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Journal of Experimental Marine Biology and Ecology 250 (2000) 51–76 www.elsevier.nl / locate / jembe Experimental ecology of rocky intertidal habitats: what are we learning? A.J. Underwood* Centre for Research on Ecological Impacts of Coastal Cities, Marine Ecology Laboratories A11, University of Sydney, Sydney, NSW 2006, Australia Abstract Experimental analyses of causes of patterns of distribution and abundance of intertidal animals and plants on rocky shores have been a major activity for many years. In this review, some of the themes and topics that have emerged from such analyses are briefly discussed to provide an up-date for practitioners and ecologists working in other habitats. Conceptual issues include the widespread occurrence of transphyletic use of the same resources (space and food), theories and experimental analyses of intermediate disturbance in relation to numbers of species, the complex but pervasive nature of indirect interactions among species, relative importance of ‘top-down’ versus ‘bottom-up’ control of assemblages and the importance to rocky intertidal species of ‘supply-side’ influences on densities and interactions. Methodological advances include experimental designs for complex and patchy, interacting sets of species, the importance of controls in experimental manipulations and methods for analyses of hierarchical scales of patterns and processes. Finally, some contributions to social issues (pollution, biodiversity) and some scenarios for future directions are briefly considered. 2000 Elsevier Science B.V. All rights reserved. Keywords: Ecology; Experiment; Methods; Rocky intertidal habitats 1. Introduction 1.1. Preamble The ecology of animals and plants on intertidal rocky shores has been a topic of interest for decades in many parts of the world. There have long been descriptions of *Tel.: 1 61-2-9351-2590; fax: 1 61-2-9351-6713. E-mail address: [email protected] (A.J. Underwood). 0022-0981 / 00 / $ – see front matter 2000 Elsevier Science B.V. All rights reserved. PII: S0022-0981( 00 )00179-9 52 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 fauna and flora (Colman, 1933; Fischer-Piette, 1936) and some manipulative experimentation goes back at least to the 1930s (Hatton, 1932). The ecological study of the organisms has been, in many ways, a source of, or, at least, a major contribution to concepts that have moved out into other areas of ecology. There have been several syntheses of the field, from the purely descriptive cataloguing of broad patterns of occupancy of shores in different parts of the world (e.g. Lewis, 1964; Morton and Miller, 1968; Stephenson and Stephenson, 1972) to those based on experimental analyses of patterns of distribution (Connell, 1972; Paine, 1977). It is often timely to revisit some of the themes and constructs of a particular discipline or sub-discipline in order to take stock of the current frameworks and local geography. Such an exercise can be self-serving — it is, in fact, intended to be in this case — as an idiosyncratic assessment of where we, the practitioners in or aficionados of a field might consider ourselves to be. It may, however, also serve to provide an up-date for those in other areas of ecology, to explain what and where current research in intertidal ecology is at present. This special edition of Journal of Experimental Marine Biology and Ecology provides an opportunity for such a brief overview. The topic is covered as a voyage of discovery about the themes intertidal ecologists are learning (hence its title) and the essay will have been worthwhile if that is all it achieves. On the other hand, if it also helps explain current activity to scientists elsewhere, it will have achieved other end-points. 1.2. The rise of experimentation One of the key factors of the past 30 years of intertidal ecology has been the rise of experimental manipulations as a crucial investigative tool. This has been quite phenomenal and represents a major shift of emphasis that has occurred during the same period as the massive increase in publications of science following the expansion of universities and grant-funding in the 1960s (Underwood, 1996a). It is worth a small amount of space on a reprise of why rocky intertidal habitats have been at the forefront of development of ecological experimentation. Most notably, a rocky intertidal shore encompasses a gradient of environmental conditions from fully marine below low tidal levels to fully terrestrial where splash and spray reach to the highest levels above high tide. Mostly, although not always, the gradient occurs over a small (metres to tens of metres) distance making some patterns of response to environmental variables relatively easy to see (Newell, 1976). Short distances across gradients also allow relatively easy observation and manipulation of environmental variables (e.g. remote sensors in different parts of the range are still quite close to each other). On such gradients, there is usually a great variety of animals and plants, often having several representatives in functional groups or guilds (Menge et al., 1986), e.g. grazing snails, predatory whelks, crabs, large fleshy, brown seaweeds, etc. Such diversity facilitates comparative studies, often allowing the generality of processes to be assessed by testing similar or the same hypotheses on several similar species. At the same time, many of the animals and plants are macroscopic, abundant, slow-moving or sessile as adults (Connell, 1972) and interact at small spatial scales A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 53 (Underwood and Chapman, 1996). Thus, small-scale studies are often at the appropriate scales, and the mechanics and logistics are usually manageable. Animals and plants may live for extraordinarily long periods, but typically turn over in relatively few years. They therefore complete their life-cycles in the sort of time-scale that is matched by cycles of grant-funding and scholarships for Ph.D.s. This has the serious down-side that studies have focussed on the conspicuous, abundant, slow-moving and shorter-lived components of assemblages. This has ignored many of the long-lived plants (Slocum, 1980) and the more active, small consumers (e.g. Brawley, 1992), although there have been exceptions (e.g. Johnson and Mann, 1988; Duffy and Hay, 1991). It is also the case that the vast majority of studies have been focussed on species characterized by relative immobility. The immobility has been emphasized by ignoring the potentially great distances travelled (Scheltema, 1971) by the dispersive larval phases of life-history, although there have been exceptions such as Shanks and Wright (1987), Shanks (1995), Eckman (1996) and Shkedy and Roughgarden (1997). Despite such biases, animals and plants on rocky shores are quite suited by visibility, size, diversity, longevity, abundance and lack of emotional appeal to be subjects of experimental tests of hypotheses about ecological patterns and processes. So, where are we now and what are we discovering from such experimentation? 