experimental ecology of rocky intertidal habitats: what are we learning?

Journal of Experimental Marine Biology and Ecology250 (2000) 51–76

www.elsevier.nl / locate / jembe

Experimental ecology of rocky intertidal habitats: what arewe learning?

*A.J. UnderwoodCentre 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 animalsand plants on rocky shores have been a major activity for many years. In this review, some of thethemes and topics that have emerged from such analyses are briefly discussed to provide anup-date for practitioners and ecologists working in other habitats. Conceptual issues include thewidespread occurrence of transphyletic use of the same resources (space and food), theories andexperimental analyses of intermediate disturbance in relation to numbers of species, the complexbut 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 ex-perimental designs for complex and patchy, interacting sets of species, the importance of controlsin experimental manipulations and methods for analyses of hierarchical scales of patterns andprocesses. Finally, some contributions to social issues (pollution, biodiversity) and some scenariosfor 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 ofinterest for decades in many parts of the world. There have long been descriptions of

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52 A.J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 –76

fauna and flora (Colman, 1933; Fischer-Piette, 1936) and some manipulative ex-perimentation goes back at least to the 1930s (Hatton, 1932).

The ecological study of the organisms has been, in many ways, a source of, or, atleast, 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 cataloguingof 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 onexperimental analyses of patterns of distribution (Connell, 1972; Paine, 1977).

It is often timely to revisit some of the themes and constructs of a particular disciplineor 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 anidiosyncratic assessment of where we, the practitioners in or aficionados of a field mightconsider ourselves to be. It may, however, also serve to provide an up-date for those inother areas of ecology, to explain what and where current research in intertidal ecologyis at present.

This special edition of Journal of Experimental Marine Biology and Ecology providesan opportunity for such a brief overview. The topic is covered as a voyage of discoveryabout the themes intertidal ecologists are learning (hence its title) and the essay willhave been worthwhile if that is all it achieves. On the other hand, if it also helps explaincurrent 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 ofexperimental manipulations as a crucial investigative tool. This has been quitephenomenal and represents a major shift of emphasis that has occurred during the sameperiod as the massive increase in publications of science following the expansion ofuniversities and grant-funding in the 1960s (Underwood, 1996a).

It is worth a small amount of space on a reprise of why rocky intertidal habitats havebeen at the forefront of development of ecological experimentation. Most notably, arocky intertidal shore encompasses a gradient of environmental conditions from fullymarine below low tidal levels to fully terrestrial where splash and spray reach to thehighest levels above high tide. Mostly, although not always, the gradient occurs over asmall (metres to tens of metres) distance making some patterns of response toenvironmental variables relatively easy to see (Newell, 1976). Short distances acrossgradients also allow relatively easy observation and manipulation of environmentalvariables (e.g. remote sensors in different parts of the range are still quite close to eachother).

On such gradients, there is usually a great variety of animals and plants, often havingseveral representatives in functional groups or guilds (Menge et al., 1986), e.g. grazingsnails, predatory whelks, crabs, large fleshy, brown seaweeds, etc. Such diversityfacilitates comparative studies, often allowing the generality of processes to be assessedby 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 appropriatescales, and the mechanics and logistics are usually manageable. Animals and plants maylive for extraordinarily long periods, but typically turn over in relatively few years. Theytherefore complete their life-cycles in the sort of time-scale that is matched by cycles ofgrant-funding and scholarships for Ph.D.s. This has the serious down-side that studieshave focussed on the conspicuous, abundant, slow-moving and shorter-lived componentsof assemblages. This has ignored many of the long-lived plants (Slocum, 1980) and themore 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 vastmajority of studies have been focussed on species characterized by relative immobility.The immobility has been emphasized by ignoring the potentially great distancestravelled (Scheltema, 1971) by the dispersive larval phases of life-history, although therehave been exceptions such as Shanks and Wright (1987), Shanks (1995), Eckman(1996) and Shkedy and Roughgarden (1997). Despite such biases, animals and plants onrocky shores are quite suited by visibility, size, diversity, longevity, abundance and lackof emotional appeal to be subjects of experimental tests of hypotheses about ecologicalpatterns and processes. So, where are we now and what are we discovering from suchexperimentation?

2. Some conceptual issues

2.1. Trans-phyletic analyses of use of resources

One area that has been extensively exploited in experimental analyses of processescausing and maintaining ecological patterns in intertidal habitats has been the role ofcompetition for resources of food and space (reviews by Branch, 1984; Underwood,1986a, 1992a). Two-dimensional space is a resource required by almost all intertidalspecies, either directly as an absolute need (Andrewartha and Birch, 1954) for space onwhich 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 totalamount of two-dimensional space available at any location and the varying and oftenunpredictable nature of abundances and mixtures of species occupying the space,competition is often intense and sustained. So, as examples, competition among andwithin species of limpets for food is a widespread and normal aspect of the ecology oflimpets on South African shores (Branch, 1981, 1984), competition for barnacles is awell-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 andintriguing complexity of interactions among similar sorts of organisms. For example,Kastendiek (1982) described an interesting situation where the turfing red alga, Halidrysdioica, outcompetes the alga, Pterocladia capillacea, by growing over and denying itaccess 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’


competition from H. dioica. Thus, competition between E. arborea and H. dioicaprevents overgrowth of P. capillacea by H. dioica.

