective function needs to be spelled out more clearly. Possible objectives include the following. 1. Minimising the value of total damages caused by a given number of dives. 2. Maximising the net benefits obtained by divers. 3. Maximising benefits for the whole community, both divers and non-divers. The analysis to this point has tended to be focused on objective 2, with consideration of objectives 1 and 3 being implicit rather than explicit. Assuming that the number of dives in a region is a fixed number, the problem is how those dives should be allocated between a fixed number of alternative dive sites. The fixed number assumptions are not restrictive because, if the numbers are variable, in the final analysis the number of divers and the number of dive sites used is a particular number. The allocation must therefore satisfy the same necessary conditions as determined in the fixed aggregate numbers’ case. The three objectives identified above are now considered. 8.1.       Solving this problem requires that the damages caused at any site are able to be related to the number of dives at that site per unit of time, and that these damage costs are valued. The costs may, in principle, be determined by obtaining information on the willingness-to-pay to avoid damage. Importantly, the optimal allocation of divers will be influenced by the nature of the damage functions. Additionally, there is a question of whose “damage value” is to be minimised. In this case, it is presumed that a management agency which is responsible for a marine protected area is concerned to minimise the value of total damages, and the analysis relates, therefore, to such an agency’s aims. Suppose that m dives are to be allocated and that n dive sites are available. Let fj(xj) represent the cost-damage function at site j in money terms where xj represents the number of dives at site j. The problem is to minimise n

C=

] f (x ) j

j=1

j

(1)

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subject to n

] x =m j

(2)

j=1

If the damage functions are everywhere differentiable and the total cost of damage increases everywhere at an increasing rate, then the necessary and sufficient condition for the optimal allocation would be f ′1(x1)=f ′2(x2)=. . .=f ′j (xj)

(3)

subject to constraint (2). This implies that the total number of dives should be allocated so that the marginal cost of damage is equal at all the alternative sites. However, if critical thresholds occur (as suggested by Dixon et al., 1993), then account should be taken of the possibility that no damage may occur until a certain threshold number of dives is evident at each site. Let the threshold value of the cost of damage at site j, when damage just commences, be represented by Aj. It follows that if n

mΖ

]A

j

(4)

j=1

then the number of dives can be allocated between the sites so as to cause no damage. Diving should not be allowed at any site on a scale likely to cause damage. If, on the other hand n

]A

m>

j

(5)

j=1

then this is impossible. If fj(xj) for xj>Aj increases at an increasing rate, optimality will then require condition (3) to be satisfied and consequently some damage will occur at all sites. Conversely, if the damage cost functions are increasing at a decreasing rate, an areal concentration in diving activities will be required once condition (5) applies. Every site should be utilised in this case up to the threshold where damage starts, except for that (one) site where damage costs, in relation to the threshold, are lowest. This site should be allocated m−R Aj dives; that is, all dives which cannot be allocated so as to avoid damage at the available sites. In this case the total cost of damage increases at a decreasing rate when it occurs.

8.2.        Divers will normally have unrestricted access to available, popular dive sites. Consequently, they may not regulate their numbers and the allocation of dives between sites so as to maximise their aggregate net benefits. Returning to the single site case,

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$

A

O

x2 x1 it ef en lb na gi ar M

Av e

ra

ge

be

ne

fit

C

B Number of dives at site Figure 3. Loss due to unrestricted access.

divers will tend to use the site until they just exhaust their net benefit by so doing (Figure 3). This is analogous to the dissipation of natural resources rent as in the case of open access fisheries (Magrath, 1989; Tietenberg, 1992). If, in Figure 3, AB represents the marginal benefit from diving at the available site and AC the average benefit, optimality calls for x1 divers, but x2 will result from unrestricted access and the deadweight loss indicated by the hatched area will occur. However, consider the case in which n dive sites are available and let Bj represent the aggregate net benefit obtained by divers at site j. Then Bj=hj(xj)

(6)

where hj represents the benefit function at site j. Assume that hj increases but at a decreasing rate. Optimality requires that h′1(x1)=h′2(x2)=. . .=h′j (xj)

(7)

that is, that marginal aggregate benefits are equal at all the sites. But if unrestricted access is the rule, the number of divers will not be distributed so as to maximise net aggregate benefits, but will allocate themselves among dive sites to maximise benefits per dive (average benefits) rather than marginal benefits at each site. That is, the relationship h1(x1)/x1=h2(x2)/x2=. . .=hj(xj)/xj

(8)

will be satisfied. This is likely to result in too many dives occurring at the sites yielding the greatest net benefits on average compared to those yielding a lower net benefit on average. This again is analogous to the fisheries case in which the most productive fishing grounds are over-exploited relative to less productive areas when unrestricted access is allowed.