2. Some conceptual issues 2.1. Trans-phyletic analyses of use of resources One area that has been extensively exploited in experimental analyses of processes causing and maintaining ecological patterns in intertidal habitats has been the role of competition for resources of food and space (reviews by Branch, 1984; Underwood, 1986a, 1992a). Two-dimensional space is a resource required by almost all intertidal species, either directly as an absolute need (Andrewartha and Birch, 1954) for space on which to settle and grow or as a relative need for space over which to feed. As a result of the fixed (because of geographical dimensions) availability of the total amount of two-dimensional space available at any location and the varying and often unpredictable nature of abundances and mixtures of species occupying the space, competition is often intense and sustained. So, as examples, competition among and within species of limpets for food is a widespread and normal aspect of the ecology of limpets on South African shores (Branch, 1981, 1984), competition for barnacles is a well-described feature of ecology of shores in Britain (Connell, 1961) and elsewhere (Wethey, 1984a). An interesting phenomenon of ecology of rocky intertidal habitats is the great and intriguing complexity of interactions among similar sorts of organisms. For example, Kastendiek (1982) described an interesting situation where the turfing red alga, Halidrys dioica, outcompetes the alga, Pterocladia capillacea, by growing over and denying it access to light. In the presence of the canopy-forming species, Eisenia arborea, however, the inferior competitor survives well under the canopy and is able to ‘resist’ 54 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 competition from H. dioica. Thus, competition between E. arborea and H. dioica prevents overgrowth of P. capillacea by H. dioica. Equally interesting and well-documented from many experimental studies is the widespread occurrence of competitive interactions between very different sorts of organisms. Without entering the labyrinthine maze of numerous types and complex taxonomy of competitive interactions (Schoener, 1983), competition on rocky shores is of three general types. Pre-emptive competition occurs wherever occupation of space by one species prevents another species from settling from the plankton (which may allow the later arrival to find space elsewhere; Underwood and Denley, 1984). Interference competition is that involving a ‘contest’ (e.g. Pielou, 1974), so that one user of space directly harms or kills another. Examples are a faster-growing barnacle that undercuts or smothers a slower-growing species (Connell, 1961) or mussels smothering other species by growing over them (Paine, 1974; Menge, 1976). Finally, there is exploitative or ‘scramble’ competition (Pielou, 1974) where several species need the same space to feed, but there is insufficient food to support all the animals needing it (e.g. Underwood, 1984). Of these types, pre-emption and interference are often across phyla (or between animals and plants). Pre-emption prevents barnacles from settling where algal fronds sweep the surfaces of the rock (Dayton, 1971) and where the cover of plants prevents settlement or attachment of larvae. Direct interference occurs where, for example, mussels encroach on the feeding territories of limpets (Stimson, 1970, 1973) or overgrow and kill barnacles (e.g. Menge, 1976; Jernakoff, 1985). Pre-emptive competition is fundamentally different from the other two types in that the outcome may not actually be any increased damage or increased risk of mortality to the ‘loser’ (members of the species arriving later). The larvae may simply go elsewhere, although it is possible that delaying settlement may lead to increased risk of predation or accidental calamity. The hypothesis that larvae prevented from settling in one spot are more likely to die before settlement than are those not so prevented has not been tested (and will be very difficult to test in the field). Competitive interactions between different sorts of species demonstrate the need for careful identification of the make-up of ‘guilds’ of species using similar resources (Root, 1967). Taxocoenes (groups of similar types of species) are often considered a ‘unit’ of study in assemblages (see discussion in Underwood, 1986b), but are not a relevant grouping where resources are used by many organisms that are not taxonomically related. 2.2. Intermediate disturbance and competitive interactions One consequence of widespread, complex competitive interactions is that any other process (disturbance, disease, predation) leading to reductions in densities or cover of competitors can have indirect effects on species not directly involved. For example, the model of intermediate disturbance was proposed, historically, by Tansley and Adamson (1925; see Jackson, 1981) and, in a more recent context, by Paine and Vadas (1969) and Connell (1978) to explain the often observed downwards concave curve of number of species across a gradient of disturbance. The model states that where disturbance is large A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 55 (or frequent or recent), only those species capable of withstanding it (or colonizing and growing since the last disturbance) can survive. This is only a subset of the species that would otherwise be found in the habitat. At the other end of the gradient, where disturbances are small (or rare or long ago), superior competitors have time to build up sizes or numbers and to dominate resources. Consequently, the number of species is, again, reduced. Hypotheses derived from this model must be tested by manipulations of the regime of disturbance. No amount of describing patterns of species richness across gradients of disturbance will help. Two very convincing series of experiments have been done in rocky intertidal habitats to test predictions about changes and the processes causing changes in numbers of species when disturbances are manipulated. Sousa (1979a,b, 1980) increased disturbance in a boulder-field by turning boulders over more frequently than occurred naturally. This did lead to reductions in abundance of some species. Reducing disturbance (by experimentally preventing boulders from being turned over) did increase elimination of green species by red algal species. The major conclusions were, however, that features of life-history (rate of colonization, nature of cycle of breeding) and responses to disturbance were important reasons why the model of intermediate disturbance did not apply very well to explain the observed patterns of numbers of species. Responses of long-lived red algae to being disturbed included ‘grab-and-hold’ strategies whereby vegetative growth from surviving remnants retained occupation of space by those species. In the end, in areas that were not much disturbed or not disturbed often, the perennial red species became dominant by occupying space vacated by other species when it became available, rather than by overgrowing and eliminating other species competitively. McGuinness (1987a,b) examined the consequences of experimentally increasing or decreasing disturbance in several boulder-fields (at two heights on each of two shores). His results found support for the model of intermediate disturbance for only some combinations of components of the fauna, under only some conditions. Intermediate disturbance did not seem a widespread explanation for patterns of difference in numbers of species across gradients. 2.3. Keystone predation and other indirect interactions Other indirect effects have, however, been more widely demonstrated (Dayton, 1971; Lubchenco, 1978). The most widely cited is ‘keystone predation’ (Paine, 1966, 1974), the situation where a predator can cause a large change in local diversity or relative abundances of species because it consumes superior competitors in an assemblage. The best-known example is the starfish, Pisaster ochraceus, which eats mussels as a major component of its food. The mussels are capable of smothering many, if not all, of the other users of primary space on the shore. So, the predators, by removing mussels, make space continuously available for other species and thereby increase diversity. There have been critical evaluations of this example in terms of the evidence (Fairweather and Underwood, 1983; Underwood and Denley, 1984) and the extent to which the process is widespread (e.g. Foster, 1990). It is also clear that areas with mussels generally support 56 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 more rather than fewer species than are found where mussels are removed (Lohse, 1993), because more species find habitat on and amongst mussels than on the rock itself. Concepts of keystone predation have spread in marine ecological studies (Mann and Breen, 1972; Mann, 1982; Estes and Duggins, 1995), with continuing criticism of their validity and the lack of