Equally interesting and well-documented from many experimental studies is thewidespread occurrence of competitive interactions between very different sorts oforganisms. Without entering the labyrinthine maze of numerous types and complextaxonomy of competitive interactions (Schoener, 1983), competition on rocky shores isof three general types. Pre-emptive competition occurs wherever occupation of space byone species prevents another species from settling from the plankton (which may allowthe later arrival to find space elsewhere; Underwood and Denley, 1984). Interferencecompetition is that involving a ‘contest’ (e.g. Pielou, 1974), so that one user of spacedirectly harms or kills another. Examples are a faster-growing barnacle that undercuts orsmothers a slower-growing species (Connell, 1961) or mussels smothering other speciesby growing over them (Paine, 1974; Menge, 1976). Finally, there is exploitative or‘scramble’ competition (Pielou, 1974) where several species need the same space tofeed, 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 betweenanimals and plants).

Pre-emption prevents barnacles from settling where algal fronds sweep the surfaces ofthe rock (Dayton, 1971) and where the cover of plants prevents settlement or attachmentof larvae. Direct interference occurs where, for example, mussels encroach on thefeeding 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 thatthe outcome may not actually be any increased damage or increased risk of mortality tothe ‘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 oraccidental calamity. The hypothesis that larvae prevented from settling in one spot aremore 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 forcareful 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’ ofstudy in assemblages (see discussion in Underwood, 1986b), but are not a relevantgrouping where resources are used by many organisms that are not taxonomicallyrelated.

2.2. Intermediate disturbance and competitive interactions

One consequence of widespread, complex competitive interactions is that any otherprocess (disturbance, disease, predation) leading to reductions in densities or cover ofcompetitors can have indirect effects on species not directly involved. For example, themodel 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) andConnell (1978) to explain the often observed downwards concave curve of number ofspecies across a gradient of disturbance. The model states that where disturbance is large


(or frequent or recent), only those species capable of withstanding it (or colonizing andgrowing since the last disturbance) can survive. This is only a subset of the species thatwould 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 thismodel must be tested by manipulations of the regime of disturbance. No amount ofdescribing patterns of species richness across gradients of disturbance will help.

Two very convincing series of experiments have been done in rocky intertidal habitatsto test predictions about changes and the processes causing changes in numbers ofspecies when disturbances are manipulated. Sousa (1979a,b, 1980) increased disturbancein a boulder-field by turning boulders over more frequently than occurred naturally. Thisdid lead to reductions in abundance of some species. Reducing disturbance (byexperimentally preventing boulders from being turned over) did increase elimination ofgreen species by red algal species. The major conclusions were, however, that featuresof life-history (rate of colonization, nature of cycle of breeding) and responses todisturbance were important reasons why the model of intermediate disturbance did notapply very well to explain the observed patterns of numbers of species. Responses oflong-lived red algae to being disturbed included ‘grab-and-hold’ strategies wherebyvegetative growth from surviving remnants retained occupation of space by thosespecies. In the end, in areas that were not much disturbed or not disturbed often, theperennial red species became dominant by occupying space vacated by other specieswhen it became available, rather than by overgrowing and eliminating other speciescompetitively.

McGuinness (1987a,b) examined the consequences of experimentally increasing ordecreasing 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 somecombinations of components of the fauna, under only some conditions. Intermediatedisturbance did not seem a widespread explanation for patterns of difference in numbersof 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 relativeabundances of species because it consumes superior competitors in an assemblage. Thebest-known example is the starfish, Pisaster ochraceus, which eats mussels as a majorcomponent of its food. The mussels are capable of smothering many, if not all, of theother users of primary space on the shore. So, the predators, by removing mussels, makespace continuously available for other species and thereby increase diversity. There havebeen critical evaluations of this example in terms of the evidence (Fairweather andUnderwood, 1983; Underwood and Denley, 1984) and the extent to which the process iswidespread (e.g. Foster, 1990). It is also clear that areas with mussels generally support


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 andBreen, 1972; Mann, 1982; Estes and Duggins, 1995), with continuing criticism of theirvalidity and the lack of attention to other explanatory models for observed patterns ofnumbers of species (Foster and Schiel, 1988; Elner and Vadas, 1990). There has alsobeen some uncritical adoption of the concept in areas far removed from its empirical

´origins, such as in discussions of issues for biological conservation (Soule andSimberloff, 1986; Terborgh, 1986). Here, again, the validity of the concept, or theuntested 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 theirimportance in the ecology of complex assemblages (reviewed by Menge, 1997). Thiswas, of course, an older tradition dating back to Darwin’s ‘‘web of complex ‘interac-tions’ ’’ (Darwin, 1859). It generated some attempt at a novel theoretical synthesis ofcomponents of an organism’s environment (Andrewartha and Birch, 1984). Certainelements (mates, food, predators) were defined by Andrewartha and Birch (1984) to bein 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 abundanceof an animal.