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8.3.      The focus in the discussion has, so far, been exclusively on scuba divers. However, a range of other groups such as recreational and commercial fishers, along with day trippers, may wish to use the same sites as divers. These other users need also to be taken into account if an optimal use of the various sites is to be achieved. Again, take the case where n sites are available, and where the sites are used also by two other groups of recreational users—day trippers (for snorkelling and viewing via, for example, semi-submersible vessels) and recreational fishers. These groups, like divers, will also have an impact on the ecological and amenity functions provided by a range of sites, although the cost functions will be different for each user group. The problem then is to optimally allocate all users across the available sites. The problem is to maximise n

] (B +F +R )

Wj=

j

j

(9)

j

j=1

where Wj=aggregate benefit to all three user groups at site j; and Bj=hj(xj, yj, zj)

(10)

Fj=hj(xj, yj, zj)

(11)

Rj=hj(xj, yj, zj)

(12)

The net benefit obtained by divers is, as before, Bj, with xj representing the number of divers at site j. Similarly, Fj is the net benefit obtained by recreational fishers, with their numbers represented by yj, and the net benefit to day trippers is Rj, associated with zj users. The nature of the functions allows for interdependence between the different types of users of sites. Maximising the net benefits to all the user groups is an extension of the case where divers are allocated between sites. That is, optimality requires that marginal aggregate benefits are equal at all sites. In an open access, unconstrained situation, however, users again will seek usage patterns such that their marginal costs equal their average benefits. Additionally, as previously discussed, concern may arise about threshold use levels and damage to biological and amenity values at individual sites. Each of the three groups of users will impose different costs on the sites being utilised. Assuming that use levels are non-zero and positive, the aggregate net social benefit (NSB) from recreational use of sites may be represented as n

] [(B −C )+(F −C )+(R −C )]

NSB=

j

j0

j

j1

j

j2

(13)

j=1

where Cj0 is the cost function associated with use by divers, and Cj1 and Cj2 the cost functions associated with use by fishers and day trippers, respectively. The costs referred to here are environmental costs, including both biological and amenity costs. At this point it becomes clear that the analysis, and decisions about optimally allocating users among natural sites, is a complex matter. Most significantly, it is

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hamstrung by the difficulties associated with estimating the critical variables. Considerable difficulties are encountered in estimating the biological and social carrying capacity of one site which is being used by only one category of users. These difficulties are compounded when there are several sites being used by a range of user groups. Furthermore, the environmental damage costs are difficult to estimate, especially given the need to distinguish natural damage from human impacts. Relatedly, defining critical threshold use levels, if they exist, remains an unresolved yet important matter. Further work by both biologists and economists is needed on these issues. This is not, however, an argument to do nothing—concerns about likely environmental damage, reduced recreational values and so on, mean that management strategies need to be devised which will safeguard the conservation and amenity values provided in MPAs.

9. Management strategies—allocating divers between sites In an effort to redress allocation problems related to the use of natural resources, governments, through their management agencies, can choose from a range of instruments. These may “range from prescriptive regulations through to direct market mechanisms, including tradeable rights and user charges” (ABARE, 1993, p. vi). However, the management of a resource may entail multiple objectives (as in the multi-use zones of some marine parks) and, therefore, it will often be necessary to use a combination of several different management instruments (ABARE, 1993). Dixon et al. (1993) concluded that Bonaire Marine Park could sustain around 200 000 dives per year. However, they did not devise strategies for ensuring that these guidelines would be met, nor did they indicate how divers could be encouraged to use lesser dived sites in preference to those currently more favoured. Dixon et al.’s work served to highlight important management issues and also developed, conceptually, possible strategies for increasing the number of divers who could use sites without degrading them. But the question of how to reallocate divers amongst sites remains. There are a variety of instruments which could be applied to the management of scuba diving in MPAs. Indirect interventions are possible by such means as varying the ease of access to individual sites (thereby raising or lowering the cost of using those sites), or via more direct means such as placing a limit on the number of divers who may use a site each year, or introducing a system of licensing individual divers. While recreational divers are certified to dive, there is no requirement in the certification system which governs particular kinds of underwater behaviour or activities. The only controls here relate to bans on activities such as collecting and spearfishing on scuba in most MPAs. It would be possible for management authorities to limit diving at certain sites to divers with a particular kind of licence or certification. Education is also likely to have a significant role to play, and various diver training organisations and some management agencies have developed some initiatives in this regard. The major training agencies offer environmental awareness programs for divers, while the Great Barrier Reef Marine Park Authority, in cooperation with the dive industry, is developing a train-the-trainer package for scuba instructors in “Ecological Diving”. Clarke et al. (1995) pointed out that visitor behaviour, as well as visitor numbers, determines environmental impacts and costs. They noted that the difference between good and bad behaviour may make a large difference to those costs. Presumably, education can reduce this “bad” behaviour. Dixon et al. (1993) noted the potential of education to reduce environmental damage and/or allow a larger number of consumers