attention to other explanatory models for observed patterns of numbers of species (Foster and Schiel, 1988; Elner and Vadas, 1990). There has also been some uncritical adoption of the concept in areas far removed from its empirical origins, such as in discussions of issues for biological conservation (Soule´ and Simberloff, 1986; Terborgh, 1986). Here, again, the validity of the concept, or the untested applicability of the concept to new situations has been seriously questioned (Mills et al., 1993). Whatever the validity or applicability of keystone predation in any particular situation, there has been a renewed interest in and understanding of indirect interactions and their importance in the ecology of complex assemblages (reviewed by Menge, 1997). This was, of course, an older tradition dating back to Darwin’s ‘‘web of complex ‘interactions’ ’’ (Darwin, 1859). It generated some attempt at a novel theoretical synthesis of components of an organism’s environment (Andrewartha and Birch, 1984). Certain elements (mates, food, predators) were defined by Andrewartha and Birch (1984) to be in the core or ‘centrum’ of an animal’s environment. Other ecological components (e.g. competitors) were considered to be in the ‘web’ of indirect influences on the abundance of an animal. The most recent synthesis has been by Wootton (1993, 1994a) who has demonstrated that indirect interactions fall into two main types, i.e., ‘interaction chains’ and ‘interaction modifications’. In the first case, a species (A) has direct effects on a second species (B), which, in turn, has direct influences on a third species (C). For example, a predatory whelk (A) consumes barnacles (B) that occupy space, making it unsuitable for grazing limpets (C) (see Dayton, 1971; Underwood et al., 1983, for examples). As a result, consumption of barnacles by the whelks can lead to local increases in numbers of limpets because of the reduction in competition for space. Predation directly negatively influences numbers of barnacles, but indirectly positively affects numbers of limpets. In the second case (an interaction modification), an indirectly acting species influences the direct interaction between two species. So, a predatory whelk (A) consumes barnacles (C), but is itself eaten by predatory crabs (B) (see examples in Hughes and Elner, 1979; Hughes and Seed, 1995). Thus, the direct reduction of numbers of whelks due to the activities of crabs may indirectly cause an increase in the number of barnacles. Analyses of these interactions can require very carefully formulated hypotheses and the appropriate manipulative experiments to test them. Methods used include path analysis (Wootton, 1994b), but such approaches are saddled with all the problems of any derivative of techniques of multiple regression (Petraitis et al., 1996). Also, there can be problems with attempting to fit particular cases into a theoretical dichotomous framework. For example, Underwood (1999a) demonstrated experimentally that whelks (Morula marginalba) shelter under the canopy of an alga (Hormosira banksii), creating reductions in densities of a prey species, the barnacle Chamaesipho tasmanica (see also related studies by Fairweather et al., 1984; Moran, 1985; Fairweather, 1988). A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 57 So, the alga directly increases the abundance of the whelk; the whelk directly decreases the numbers of the barnacle. Removal of the algal canopy causes local increases in abundances of the barnacle, an indirect effect of the first type identified as an interaction chain by Wootton (1993). At the same time, however, the rate and intensity of the direct predatory actions of the whelks (i.e. reducing the numbers of barnacles) are influenced by the size of the canopy, so the canopy has an indirect effect of the second type identified by Wootton (1993). Despite potential problems with analyses and interpretations, complex chains of indirect interactions are an important area for study, particularly because they will prove crucial for effective understanding of functional aspects of biodiversity in marine habitats. 2.4. ‘ Top-down /bottom-up’ regulation of assemblages A well-developed framework for understanding assemblages in some habitats is the idea that structure of assemblages can be regulated by ‘bottom-up’ processes. In such processes, there may be quantitative or qualitative differences in the structure of assemblages with different levels of nutrients in the system. This has been a feature of some interpretations of factors controlling structure and composition of assemblages in freshwater habitats (e.g. Hall et al., 1970; Power, 1990). The argument is that, where primary production is greater, there can be greater abundances and / or greater diversity of grazers exploiting the large primary production. This idea contrasts with well-established notions in intertidal ecology that are interpreted as ‘top-down’ control. So, for example, predatory animals may consume sufficient grazers in any area, thereby preventing excessive numbers in or, sometimes, eliminating species from patches of habitat. Alternatively, competitive interactions may directly limit the numbers of all species that compete for a particular resource. The degree to which either is the major influence on structure of an assemblage is emerging as an important issue in the analysis of complex ecological systems (Fretwell, 1987; Menge, 1997; reviewed by Menge, 2000). Determining how important either type of process may be under different circumstances can be advanced by the experimental opportunities offered by experiments on rocky shores. There are potential managerial or conservatory issues associated with this concept. For example, where top-down processes matter, reductions in density or removal of predators or large, competitively dominant grazers or users of space can have profound impacts on other components of the assemblage (Paine and Vadas, 1969; Moreno et al., 1984; Castilla and Duran, 1985; Castilla and Bustamante, 1989). Species that have abundances regulated by predators can explode in numbers with concomitant alterations in density-dependent processes influencing other species. So, management for conservation needs to be concerned with harvesting, fishing and any disruptive processes that may alter relative abundances of the top-down, regulating species. Where bottom-up processes are more important, the managerial issues must revolve around preventing alterations to productivity, in particular, to aspects of eutrophication and other manifestations of overabundant nutrients. Menge (1992, 2000) perceptively pointed out that no ecological system is likely to be controlled solely in one or other direction. There is also the well-known problem that primary production may, itself, regulate the number of trophic levels in an assemblage 58 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 (Hall et al., 1970; Fretwell, 1987). In such a case, the increased trophic complexity may allow for increased types of predators and increased variety, interactions and strengths of predatory or other top-down controlling processes. Were this to be the case, it becomes philosophically unclear how top-down processes could control an assemblage, given that the diversity of higher trophic levels, where many of the top-down controlling species are found, is itself controlled by bottom-up processes. Nevertheless, more study is needed of the relationships between and synthesis of regulatory processes operating in opposite directions in assemblages. 