The most recent synthesis has been by Wootton (1993, 1994a) who has demonstratedthat 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 secondspecies (B), which, in turn, has direct influences on a third species (C). For example, apredatory whelk (A) consumes barnacles (B) that occupy space, making it unsuitable forgrazing limpets (C) (see Dayton, 1971; Underwood et al., 1983, for examples). As aresult, consumption of barnacles by the whelks can lead to local increases in numbers oflimpets because of the reduction in competition for space. Predation directly negativelyinfluences numbers of barnacles, but indirectly positively affects numbers of limpets.

In the second case (an interaction modification), an indirectly acting species influencesthe direct interaction between two species. So, a predatory whelk (A) consumesbarnacles (C), but is itself eaten by predatory crabs (B) (see examples in Hughes andElner, 1979; Hughes and Seed, 1995). Thus, the direct reduction of numbers of whelksdue to the activities of crabs may indirectly cause an increase in the number ofbarnacles.

Analyses of these interactions can require very carefully formulated hypotheses andthe appropriate manipulative experiments to test them. Methods used include pathanalysis (Wootton, 1994b), but such approaches are saddled with all the problems of anyderivative of techniques of multiple regression (Petraitis et al., 1996).

Also, there can be problems with attempting to fit particular cases into a theoreticaldichotomous framework. For example, Underwood (1999a) demonstrated experimentallythat whelks (Morula marginalba) shelter under the canopy of an alga (Hormosirabanksii), creating reductions in densities of a prey species, the barnacle Chamaesiphotasmanica (see also related studies by Fairweather et al., 1984; Moran, 1985; Fair-weather, 1988).


So, the alga directly increases the abundance of the whelk; the whelk directlydecreases the numbers of the barnacle. Removal of the algal canopy causes localincreases in abundances of the barnacle, an indirect effect of the first type identified asan interaction chain by Wootton (1993). At the same time, however, the rate andintensity of the direct predatory actions of the whelks (i.e. reducing the numbers ofbarnacles) are influenced by the size of the canopy, so the canopy has an indirect effectof the second type identified by Wootton (1993). Despite potential problems withanalyses and interpretations, complex chains of indirect interactions are an importantarea for study, particularly because they will prove crucial for effective understanding offunctional 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 theidea that structure of assemblages can be regulated by ‘bottom-up’ processes. In suchprocesses, there may be quantitative or qualitative differences in the structure ofassemblages with different levels of nutrients in the system. This has been a feature ofsome interpretations of factors controlling structure and composition of assemblages infreshwater habitats (e.g. Hall et al., 1970; Power, 1990). The argument is that, whereprimary production is greater, there can be greater abundances and/or greater diversityof grazers exploiting the large primary production.

This idea contrasts with well-established notions in intertidal ecology that areinterpreted as ‘top-down’ control. So, for example, predatory animals may consumesufficient grazers in any area, thereby preventing excessive numbers in or, sometimes,eliminating species from patches of habitat. Alternatively, competitive interactions maydirectly 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 isemerging as an important issue in the analysis of complex ecological systems (Fretwell,1987; Menge, 1997; reviewed by Menge, 2000). Determining how important either typeof process may be under different circumstances can be advanced by the experimentalopportunities offered by experiments on rocky shores.

There are potential managerial or conservatory issues associated with this concept. Forexample, where top-down processes matter, reductions in density or removal ofpredators or large, competitively dominant grazers or users of space can have profoundimpacts on other components of the assemblage (Paine and Vadas, 1969; Moreno et al.,1984; Castilla and Duran, 1985; Castilla and Bustamante, 1989). Species that haveabundances regulated by predators can explode in numbers with concomitant alterationsin density-dependent processes influencing other species. So, management for conserva-tion needs to be concerned with harvesting, fishing and any disruptive processes thatmay alter relative abundances of the top-down, regulating species.

Where bottom-up processes are more important, the managerial issues must revolvearound preventing alterations to productivity, in particular, to aspects of eutrophicationand other manifestations of overabundant nutrients.

Menge (1992, 2000) perceptively pointed out that no ecological system is likely to becontrolled solely in one or other direction. There is also the well-known problem thatprimary production may, itself, regulate the number of trophic levels in an assemblage


(Hall et al., 1970; Fretwell, 1987). In such a case, the increased trophic complexity mayallow for increased types of predators and increased variety, interactions and strengths ofpredatory or other top-down controlling processes. Were this to be the case, it becomesphilosophically unclear how top-down processes could control an assemblage, given thatthe diversity of higher trophic levels, where many of the top-down controlling speciesare found, is itself controlled by bottom-up processes.

Nevertheless, more study is needed of the relationships between and synthesis ofregulatory processes operating in opposite directions in assemblages.

2.5. Supply-side ecology

Another area of intertidal investigation that has been of some influence in thedevelopment of conceptual methods is the notion that supply of recruits into any patchof habitat is an important influence. This is, in no way, a new idea, but it wasunfashionable 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 byLewin (1986) to summarize, in a pun, the fact that various important processes can onlyoccur or can only take place at relevant magnitudes and rates if the species involved inthem are present in sufficient numbers.

As an example, predatory starfish or whelks cannot be involved in top-down controlof 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/orplace to place, there will be great variation in the duration, timing or frequency of anyparticular process.