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to use a site. Educational programs, however, unlike licensing, regulation or changing ease of access, will not change the allocation of divers amongst sites. An alternative approach could be to implement one or more economic instruments in order to change demand patterns and, consequently, to reallocate divers amongst sites. There is a wide range of such instruments which can be used to control environmental impacts, reduce market failure and, subsequently, externalities in natural settings. These have been reviewed by Tisdell (1993), James (1993), Turner et al. (1994), and other authors. Their use specifically in a tourism setting has been examined also by Clarke et al. (1995). The best known group of economic instruments is the price-based instruments, particularly user-pays. Clarke et al. (1995) interpreted this as the application of a charge to use a fragile environment. Desirably, such charges would be related to the costs of environmental damage, or the monetary costs to avoid such damage. But, as Clarke et al. pointed out, user charges are often like a Pigovian tax on the product that generates the externality, not the externality itself. This would be the case for damage caused by recreational scuba diving where a cost might be applied to use a site, in the expectation that this will reduce demand for, and usage of the site. The difficulties likely to be encountered in putting a monetary cost on the damage to, say, coral reefs, means also that this would be a pragmatic approach to charging users. Care needs to be taken in varying the relative cost, accessibility or surrogate price for access to an over-utilised diving site. If everything is left unchanged, except that the cost or price of access to an under-utilised site(s) is reduced and if consumption of experiences at diving sites is normal (normal goods), the “income” effect could result in utilisation increasing at all dive sites. By the same token, merely raising the price of access to the over-utilised site, other things left unchanged, could reduce diving at all sites because of the “income” effect. Detailed empirical analysis is needed to determine how large the income effect is likely to be relative to the substitution effect. As outlined previously, to achieve the aim of reallocating divers from one site to another while keeping the number of aggregate visits constant, both prices need to be varied. This follows from the normal properties of the economic analysis of choice based on indifference curves. The adoption of a user-pays approach assumes also that it is possible to estimate the prices needed to bring about the desired responses. In other words, information on demand elasticities for each site is needed. While it will be possible to estimate these elasticities when there are multiple sites (and as yet unused sites) available to divers, this will be a substantial undertaking. Further, the elasticities may change over time as incomes or tastes change. A related issue is that of how the prices for diving at different sites can be varied. One approach may be to charge different amounts for permits to charter operators for the use of different sites in the expectation that these charges will be passed on to their clients. In conducting dive holidays which involve multiple dives, these operators will weigh the differing permit costs against their other costs such as travelling time between sites and ease of access to different locations. Consequently, it is not immediately obvious what the net effect of differential permit fees might be, given also that divers will not necessarily be aware of those differentials. Furthermore, this strategy does not affect the use patterns of private divers. Finally, this is quite an indirect method of charging for the use of dive sites and, as observed by Clarke et al. (1995), indirect pricing solutions are likely to be far less efficient than direct charging approaches. Therefore, attempts to use economic instruments will almost certainly need to be