2.5. Supply-side ecology Another area of intertidal investigation that has been of some influence in the development of conceptual methods is the notion that supply of recruits into any patch of habitat is an important influence. This is, in no way, a new idea, but it was unfashionable for a while and had to wait its turn to be ‘rediscovered’ (Young, 1987; Underwood and Fairweather, 1989). The term ‘supply-side ecology’ was coined by Lewin (1986) to summarize, in a pun, the fact that various important processes can only occur or can only take place at relevant magnitudes and rates if the species involved in them are present in sufficient numbers. As an example, predatory starfish or whelks cannot be involved in top-down control of the structure of an assemblage if they have not arrived in the assemblage as larvae or, having arrived, fail to survive to sizes large enough to become dominant predators. Where there is great variation in the numbers of larvae arriving from time to time and / or place to place, there will be great variation in the duration, timing or frequency of any particular process. Sometimes, as explained for a variety of examples by Underwood (1979) and Underwood and Denley (1984), prevailing explanations of observed patterns in intertidal assemblages fail to include the possibility that larval supply or recruitment of juveniles could be important. So, for one example, Connell (1975) described a conceptual model for recolonization of a disturbed patch of habitat. In relatively benign parts of the environment, predators were presumed to be able to eliminate most, if not all, of prey species under most prevailing patterns of weather. Occasionally, however, the predators are absent (e.g. whelks are missing because they were killed by an unusual period of harsh weather; Dayton, 1971). Consequently, in those periods, prey arrive as larvae, settle and survive. If they then survive long enough to become sufficiently large, they will escape being consumed by predators when the predators finally re-appear. This model can explain why there are intermittent ‘pulses’ of appearance of species of prey in some intertidal habitats. The alternative, supply-side explanation (Underwood and Denley, 1984) is that predators fluctuate in abundance because of variations in their own recruitment and not because of occasional periods of harsh weather. Regardless of the causes or amounts of such variation, numbers of prey fluctuate a lot because of the vagaries of larval production, dispersal and survival. Occasionally, they will arrive in large numbers when numbers of predators recruiting at some earlier time happen to be small. Alternatively, they may arrive in very large numbers, so that they ‘swamp’ their predators. Despite A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 59 intense predation, sufficient prey survive long enough to reach sizes at which predators can no longer harm them. Under this explanatory model, it is not unusual weather events that reduce numbers of predators. Instead, larger-than-normal recruitment of prey dictates temporal variation in the presence or in the numbers of a prey species present, from time to time in an area. The concept that larval variation drives abundances of adults of marine species has been widely understood in fisheries science. In fact (apart from being understood with respect to agriculture in biblical times; King James Bible), the earliest reference on the topic known to me was Hjort (1914) discussing fisheries. There have been many observations of major variations in abundances of intertidal or shallow coastal species (Coe, 1956; Loosanoff, 1964, 1966) and, more than 50 years ago, there were reviews of the consequences (e.g. Orton, 1937). Thorson (1946, 1950) was the first marine ecologist to try to use variation in recruitment as a mechanism in models explaining temporal and spatial variations in abundances of animals. He noted that three species of bivalves with long larval periods of planktonic development had abundances of adults that fluctuated much more than was the case for three species living in the same habitats, but which had a short period or no pelagic development. He also commented on the consequences of the timing of recruitment. When the larvae of a bivalve arrived before the larvae of one of their predators (a starfish), they were able to survive for long enough to be too large to be consumed by the predators when these eventually arrived. In contrast, if the bivalves recruited after the starfish, many more were consumed. So, timing of recruitment could also influence the sizes of populations of adults. The notion of supply-side ecology has become more important in recent developments of meta-models of patchy populations of marine invertebrates (Underwood and Fairweather, 1989). The best developments have been the models developed by Roughgarden (e.g. Roughgarden et al., 1985). She developed a model for space-limited habitats that linked the numbers of barnacles settling on the shore to the amount of free space available for settlement on that shore and to the probability of recruits surviving from one time to another. The model was highly successful in some situations. For example, where competitors for space occupy areas that have generally poor recruitment, the competitors will have abundances of adults regulated by processes post-recruitment. In contrast, where recruitment is generally great, sizes of populations tend to be regulated by densities and fluctuations of recruits. These results conform well to some empirical observations (see particularly Connell, 1985; Hughes, 1990; Sutherland, 1990). The models are less successful at providing useful insights or predictive capacity for species that have variable rates of recruitment from time to time (e.g. barnacles in New South Wales discussed in Underwood and Denley, 1984; Underwood, 1999b). Nevertheless, more recent advances have begun to demystify the vagaries of what Spight (1974) called the planktonic mystery stage (see review by Eckman, 1996). As a particular example, the consequences of variations in Roughgarden et al.’s (1985) ‘recruitment parameter’ (the number of cyprids settling within a period of time per unit area of free space that survive to the end of that period) can be predicted for some species of barnacles on the coast of California. The numbers likely to be available for settlement can be very satisfactorily predicted from a polynomial regression of numbers 60 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 of recruits on sea-surface temperature (to the 11 th power; Shkedy and Roughgarden, 1997). The regression was associated with 59% of the variation in numbers of recruits, which is a remarkably good fit for field-derived empirical data. So, in those systems that conform persistently to one or other end of the gradient of magnitude of recruitment, supply-side models and coupling with upwelling and other coastal oceanographic processes can provide considerable understanding of and capacity to predict numbers of adults in intertidal populations. 3. Some methodological issues in experimental design As with many areas of modern ecology, intertidal ecology has been concerned with improvements to its methods, particularly those concerned with the design, analysis and interpretation of quantitative and experimental data. Some of these contributions are briefly described here. 3.1. Experimental designs: competition The formal analysis of competitive interactions has been fraught with difficulties of understanding the relevant processes and scales (Connell, 1983; Schoener, 1983) and of logical structures in the design of the experiments (reviewed by Underwood, 1986a, 1992a). There is no need to repeat the issues for design, but the main points are worth iterating once more because they seem to have eluded some authors. If it is proposed that two species (A and B) have negative effects on each other because of their joint needs for some resources (Birch, 1957), there are two major procedures. First, the amount of resource may be manipulated. For example, the amount of food in areas can be experimentally increased (or decreased) to test the hypothesis that outcomes of competition will be less (or more) stark. Second, the numbers of consumers of resources can be manipulated, again to test hypotheses about directions and magnitudes of interactions among the consumers. In the simplest experiments, to examine the influence of species B on species A, some relevant density of A