Sometimes, as explained for a variety of examples by Underwood (1979) andUnderwood and Denley (1984), prevailing explanations of observed patterns in intertidalassemblages fail to include the possibility that larval supply or recruitment of juvenilescould be important. So, for one example, Connell (1975) described a conceptual modelfor recolonization of a disturbed patch of habitat. In relatively benign parts of theenvironment, predators were presumed to be able to eliminate most, if not all, of preyspecies under most prevailing patterns of weather. Occasionally, however, the predatorsare absent (e.g. whelks are missing because they were killed by an unusual period ofharsh weather; Dayton, 1971). Consequently, in those periods, prey arrive as larvae,settle and survive. If they then survive long enough to become sufficiently large, theywill escape being consumed by predators when the predators finally re-appear. Thismodel can explain why there are intermittent ‘pulses’ of appearance of species of prey insome intertidal habitats.

The alternative, supply-side explanation (Underwood and Denley, 1984) is thatpredators fluctuate in abundance because of variations in their own recruitment and notbecause of occasional periods of harsh weather. Regardless of the causes or amounts ofsuch variation, numbers of prey fluctuate a lot because of the vagaries of larvalproduction, dispersal and survival. Occasionally, they will arrive in large numbers whennumbers 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


intense predation, sufficient prey survive long enough to reach sizes at which predatorscan no longer harm them. Under this explanatory model, it is not unusual weather eventsthat reduce numbers of predators. Instead, larger-than-normal recruitment of preydictates 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 hasbeen widely understood in fisheries science. In fact (apart from being understood withrespect to agriculture in biblical times; King James Bible), the earliest reference on thetopic known to me was Hjort (1914) discussing fisheries. There have been manyobservations 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 ofthe consequences (e.g. Orton, 1937).

Thorson (1946, 1950) was the first marine ecologist to try to use variation inrecruitment as a mechanism in models explaining temporal and spatial variations inabundances of animals. He noted that three species of bivalves with long larval periodsof planktonic development had abundances of adults that fluctuated much more than wasthe case for three species living in the same habitats, but which had a short period or nopelagic development. He also commented on the consequences of the timing ofrecruitment. When the larvae of a bivalve arrived before the larvae of one of theirpredators (a starfish), they were able to survive for long enough to be too large to beconsumed by the predators when these eventually arrived. In contrast, if the bivalvesrecruited after the starfish, many more were consumed. So, timing of recruitment couldalso influence the sizes of populations of adults.

The notion of supply-side ecology has become more important in recent developmentsof meta-models of patchy populations of marine invertebrates (Underwood andFairweather, 1989). The best developments have been the models developed byRoughgarden (e.g. Roughgarden et al., 1985). She developed a model for space-limitedhabitats that linked the numbers of barnacles settling on the shore to the amount of freespace available for settlement on that shore and to the probability of recruits survivingfrom one time to another. The model was highly successful in some situations. Forexample, 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 beregulated by densities and fluctuations of recruits. These results conform well to someempirical observations (see particularly Connell, 1985; Hughes, 1990; Sutherland, 1990).The models are less successful at providing useful insights or predictive capacity forspecies that have variable rates of recruitment from time to time (e.g. barnacles in NewSouth Wales discussed in Underwood and Denley, 1984; Underwood, 1999b).

Nevertheless, more recent advances have begun to demystify the vagaries of whatSpight (1974) called the planktonic mystery stage (see review by Eckman, 1996). As aparticular example, the consequences of variations in Roughgarden et al.’s (1985)‘recruitment parameter’ (the number of cyprids settling within a period of time per unitarea of free space that survive to the end of that period) can be predicted for somespecies of barnacles on the coast of California. The numbers likely to be available forsettlement can be very satisfactorily predicted from a polynomial regression of numbers


thof recruits on sea-surface temperature (to the 11 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 ofmagnitude of recruitment, supply-side models and coupling with upwelling and othercoastal oceanographic processes can provide considerable understanding of and capacityto 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 withimprovements to its methods, particularly those concerned with the design, analysis andinterpretation of quantitative and experimental data. Some of these contributions arebriefly described here.

3.1. Experimental designs: competition

The formal analysis of competitive interactions has been fraught with difficulties ofunderstanding the relevant processes and scales (Connell, 1983; Schoener, 1983) and oflogical 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 worthiterating once more because they seem to have eluded some authors. If it is proposedthat two species (A and B) have negative effects on each other because of their jointneeds for some resources (Birch, 1957), there are two major procedures. First, theamount of resource may be manipulated. For example, the amount of food in areas canbe experimentally increased (or decreased) to test the hypothesis that outcomes ofcompetition will be less (or more) stark. Second, the numbers of consumers of resourcescan be manipulated, again to test hypotheses about directions and magnitudes ofinteractions among the consumers.