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supported by some regulation of dive sites. This is the approach used, for example, by the Great Barrier Reef Marine Park Authority in allocating permits to tourism operators for the use of certain reef sites for day trips. An extension of the above approaches could be to supplement user charges with a regulatory requirement that charter operators charge this as a separate cost to divers, thereby making it clear that there are different prices which apply to different dive sites. Even on multiple dive trip holidays this may encourage some divers to not use certain sites. Again, this approach does not prevent open access use by individual divers, unless there is a very high level of policing of dive sites, which would be both difficult and expensive in most cases. Neither of the approaches just described deal with the problem of continually increasing demand for diving. While they may result in some reallocation of divers from one site to another in an optimal way, there is no guarantee that, ultimately, all sites will not be over-used. Is there another approach? If carrying capacities for individual dive sites could in fact be estimated, it may be possible to combine economic instruments and regulation in the following way. Begin by obtaining an estimate of the number of private divers. In most cases this will be a relatively small proportion of the total divers using a site. The difference between the carrying capacity and private diver numbers would be the “total allowable dives” available to charter operators at a site. These operators could purchase the rights for a certain number of dives at a particular site and pay an annual permit fee to retain those rights. They would be required to keep a log of the dives conducted at that site, with periodic checks by the managing authority. Presumably the funds realised from the sale of “dive site rights” could be used to pay for the managing agency’s policing and monitoring activities. Importantly, the rights should be made fully transferable between charter operators to ensure that they are allocated efficiently. This would have the added benefit that the “private property” nature of the rights would provide a strong incentive for the owners to protect the dive sites which are a capital asset and one which may appreciate in value if carefully managed. Dive operators would be expected to pass the costs they incur on to their diver clients, thereby internalising any external environmental costs. It is such a mixture of regulation and use of economic instruments which has the best chance of achieving an optimal allocation of divers amongst sites. Even then an equity problem arises in that this system gives preference to private divers. Ultimately, it may be important also to consider a permit system for this group. This last point serves also to highlight the boundary problems which arise in marine areas. Clarke et al. (1995), following Lipsey and Lancaster’s (1956) second-best theory, noted that user-pays is a useful policy “for a particular resource if it can also be applied to all other resources which could substitute for the competing resource” (p. 165). In marine areas, even protected areas, it will be difficult to include or exclude certain user groups. User-charges within a defined area may provide an incentive to use areas outside the boundary, possibly leading to overuse of those areas. Controlling entry to marine areas such as the Great Barrier Reef Marine Park will be both difficult to implement and monitor (Clarke et al., 1995). Clarke and his colleagues concluded that boundary problems cause both price and quantitative solutions to lose their effectiveness and even to become counter-productive in some instances. In the case of marine protected areas this will be influenced by the size of the area and the potential to access substitute sizes. It is likely to be less of a problem in the 3·49 million hectares of the Great Barrier Reef Marine Park than the 80 hectares of Julian Rocks Aquatic Reserve. It means also that education programs

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and the development of codes of conduct have an important role. Train-the-trainer (and operator) will be important in recreational diving, leading to better diver education and behaviour, as well as improved monitoring of that behaviour (cf. Clarke et al., 1995). Basing user rights on transferable permits, thereby granting an element of private property rights, should contribute also as dive operators then have a vested interest in preserving the environment from which they earn their income and which is also a capital asset. Considering user rights and diver behaviour in conjunction, one positive result may be diver briefings in which a strong emphasis is placed on care of the underwater environment. Harriott et al. (in press), in monitoring diver impacts, found a positive correlation between diver briefings and a lack of environmental damage. 10. Conclusion Increasing usage of MPAs for scuba diving, combined with concerns about overcrowding and biological damage at popular dive sites, implies a need for better site management. This may mean, amongst other things, that there is a need to seek an optimal distribution of divers and other users between sites. No one policy instrument will achieve such a result. Rather, it is likely that a judicious blend of regulation and economic instruments will be needed. For example, Geen and Lal (1991) suggested that charging users of the Great Barrier Reef may have a rationing effect, yet there may also be a need to place restrictions on very heavily used reefs. Geen and Lal suggested also that user charges should be related to the amount of damage being caused at particular locations by recreational users. Presumably, such an approach would have the further effect of inducing users to move to alternative, cheaper sites. In order to frame suitable management approaches, further information is required. For example, information on the factors underlying the demand for diving and dive sites, the demand elasticities associated with various sites and the degree of substitutability between sites is needed. Strategies need also to be devised which take care of boundary problems. Further, the important question of thresholds, raised by Dixon et al. (1993) requires investigation. This issue is complicated by there being two possible types of thresholds—biological and crowding thresholds. Long-term monitoring studies are required to gain further information about biological thresholds, while divers, as consumers, need to be studied in terms of their reactions to overcrowding and their willingness-to-pay for “wilderness” and other significant attributes of the diving experience. The research leading to this paper was partially funded under the National Ecotourism Program of the Australian Department of Tourism.

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