must be created in the absence of B. As the second experimental treatment, the same density of A must be established with a relevant density of B. Thus, the minimal experiment has independently replicated arenas with density NA of A and independently replicated areas with NA plus density MB of B. To determine the reciprocal influence of species A on species B, there must also be independently replicated arenas with MB of species B alone and MB of species B with NA of species A. In some studies, the treatments with the two species together can be the same arenas (using results for species A and results of species B in analyses; Underwood, 1978a, 1984, 1986a, 1992a; Creese and Underwood, 1982). Sometimes, because of issues of non-independence of data, it will be more appropriate to establish two sets of replicate arenas for the combined treatment and use one set to provide data for species A and the other set to provide data for species B. Attempts to do such experiments by holding the total density of organisms constant are confounded. So, some authors investigate the effect of B on A with N individuals of species A in one treatment and a total of N A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 61 summed from species A and B in the other treatment (see Underwood, 1986a). The latter alters the mix of species (as required by the hypothesis) and simultaneously alters the density of species A. Any comparison between the treatments cannot test the hypothesis as stated (see detailed discussion in Underwood, 1986a, 1992a). It is often the case by hypothesis and wherever per capita influences of competitive interactions must be determined, that the influences of each species must be investigated at several densities. Thus, there must now be different densities of species A (NA 1 , NA 2 , etc.), as required. Each of these must also be established at the appropriate densities of species B. This creates a two-factorial matrix with (NA 1 , NA 1 1 MB 1 , NA 1 1 MB 2 , . . . . . . . . . , NA 2 , NA 2 1 MB 1 , NA 2 1 MB 2 , . . . . . . . . . , etc.) as treatments. The analysis of this sort of experiment remains straightforward and can be extended to make simultaneous comparisons of the influence of several potential pairwise interactions. The experiment becomes a little more complex where possible asymmetries (Lawton and Hassell, 1981; Connell, 1983; Schoener, 1983) in the intensity of competition between two or more species must also be investigated. For this to be possible, an experiment must simultaneously include treatments with the same additions as before, but this time of the same species (A added to A; B added to B). For the simplest case, of one density of species A (NA ) and one density of species B (MB ), there must be treatments NA , NA 1 MB ) as before, plus (NA 1 MA ) to determine the magnitude of intraspecific competition, i.e., species A on species A, relative to that of interspecific competition, i.e., species B on species A. Then, there must also be (MB , MB 1 NA , MB 1 NB ) to determine the influences on species B. If the asymmetry is to be compared between the two species, in addition to measuring it for each species separately, there should really also be treatments (MA ; MA 1 NA , MA 1 NB ; NB ; NB 1 MA , NB 1 MB ) to ensure that a comparison of per capita influences does not confound inter- and intraspecific differences with differences in the density of the two species. These designs and their interpretation have become quite standard in studies of competition on rocky shores. They are well-suited to the manipulation of densities and composition of grazing species, particularly where the species are abundant, so that experimental plots are small and there are many individuals to make up the experimental densities. 3.2. Experimental designs: transplantation Another quite common requirement of studies of distributions of organisms across gradients is to be able to transplant individuals from one part of the gradient to another to test specific types of hypotheses (reviewed by Chapman, 1986, 1999; Underwood, 1988; Chapman and Underwood, 1992). For example, suppose there is a gradient in size of mobile animals with smaller ones at higher levels on the shore. In some areas, an appropriate model to explain the observed pattern is that small animals are more likely (for whatever reason) to move upshore than are large individuals. Any large individual wandering at random downshore will tend to stay there, whereas a small animal will move upwards. Reasons for such behaviour include perception of an increased risk of 62 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 predation, even though more food is available at lower levels (e.g. Paine, 1969). Larger animals can move to and survive at lower levels because they are less likely to be eaten by a predator than are the small individuals. Alternatively, in some cases, more food may be available to competitively inferior small individuals at higher levels on the shore (e.g. Wolcott, 1973). So, small individuals keep moving until they find food at the higher levels. There are, of course, other models, but these examples will illustrate the point. From this model to explain the gradient in size, an appropriate hypothesis is that if some large and some small individuals are transplanted downshore, small individuals will change their behaviour and will show greater tendencies to move upshore than do large animals, and shown by small ones in the original habitat. To test the above hypothesis requires appropriate controls for disturbing the animals whilst moving them and for moving them to a new habitat, regardless of it being at a lower level. Suppose, for example, that individuals put in a new location, wherever it is, are seriously dismayed by unfamiliar surroundings and / or by unfamiliar individuals in those surroundings. As a result, they become disorientated and tend to move more often and to move greater distances upshore, as a response to such disorientation. Under these circumstances, discovering a greater tendency to move upshore does not unambiguously support the hypothesis. Such a conclusion is potentially confounded with any effects of disorientation. Appropriate controls must include translocations, i.e., the disturbance of animals that are moved to a new location in the same (upper) part of the gradient where they were originally found (Chapman, 1986; Chapman and Underwood, 1992). Where such controls have been properly incorporated in experimental designs, they have often revealed the existence of the potential artefacts (Underwood, 1988; Chapman, 1999). The roles, needs and natures of the appropriate controls in such ecological experiments (wherever they are to be done) have been greatly elucidated by experimentation on rocky shores. 3.3. Variation in processes and hierarchies in patterns Ecologists studying rocky intertidal habitats have been very concerned with spatial and temporal variability in the patterns and processes that influence distributions and abundances of animals and plants. Early attempts to fit simple models of zonation (e.g. Colman, 1933; Lewis, 1964), i.e., the replacement of one sub-assemblage by another in discrete and abrupt boundaries between low and high tide, persist in the literature. They have, however, never been supported by quantitative data and have been refuted by quantitative tests of their predictions (Underwood, 1978b; Chaloupka and Hall, 1985). They are as inaccurate a description of the distributions of species across intertidal gradients as were descriptions of series of communities of plants at different heights on the Smoky Mountains. The latter were demolished by quantitative sampling by Whittaker (1956). Instead, there has more recently been a focus on patch dynamics and attempts to understand and model responses to mixtures of disturbances, physical factors and variable rates and intensities of competition and predation (see particularly the mix of theory and experimentation by Levin and Paine (1974) and Paine and Levin (1981)). All of these interacting ecological processes are affected by issues of recruitment (see A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 63 particularly the experimental work of Dayton, 1971; Menge, 1976; Sousa, 1979a,b, 1980b; Underwood et al., 1983). In some parts of the world, however, this is being replaced by an increasing interest in processes operating at different spatial and temporal scales as hierarchies, rather than as interactions at one place or time of investigation. Some of the background to considering ecological scales in hierarchies was summarized by Allen and Starr (1982). Examples of the sort of ecological models that involve hierarchies of processes operating at different scales were provided for subtidal kelp-beds in studies by Dayton and Tegner (1984). In analyses of intertidal habitats, there have been several approaches to considering hierarchies of spatial scales. One was the survey done by Foster (1990). Along the coast of California, Foster examined a series of typical rocky headlands chosen because they had similar physical characteristics. On each shore, he examined a series of patterns to determine how widespread or how frequent they were. The object of the exercise was to test predictions derived from models about processes influencing the local structure of assemblages. So, predation on superior competitive mussels had been proposed as a widespread and important influence on distributions (Paine, 1974; see earlier discussion of keystone predation). Similarly, Foster’s (1982) own work on competition for space between algae was thought to be an important process. If any of the processes considered was, in fact, important, Foster (1990) hypothesized that the patterns resulting from the processes should be found frequently over a set of shores for which the processes were claimed to be operating. This hypothesis was not generally supported by the data. Despite objections to this approach (Paine, 1991), it has great merit. If the patterns that are supposed to be the result of some process are not widespread, it is difficult to maintain an argument that the process occurs widely. It is not ‘nihilist’ (Paine, 1991) to question dogma by testing hypotheses about outcomes of supposedly general processes. When predictions fail, new models and understanding are needed (Popper, 1968; Simberloff, 1983; Underwood, 1990). There must usually be a delay between discovering that some previous paradigm must be overthrown because it has failed and proposing new models that incorporate the older ideas and the new observations that failed to confirm them (Kuhn, 1970). So, this approach examines the frequency of patterns that should result from various processes. An alternative, sometimes called a comparative experimental approach (Menge et al., 1994), uses experimental procedures, done at small scales, but arranged across larger spatial scales. This was used successfully to identify the variable responses to removals of predatory starfish along a coast-line (discussed earlier) and to compare this interaction across coastlines (Paine et al., 1985). It has also served well to demonstrate the inconsistencies in colonization and development of algal assemblages on low-shore rocky habitats on the exposed coast-line of New South Wales (Chapman and Underwood, 1998). It was essential for Wethey’s (1984b) analysis of short-term variation in settlement of barnacles on British shores and Caffey’s (1982) experimental tests of hypotheses about the influences of different types of rock on the settlement of barnacles. This approach has, however, not yet managed to synthesize results of some complex intertidal issues, for example, the timing, frequency and duration of foraging by intertidal homing limpets on British shores (Hartnoll and Wright, 1977; Little et al., 1990; Gray and Naylor, 1996). 64 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 One of the problems with this comparative approach is that it is better to plan the experiments to make specific contrasts interpretable, i.e., to propose and test specific hypotheses about different types of habitats (or weathers, or seasons, etc.). This has numerous advantages over the attempt to gather together the outcomes of various experiments on similar topics, but done for different purposes without directly comparable designs and without the appropriate spatial and temporal replication (see the discussion in Underwood and Petraitis, 1993). So, planned comparative experimentation can reveal not only the relevant variation in processes at large spatial scales (for example, variations in diets and growth of predators on shores dominated by different types of prey; Moran et al., 1984), but also the similarities in shores of similar type compared to the differences from one habitat to another (for example, the influences of experimental removals of predatory whelks from two different intertidal habitats; Fairweather and Underwood, 1991). Doing experiments over short periods, but repeating them in a planned manner over several time-periods had similar advantages for understanding temporal variation (Underwood and Chapman, 1992). The third and final approach to investigating hierarchies in ecological processes and their resulting patterns has been the analysis of the spatial (or temporal) hierarchy itself. Some examples showing methods of analysis and results for spatial variation in abundances of intertidal snails and barnacles have been described in Underwood (1996b), and Underwood and Chapman (1996, 1998). In a comparison of relevant procedures (spatial autocorrelation, fractal analysis, block mean square analysis and hierarchical analysis of variance), Underwood and Chapman (1996) found that the traditional hierarchical analyses had considerable advantages. So, for sessile species and species with limited mobility that are typical of denizens on rocky shores, investigations to test hypotheses at scales from tens of kilometres down to centimetres can be done by sets of experimental or sampling units at small spatial intervals repeated at sites, locations, etc., that are different distances apart. The other procedures investigated all required much more effort, were very time-consuming to replicate and impossible to do over very large spatial scales. The second important result from these types of analyses is that specific hypotheses about a hierarchical series of processes operating simultaneously can be tested in comparison with each other. So, very small-scale (centimetres to metres) variation in densities of an intertidal snail can be shown to be the result of small-scale behavioural responses by individuals to local topography, food, micro-climate. Larger-scale (metres to tens of metres) variability can be attributed to variation in biological processes of pre-emption, interference, predation. At yet larger scales (tens to hundreds of metres), variation may largely be affected by variations in recruitment or physical disturbances due to weather. At even larger scales of hundreds of metres to kilometres, there may be consistent variation due to wave-action and storms. Finally, at very large spatial scales (tens to hundreds of kilometres), there may be biogeographic variation caused by consistent latitudinal differences in climate. Hierarchical analysis of variance is a robust tool for extracting estimates of variance from data collected at these scales, so that comparisons can be made about the relative magnitudes of such variance (e.g. Burdick and Graybill, 1992; Searle et al., 1992; A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 65 Underwood, 1997). Where this has been used for intertidal species (Underwood, 1996b; Underwood and Chapman, 1996), the outcome has almost universally been that variation is very great at the smallest spatial scales. This not only implies a very great importance for small-scale interactions of behaviour and ecology in response to food, micro-climate and topography. It also, fortunately, justifies assumptions of independence, thus validating experimental manipulations where replicate experimental units have been separated by a few metres (see also Underwood, 1998a). 