In the simplest experiments, to examine the influence of species B on species A, somerelevant density of A must be created in the absence of B. As the second experimentaltreatment, the same density of A must be established with a relevant density of B. Thus,the minimal experiment has independently replicated arenas with density N of A andA

independently replicated areas with N plus density M of B. To determine theA B

reciprocal influence of species A on species B, there must also be independentlyreplicated arenas with M of species B alone and M of species B with N of species A.B B 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 ofnon-independence of data, it will be more appropriate to establish two sets of replicatearenas for the combined treatment and use one set to provide data for species A and theother set to provide data for species B. Attempts to do such experiments by holding thetotal density of organisms constant are confounded. So, some authors investigate theeffect of B on A with N individuals of species A in one treatment and a total of N


summed from species A and B in the other treatment (see Underwood, 1986a). The latteralters the mix of species (as required by the hypothesis) and simultaneously alters thedensity of species A. Any comparison between the treatments cannot test the hypothesisas stated (see detailed discussion in Underwood, 1986a, 1992a).

It is often the case by hypothesis and wherever per capita influences of com-petitive interactions must be determined, that the influences of each species mustbe investigated at several densities. Thus, there must now be different densities ofspecies A (N , N , etc.), as required. Each of these must also be established atA A1 2

the appropriate densities of species B. This creates a two-factorial matrix with(N , N 1 M , N 1 M , . . . . . . . . . , N , N 1 M , N 1 M , . . . . . . . . . ,A A B A B A A B A B1 1 1 1 2 2 2 1 2 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 interac-tions. The experiment becomes a little more complex where possible asymmetries(Lawton and Hassell, 1981; Connell, 1983; Schoener, 1983) in the intensity ofcompetition between two or more species must also be investigated. For this to bepossible, an experiment must simultaneously include treatments with the same additionsas before, but this time of the same species (A added to A; B added to B). For thesimplest case, of one density of species A (N ) and one density of species B (M ), thereA B

must be treatments N , N 1 M ) as before, plus (N 1 M ) to determine the magnitudeA A B A A

of intraspecific competition, i.e., species A on species A, relative to that of interspecificcompetition, i.e., species B on species A. Then, there must also be (M , M 1 N ,B B A

M 1 N ) to determine the influences on species B. If the asymmetry is to be comparedB B

between the two species, in addition to measuring it for each species separately, thereshould really also be treatments (M ; M 1 N , M 1 N ; N ; N 1 M , N 1 M ) toA A A A B B B A B B

ensure that a comparison of per capita influences does not confound inter- and intra-specific differences with differences in the density of the two species.

These designs and their interpretation have become quite standard in studies ofcompetition on rocky shores. They are well-suited to the manipulation of densities andcomposition of grazing species, particularly where the species are abundant, so thatexperimental plots are small and there are many individuals to make up the experimentaldensities.

3.2. Experimental designs: transplantation

Another quite common requirement of studies of distributions of organisms acrossgradients is to be able to transplant individuals from one part of the gradient to anotherto test specific types of hypotheses (reviewed by Chapman, 1986, 1999; Underwood,1988; Chapman and Underwood, 1992). For example, suppose there is a gradient in sizeof mobile animals with smaller ones at higher levels on the shore. In some areas, anappropriate 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 individualwandering at random downshore will tend to stay there, whereas a small animal willmove upwards. Reasons for such behaviour include perception of an increased risk of


predation, even though more food is available at lower levels (e.g. Paine, 1969). Largeranimals can move to and survive at lower levels because they are less likely to be eatenby a predator than are the small individuals. Alternatively, in some cases, more foodmay 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 higherlevels. 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 ifsome large and some small individuals are transplanted downshore, small individualswill change their behaviour and will show greater tendencies to move upshore than dolarge animals, and shown by small ones in the original habitat.

To test the above hypothesis requires appropriate controls for disturbing the animalswhilst moving them and for moving them to a new habitat, regardless of it being at alower 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 inthose surroundings. As a result, they become disorientated and tend to move more oftenand to move greater distances upshore, as a response to such disorientation.

Under these circumstances, discovering a greater tendency to move upshore does notunambiguously support the hypothesis. Such a conclusion is potentially confounded withany effects of disorientation. Appropriate controls must include translocations, i.e., thedisturbance of animals that are moved to a new location in the same (upper) part of thegradient 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 suchecological experiments (wherever they are to be done) have been greatly elucidated byexperimentation on rocky shores.

3.3. Variation in processes and hierarchies in patterns

Ecologists studying rocky intertidal habitats have been very concerned with spatialand temporal variability in the patterns and processes that influence distributions andabundances 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 indiscrete and abrupt boundaries between low and high tide, persist in the literature. Theyhave, however, never been supported by quantitative data and have been refuted byquantitative tests of their predictions (Underwood, 1978b; Chaloupka and Hall, 1985).They are as inaccurate a description of the distributions of species across intertidalgradients as were descriptions of series of communities of plants at different heights onthe Smoky Mountains. The latter were demolished by quantitative sampling byWhittaker (1956).

Instead, there has more recently been a focus on patch dynamics and attempts tounderstand and model responses to mixtures of disturbances, physical factors andvariable rates and intensities of competition and predation (see particularly the mix oftheory and experimentation by Levin and Paine (1974) and Paine and Levin (1981)). Allof these interacting ecological processes are affected by issues of recruitment (see


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 inprocesses operating at different spatial and temporal scales as hierarchies, rather than asinteractions at one place or time of investigation. Some of the background to consideringecological scales in hierarchies was summarized by Allen and Starr (1982). Examples ofthe sort of ecological models that involve hierarchies of processes operating at differentscales were provided for subtidal kelp-beds in studies by Dayton and Tegner (1984).

In analyses of intertidal habitats, there have been several approaches to consideringhierarchies of spatial scales. One was the survey done by Foster (1990). Along the coastof California, Foster examined a series of typical rocky headlands chosen because theyhad similar physical characteristics. On each shore, he examined a series of patterns todetermine how widespread or how frequent they were. The object of the exercise was totest predictions derived from models about processes influencing the local structure ofassemblages. So, predation on superior competitive mussels had been proposed as awidespread and important influence on distributions (Paine, 1974; see earlier discussionof keystone predation). Similarly, Foster’s (1982) own work on competition for spacebetween algae was thought to be an important process.

If any of the processes considered was, in fact, important, Foster (1990) hypothesizedthat the patterns resulting from the processes should be found frequently over a set ofshores for which the processes were claimed to be operating. This hypothesis was notgenerally supported by the data. Despite objections to this approach (Paine, 1991), it hasgreat merit. If the patterns that are supposed to be the result of some process are notwidespread, it is difficult to maintain an argument that the process occurs widely. It isnot ‘nihilist’ (Paine, 1991) to question dogma by testing hypotheses about outcomes ofsupposedly general processes. When predictions fail, new models and understanding areneeded (Popper, 1968; Simberloff, 1983; Underwood, 1990). There must usually be adelay between discovering that some previous paradigm must be overthrown because ithas failed and proposing new models that incorporate the older ideas and the newobservations that failed to confirm them (Kuhn, 1970).

So, this approach examines the frequency of patterns that should result from variousprocesses. An alternative, sometimes called a comparative experimental approach(Menge et al., 1994), uses experimental procedures, done at small scales, but arrangedacross larger spatial scales. This was used successfully to identify the variable responsesto removals of predatory starfish along a coast-line (discussed earlier) and to comparethis interaction across coastlines (Paine et al., 1985). It has also served well todemonstrate the inconsistencies in colonization and development of algal assemblageson low-shore rocky habitats on the exposed coast-line of New South Wales (Chapmanand Underwood, 1998). It was essential for Wethey’s (1984b) analysis of short-termvariation in settlement of barnacles on British shores and Caffey’s (1982) experimentaltests of hypotheses about the influences of different types of rock on the settlement ofbarnacles. This approach has, however, not yet managed to synthesize results of somecomplex intertidal issues, for example, the timing, frequency and duration of foraging byintertidal homing limpets on British shores (Hartnoll and Wright, 1977; Little et al.,1990; Gray and Naylor, 1996).


One of the problems with this comparative approach is that it is better to plan theexperiments to make specific contrasts interpretable, i.e., to propose and test specifichypotheses about different types of habitats (or weathers, or seasons, etc.). This hasnumerous advantages over the attempt to gather together the outcomes of variousexperiments on similar topics, but done for different purposes without directlycomparable designs and without the appropriate spatial and temporal replication (see thediscussion in Underwood and Petraitis, 1993).

So, planned comparative experimentation can reveal not only the relevant variation inprocesses at large spatial scales (for example, variations in diets and growth of predatorson shores dominated by different types of prey; Moran et al., 1984), but also thesimilarities in shores of similar type compared to the differences from one habitat toanother (for example, the influences of experimental removals of predatory whelks fromtwo different intertidal habitats; Fairweather and Underwood, 1991). Doing experimentsover short periods, but repeating them in a planned manner over several time-periodshad similar advantages for understanding temporal variation (Underwood and Chapman,1992).

The third and final approach to investigating hierarchies in ecological processes andtheir resulting patterns has been the analysis of the spatial (or temporal) hierarchy itself.Some examples showing methods of analysis and results for spatial variation inabundances 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, blockmean 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 rockyshores, investigations to test hypotheses at scales from tens of kilometres down tocentimetres can be done by sets of experimental or sampling units at small spatialintervals repeated at sites, locations, etc., that are different distances apart. The otherprocedures investigated all required much more effort, were very time-consuming toreplicate and impossible to do over very large spatial scales.

The second important result from these types of analyses is that specific hypothesesabout a hierarchical series of processes operating simultaneously can be tested incomparison with each other. So, very small-scale (centimetres to metres) variation indensities of an intertidal snail can be shown to be the result of small-scale behaviouralresponses by individuals to local topography, food, micro-climate. Larger-scale (metresto tens of metres) variability can be attributed to variation in biological processes ofpre-emption, interference, predation. At yet larger scales (tens to hundreds of metres),variation may largely be affected by variations in recruitment or physical disturbancesdue to weather. At even larger scales of hundreds of metres to kilometres, there may beconsistent variation due to wave-action and storms. Finally, at very large spatial scales(tens to hundreds of kilometres), there may be biogeographic variation caused byconsistent latitudinal differences in climate.