3.4. Contributions to social uses of ecology Ecological methods used routinely on rocky coasts have also been adapted to help solve various problems of a practical, environmental nature. Three examples will illustrate the point. First, there have been developments of the asymmetrical sampling designs needed for detecting and estimating the sizes of environmental impacts (reviewed by Underwood, 1994). It has long been realized that impacts can only be defined and detected as a statistical interaction in time and space (Green, 1979). There must be a different pattern of change in some relevant variable(s) from before to after a human disturbance in the disturbed site compared to undisturbed, reference areas. Routinely, such interactions have been detected using a comparison of the disturbed to a single undisturbed site (BACI procedures; Bernstein and Zalinski, 1983; Stewart-Oaten et al., 1986). These procedures are unreplicated, so the comparison is always potentially confounded. ‘Beyond BACI’ procedures compare the site disturbed with a sample of undisturbed sites (Underwood, 1992b, 1993, 1994) and arose from the asymmetrical analytical procedures used to analyse competitive interactions (Underwood, 1978a, 1984). These procedures have been extended to situations where there are no data before the disturbance (Glasby, 1997). A more recent approach to this latter problem involves a meta-analysis of a series of paired comparisons, each of one disturbed and a paired undisturbed site (McDonald et al., 1993). Any interpretation of the result of each comparison would be confounded (as above), but the whole set of comparisons provides independent replication of the tests, providing an unconfounded interpretation. In the case of oil-spills, for which these methods were originally used, there may be problems with finding independent sets of oiled and unoiled sites that did not originally differ (and therefore subsequently interact) in some important way that has nothing to do with oil-spills (Underwood, 1999c). Nevertheless, these are promising methods in the detection of impacts. The third example is the recent development (Underwood and Chapman, 1998) of univariate analyses of measures of composition and relative abundance of species in intertidal assemblages to be able to use the previously mentioned hierarchical analyses of spatial variation. The methods generate independent measures of multivariate differences among replicates in samples at different scales. These can then be analysed by the versatile procedures available for univariate measures (Winer et al., 1991; Underwood, 1997). Such techniques may be helpful in assessments of scales of variation of ecological diversity and for such problems as divisions of coast-lines, habitats, regions, etc., for conservation, management, establishing marine reserves, etc. The development of analytical and experimental methodologies capable of dealing 66 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 with the very variable abundances of populations, patchiness of occupation of habitat, rates and intensities of processes influencing densities and sizes and the composition of assemblages in intertidal habitats is an on-going research programme in intertidal ecology. It is not surprising that some of the outcomes spill into areas of applied problem-solving. 4. Conclusions The examples discussed briefly above serve to illustrate three themes in experimental intertidal ecology. The consequences of life-history and vagaries of weather cause there to be great variation in intensity and outcomes of interactions among species and cause intrinsically large variation in abundances of species. As a result, ecologists working in intertidal habitats are continually grappling with the search for general understanding. Second, experimental analyses against this background of variation have shed light on many processes. Some are undoubtedly more important in intertidal habitats than elsewhere, but nevertheless provide a framework of possibilities for interpreting ecological patterns in any habitat. The third point is that, in common with other areas of ecology, ingenuity and inventiveness in the development of methods have been very successful in finding ways to extract signals from the intrinsic noise of the system. Where might the illustrated themes be leading? There are two components of the future of intertidal ecology. The first involves development and synthesis of other types of biology into ecological understanding. So, the distances dispersed by species with short- or long-term larval stages may be amenable to analysis by the methods of molecular genetics (Grosberg and Quinn, 1986; Gosling and McGrath, 1990; Gallardo and Carrasco, 1996). Certainly, the extent to which individuals in any place originate together (e.g. Hedgecock, 1979) and the distances across which there are coherent scales of connectedness among populations along coast-lines (Yamada, 1987) should be more usefully resolved by genetical analyses. Another issue requiring inputs from non-ecologists is the development of methods to understand the functional roles of diversity seen in rocky intertidal habitats. To what extent the species found are ‘redundant’ is not clear. For example, Menge et al. (1986), because of logistic constraints imposed by the diversity of species in the habitat, analysed predators in functional guilds. The extent to which the analysis was successful indicates that groups of fast-acting or slowly moving predators may have internal redundancies. Some of the species have such overlapping functional roles that their removal or disappearance would make little difference. It is also true that competition for food among micro-algal grazers (limpets and snails) is ‘diffuse’. Many species overlap in use of the same resources. So, again, it may not matter if all the species are present. In contrast, if there are widespread influences of keystone species, there already exists evidence that, at least for some components of assemblages, there are crucial functional roles. This idea has already been embraced in conservation biology as a tool to help focus on species needing priority for protection (Soule´ and Simberloff, 1986; Terborgh, 1986). Needless to say, the dangers of such an approach have also been pointed out (Mills et al., 1993). Attempts to analyse functional redundancy involve very particular types of experi- A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 67 ments (Tilman and Downing, 1994; Naeem et al., 1995; Tilman, 1997). These are fraught with difficulties. For example, some of the experiments have been criticized because results may have been be due to the specific mix of species put together as an experimental assemblage rather than because some set of species was excluded (Huston, 1997). So, for intertidal systems, much more will need to be understood about direct and indirect interactions (an ecological task) and about methods of feeding, reproduction, etc. (which are tasks for biological or zoological analysis) before potential redundancy can be well understood. The second component of the future of experimental intertidal ecology involves integration of ecological understanding into a better, more coherent whole. Peters (1991) roundly castigated ecologists for failing to develop better predictive models. A major task of intertidal ecologists has been to develop and test by experiment models that explain what processes operate locally to create patterns. Better predictive capacity will come from changes in focus to understand how often, where, when and in what combinations the various processes actually operate. This requires a considerable change of scaling of the types of study being done. In particular, it will require at least the five following developments. First, there is increasing evidence that small-scale variations in patterns of dispersion and abundance