Hierarchical analysis of variance is a robust tool for extracting estimates of variancefrom data collected at these scales, so that comparisons can be made about the relativemagnitudes of such variance (e.g. Burdick and Graybill, 1992; Searle et al., 1992;


Underwood, 1997). Where this has been used for intertidal species (Underwood, 1996b;Underwood and Chapman, 1996), the outcome has almost universally been that variationis very great at the smallest spatial scales. This not only implies a very great importancefor small-scale interactions of behaviour and ecology in response to food, micro-climateand topography. It also, fortunately, justifies assumptions of independence, thusvalidating experimental manipulations where replicate experimental units have beenseparated 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 helpsolve various problems of a practical, environmental nature. Three examples willillustrate the point. First, there have been developments of the asymmetrical samplingdesigns needed for detecting and estimating the sizes of environmental impacts(reviewed by Underwood, 1994). It has long been realized that impacts can only bedefined and detected as a statistical interaction in time and space (Green, 1979). Theremust be a different pattern of change in some relevant variable(s) from before to after ahuman disturbance in the disturbed site compared to undisturbed, reference areas.Routinely, such interactions have been detected using a comparison of the disturbed to asingle undisturbed site (BACI procedures; Bernstein and Zalinski, 1983; Stewart-Oatenet al., 1986). These procedures are unreplicated, so the comparison is always potentiallyconfounded. ‘Beyond BACI’ procedures compare the site disturbed with a sample ofundisturbed sites (Underwood, 1992b, 1993, 1994) and arose from the asymmetricalanalytical procedures used to analyse competitive interactions (Underwood, 1978a,1984). These procedures have been extended to situations where there are no data beforethe disturbance (Glasby, 1997).

A more recent approach to this latter problem involves a meta-analysis of a series ofpaired comparisons, each of one disturbed and a paired undisturbed site (McDonald etal., 1993). Any interpretation of the result of each comparison would be confounded (asabove), but the whole set of comparisons provides independent replication of the tests,providing an unconfounded interpretation. In the case of oil-spills, for which thesemethods were originally used, there may be problems with finding independent sets ofoiled 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) ofunivariate analyses of measures of composition and relative abundance of species inintertidal assemblages to be able to use the previously mentioned hierarchical analysesof spatial variation. The methods generate independent measures of multivariatedifferences among replicates in samples at different scales. These can then be analysedby the versatile procedures available for univariate measures (Winer et al., 1991;Underwood, 1997). Such techniques may be helpful in assessments of scales of variationof 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


with the very variable abundances of populations, patchiness of occupation of habitat,rates and intensities of processes influencing densities and sizes and the composition ofassemblages in intertidal habitats is an on-going research programme in intertidalecology. It is not surprising that some of the outcomes spill into areas of appliedproblem-solving.

4. Conclusions

The examples discussed briefly above serve to illustrate three themes in experimentalintertidal ecology. The consequences of life-history and vagaries of weather cause thereto be great variation in intensity and outcomes of interactions among species and causeintrinsically large variation in abundances of species. As a result, ecologists working inintertidal habitats are continually grappling with the search for general understanding.Second, experimental analyses against this background of variation have shed light onmany processes. Some are undoubtedly more important in intertidal habitats thanelsewhere, but nevertheless provide a framework of possibilities for interpretingecological patterns in any habitat. The third point is that, in common with other areas ofecology, ingenuity and inventiveness in the development of methods have been verysuccessful 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 thefuture of intertidal ecology. The first involves development and synthesis of other typesof biology into ecological understanding. So, the distances dispersed by species withshort- or long-term larval stages may be amenable to analysis by the methods ofmolecular genetics (Grosberg and Quinn, 1986; Gosling and McGrath, 1990; Gallardoand Carrasco, 1996). Certainly, the extent to which individuals in any place originatetogether (e.g. Hedgecock, 1979) and the distances across which there are coherent scalesof connectedness among populations along coast-lines (Yamada, 1987) should be moreusefully resolved by genetical analyses.

Another issue requiring inputs from non-ecologists is the development of methods tounderstand the functional roles of diversity seen in rocky intertidal habitats. To whatextent 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 successfulindicates that groups of fast-acting or slowly moving predators may have internalredundancies. Some of the species have such overlapping functional roles that theirremoval or disappearance would make little difference. It is also true that competitionfor food among micro-algal grazers (limpets and snails) is ‘diffuse’. Many speciesoverlap in use of the same resources. So, again, it may not matter if all the species arepresent. In contrast, if there are widespread influences of keystone species, there alreadyexists evidence that, at least for some components of assemblages, there are crucialfunctional 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 beenpointed out (Mills et al., 1993).

Attempts to analyse functional redundancy involve very particular types of experi-


ments (Tilman and Downing, 1994; Naeem et al., 1995; Tilman, 1997). These arefraught with difficulties. For example, some of the experiments have been criticizedbecause results may have been be due to the specific mix of species put together as anexperimental 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 andindirect interactions (an ecological task) and about methods of feeding, reproduction,etc. (which are tasks for biological or zoological analysis) before potential redundancycan be well understood.

The second component of the future of experimental intertidal ecology involvesintegration of ecological understanding into a better, more coherent whole. Peters (1991)roundly castigated ecologists for failing to develop better predictive models. A majortask of intertidal ecologists has been to develop and test by experiment models thatexplain what processes operate locally to create patterns. Better predictive capacity willcome from changes in focus to understand how often, where, when and in whatcombinations the various processes actually operate. This requires a considerable changeof scaling of the types of study being done. In particular, it will require at least the fivefollowing developments.