of intertidal species are important and must be understood in the context of inertia, resilience and stability of populations (reviewed by Underwood, 1989). Experiments repeatedly demonstrate the sorts of behaviours that influence local dispersion and numbers of mobile species (Chapman and Underwood, 1994; Chapman, 1999). Modelling of individual behaviours is becoming increasingly successful (Burrows and Hughes, 1989, 1991) and depends on good experimental understanding of the sorts of responses individuals make to various features of their environment. Second, there must be an increased spatial scale of investigation. This has two separate components: increasing the actual scale and increasing the range of places and the range of types of habitats over which experiments are done. The rationale for those were reviewed by Underwood and Petraitis (1993) in the context of comparisons of ecology from one coast-line to another. Until valid comparisons can be made, it will remain impossible to understand how similar or different are processes from one place to another within or between geographical regions. Without this understanding, general and generally predictive understanding are unlikely to be achieved. There have been successful examples of studies over a wide range of habitats or places on rocky shores (Dayton, 1971; Menge, 1976; Wethey, 1984b; Castilla and Paine, 1987; Fairweather and Underwood, 1991; Chapman and Underwood, 1998). There have been fewer examples of studies in which the actual spatial scale of investigation was large. One exception has been Bustamante and Branch’s (1996) study of the ecology of limpets on two coast-lines around South Africa, which analysed the inter-relationships between diversity of limpets in relation to productivity of coastal waters. More such larger-scale analyses will be necessary, provided that care is taken to integrate the comparative experimental approach at smaller scales nested in the larger scales (Menge et al., 1994; Menge, 2000). Care must also be taken to ensure that the analysis at larger scales is commensurable with similar studies in other places or habitats (Underwood and Petraitis, 1993). The remaining two of the five developments concern temporal scale. The majority of 68 A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 ecological studies are quite short, usually three to five years. This is often long enough for important features, such as indirect interactions to appear (Menge, 1997), but not nearly long enough to understand some major processes. Connell and Sousa (1983) have made a plea for longer-term studies, at least over a turn-over of individuals in a species in any habitat. Unless such information is accumulated, it will continue to be difficult to understand whether there is persistence of populations, or the extent to which there is temporal equilibrium in their abundances (Connell and Sousa, 1983). For many intertidal species in south-eastern Australia, a period of five–seven years seems about the span needed to observe a turn-over of the individuals. So, studies will need to double in length! In fact, analyses of equilibria may need much longer study to gain sufficient temporal independence of observation to allow any statistically valid analysis (Keough and Butler, 1983). Only where long-term studies have been done with consistent methods and rationales is there any hope of detecting long-term patterns. For example, surveys of populations of intertidal barnacles have been on-going around the south of Britain for many years (Southward and Crisp, 1954; Southward, 1967, 1991; Southward et al., 1995), leading now to detection of some compelling patterns of change with large-scale climatic cycles (Southward et al., 1995). Long-term studies are also needed because of temporal hysteresis in the ways assemblages respond to disturbances. As one example, Underwood (1998b, 1999a) documented some responses to storms over a very short period (weeks) in 1974. Recovery of algal cover required about six–eight years. Other components of the assemblage have not yet (to 1998) shown any recovery (summarized in Underwood, 1999a). So, short-term changes can have long-term responses. Finally, as with larger spatial scales, there needs to be an increase in experimental analyses repeated over time. Too frequently, studies in one place at one time are published. Because of constraints of time (due to programmes for grants and, in many countries, Ph.D.s being modally about three–four years), any repetition or comparison is in space. It is quite uncommon to have experiments in the same habitat done several times. Where this has been done, it provides important evidence about the general validity of results from any one experimental series (Underwood and Barrett, 1990). It also helps validate comparisons across habitats by unconfounding temporal and spatial variation (Crowe, 1996). Where experiments have been repeated, they can illustrate temporal and spatial inconsistencies (Chapman and Underwood, 1998), but can also assist with interpretation of results that are too variable at any one time to make much sense. An example was a test of hypotheses about responses of the intertidal snail Littorina unifasciata to experimental manipulations of local topography. Results at any one time (even though there was considerable spatial replication) were very unclear. After numerous (12) repeats of the experiment over several years, results were exceedingly coherent (Underwood and Chapman, 1992). The requirements impose some logistic challenges, but there is evidence that they can be overcome. For example, the Eurorock experiments on roles of intertidal grazing gastropods and recruitment of several species of barnacles have been done using comparable designs over a huge coast-line from Sweden via Ireland, U.K., northern A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76 69 Spain, Portugal and Italy (G. Chelazzi, S. Hawkins and L. Benedetti-Cecchi, pers. comm.). Results are not yet all published, but the way forward is clear given that logistic constraints have been shown to be resolvable. The greater spatial and temporal scales also require developments of sure ways to integrate results from repeated experimentation (Underwood, 1997), including development of meta-analytical techniques (Gurevitch et al., 1992). Despite the problems, experimental intertidal ecology would seem to be thriving and robust and to have a promising future. To what extent it can offer anything to other fields is up to the practitioners in those fields. Intertidal ecologists will continue to absorb and parasitize concepts and methods from elsewhere. Having a substantial experimental base on which to build, some solid and long-lasting edifices of theory and understanding should be produced over the next few years. There are grounds for optimism that these will have graceful designs and will be functionally practical, so that (to pursue the architectural theme) future ecologists will be comfortable living within them. Acknowledgements Preparation of this paper was supported by the Australian Research Council through the Centre for Research on Ecological Impacts of Coastal Cities. I thank many colleagues in the Centre for discussions of relevant issues, Tas Crowe for making me think better about repeated experiments, Lisandro Benedetti-Cecchi and Steve Hawkins for keeping me up-to-date on Eurorock and, above all, Gee Chapman for advice, discussion and criticism. I thank Nando Boero at the Universita` di Lecce, Italy, who provided accommodation during the final preparation of the paper. The paper was reviewed by two anonymous referees by being first submitted to another journal. The paper has been extensively modified using the comments of these referees. References Allen, T.F.H., Starr, T.B., 1982. Hierarchy: Perspectives for Ecological Complexities. University of Chicago Press, Chicago. Andrewartha, H.G., Birch, L.C., 1954. The Distribution and Abundance of Animals. 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