First, there is increasing evidence that small-scale variations in patterns of dispersionand abundance of intertidal species are important and must be understood in the contextof inertia, resilience and stability of populations (reviewed by Underwood, 1989).Experiments repeatedly demonstrate the sorts of behaviours that influence localdispersion and numbers of mobile species (Chapman and Underwood, 1994; Chapman,1999). Modelling of individual behaviours is becoming increasingly successful (Burrowsand Hughes, 1989, 1991) and depends on good experimental understanding of the sortsof responses individuals make to various features of their environment.

Second, there must be an increased spatial scale of investigation. This has twoseparate components: increasing the actual scale and increasing the range of places andthe range of types of habitats over which experiments are done. The rationale for thosewere reviewed by Underwood and Petraitis (1993) in the context of comparisons ofecology from one coast-line to another. Until valid comparisons can be made, it willremain impossible to understand how similar or different are processes from one place toanother within or between geographical regions. Without this understanding, general andgenerally predictive understanding are unlikely to be achieved.

There have been successful examples of studies over a wide range of habitats orplaces on rocky shores (Dayton, 1971; Menge, 1976; Wethey, 1984b; Castilla and Paine,1987; Fairweather and Underwood, 1991; Chapman and Underwood, 1998). There havebeen fewer examples of studies in which the actual spatial scale of investigation waslarge. One exception has been Bustamante and Branch’s (1996) study of the ecology oflimpets on two coast-lines around South Africa, which analysed the inter-relationshipsbetween diversity of limpets in relation to productivity of coastal waters.

More such larger-scale analyses will be necessary, provided that care is taken tointegrate the comparative experimental approach at smaller scales nested in the largerscales (Menge et al., 1994; Menge, 2000). Care must also be taken to ensure that theanalysis 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


ecological studies are quite short, usually three to five years. This is often long enoughfor important features, such as indirect interactions to appear (Menge, 1997), but notnearly long enough to understand some major processes. Connell and Sousa (1983) havemade a plea for longer-term studies, at least over a turn-over of individuals in a speciesin any habitat. Unless such information is accumulated, it will continue to be difficult tounderstand whether there is persistence of populations, or the extent to which there istemporal equilibrium in their abundances (Connell and Sousa, 1983). For many intertidalspecies in south-eastern Australia, a period of five–seven years seems about the spanneeded to observe a turn-over of the individuals. So, studies will need to double inlength! In fact, analyses of equilibria may need much longer study to gain sufficienttemporal independence of observation to allow any statistically valid analysis (Keoughand Butler, 1983).

Only where long-term studies have been done with consistent methods and rationalesis there any hope of detecting long-term patterns. For example, surveys of populations ofintertidal barnacles have been on-going around the south of Britain for many years(Southward and Crisp, 1954; Southward, 1967, 1991; Southward et al., 1995), leadingnow 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 waysassemblages 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 theassemblage 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 experimentalanalyses repeated over time. Too frequently, studies in one place at one time arepublished. Because of constraints of time (due to programmes for grants and, in manycountries, Ph.D.s being modally about three–four years), any repetition or comparison isin space. It is quite uncommon to have experiments in the same habitat done severaltimes. Where this has been done, it provides important evidence about the generalvalidity of results from any one experimental series (Underwood and Barrett, 1990). Italso helps validate comparisons across habitats by unconfounding temporal and spatialvariation (Crowe, 1996).

Where experiments have been repeated, they can illustrate temporal and spatialinconsistencies (Chapman and Underwood, 1998), but can also assist with interpretationof results that are too variable at any one time to make much sense. An example was atest of hypotheses about responses of the intertidal snail Littorina unifasciata toexperimental manipulations of local topography. Results at any one time (even thoughthere 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 canbe overcome. For example, the Eurorock experiments on roles of intertidal grazinggastropods and recruitment of several species of barnacles have been done usingcomparable designs over a huge coast-line from Sweden via Ireland, U.K., northern


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 logisticconstraints have been shown to be resolvable.

The greater spatial and temporal scales also require developments of sure ways tointegrate results from repeated experimentation (Underwood, 1997), including develop-ment of meta-analytical techniques (Gurevitch et al., 1992).

Despite the problems, experimental intertidal ecology would seem to be thriving androbust and to have a promising future. To what extent it can offer anything to otherfields is up to the practitioners in those fields. Intertidal ecologists will continue toabsorb and parasitize concepts and methods from elsewhere. Having a substantialexperimental base on which to build, some solid and long-lasting edifices of theory andunderstanding should be produced over the next few years. There are grounds foroptimism that these will have graceful designs and will be functionally practical, so that(to pursue the architectural theme) future ecologists will be comfortable living withinthem.

Acknowledgements

Preparation of this paper was supported by the Australian Research Council throughthe Centre for Research on Ecological Impacts of Coastal Cities. I thank manycolleagues in the Centre for discussions of relevant issues, Tas Crowe for making methink better about repeated experiments, Lisandro Benedetti-Cecchi and Steve Hawkinsfor 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, whoprovided accommodation during the final preparation of the paper. The paper wasreviewed by two anonymous referees by being first submitted to another journal. Thepaper has been extensively modified using the comments of these referees.

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