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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
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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’
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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
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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
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(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
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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
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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).
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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
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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
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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. University of Chicago
Press, Chicago.
Andrewartha, H.G., Birch, L.C., 1984. The Ecological Web. University of Chicago Press, Chicago.
Bernstein, B.B., Zalinski, J., 1983. An optimum sampling design and power tests for environmental biologists.
J. Environ. Manage. 16, 335–343.
Birch, L.C., 1957. The meanings of competition. Am. Nat. 91, 5–18.
Branch, G.M., 1981. The biology of limpets: physical factors, energy flow, and ecological interactions. Ann.
Rev. Oceanogr. Mar. Biol. 19, 235–380.
Branch, G.M., 1984. Competition between marine organisms: ecological and evolutionary implications. Ann.
Rev. Oceanogr. Mar. Biol. 22, 429–593.
Brawley, S.H., 1992. Mesoherbivores. In: John, D.M., Hawkins, S.J., Price, J.H. (Eds.), Plant Animal
Interactions in the Marine Benthos. Clarendon Press, Oxford, pp. 235–263.
Burdick, R.K., Graybill, F.A., 1992. Confidence Intervals on Variance Components. Marcel Dekker, New York.
70
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
Burrows, M.T., Hughes, R.N., 1989. Natural foraging of the dogwhelk Nucella lapillus (Linnaeus): the weather
and whether to feed. J. Moll. Stud. 55, 285–296.
Burrows, M.T., Hughes, R.N., 1991. Variation in foraging behaviour among individuals and populations of
dogwhelks, Nucella lapillus: natural constraints on energy intake. J. Anim. Ecol. 60, 497–514.
Bustamante, R.H., Branch, G.M., 1996. Large scale patterns and trophic structure of southern African rocky
shores: the role of geographic variation and wave-exposure. J. Biogeog. 23, 339–351.
Caffey, H.M., 1982. No effect of naturally-occurring rock types on settlement or survival in the intertidal
barnacle Tesseropora rosea (Krauss). J. Exp. Mar. Biol. Ecol. 63, 119–132.
Castilla, J.C., Bustamante, R.H., 1989. Human exclusion from rocky intertidal of las Cruces, central Chile:
effects on Durvillaea antarctica (Phaeophyta Durvilleales. Mar. Ecol. Prog. Ser. 50, 203–214.
Castilla, J.C., Duran, L.R., 1985. Human exclusion from the rocky intertidal zone of central Chile: the effects
of Concholepas concholepas (Gastropoda). Oikos 45, 391–399.
Castilla, J.C., Paine, R.T., 1987. Predation and community organization on Eastern Pacific, temperate zone,
rocky intertidal shores. Rev. Chil. Hist. Nat. 60, 131–151.
Chaloupka, M.Y., Hall, D.N., 1985. An examination of species dispersion patterns along the intertidal gradient
on Macquarie Island (Sub-Antarctica) using a restricted occupancy model. J. Exp. Mar. Biol. Ecol. 84,
133–144.
Chapman, M.G., 1986. Assessment of some controls in experimental transplants of intertidal gastropods. J.
Exp. Mar. Biol. Ecol. 103, 181–201.
Chapman, M.G., 1999. Assessment of variability in responses of intertidal periwinkles to experimental
transplantation. J. Exp. Mar. Biol. Ecol. 236, 171–190.
Chapman, M.G., Underwood, A.J., 1992. Experimental designs for analyses of movements by molluscs. In:
Grahame, J., Mill, P.J., Reid, D.G. (Eds.), Proceedings of the Third International Symposium on Littorinid
Biology. The Malacological Society of London, London, pp. 169–180.
Chapman, M.G., Underwood, A.J., 1994. Dispersal of the intertidal snail, Nodilittorina pyramidalis, in
response to topographic complexity of the substratum. J. Exp. Mar. Biol. Ecol. 179, 145–169.
Chapman, M.G., Underwood, A.J., 1998. Inconsistency and variation in the development of rocky intertidal
algal assemblages. J. Exp. Mar. Biol. Ecol. 224, 265–289.
Coe, W.R., 1956. Fluctuations in populations of littoral marine invertebrates. J. Mar. Res. 15, 212–232.
Colman, J.S., 1933. The nature of the intertidal zonation of plants and animals. J. Mar. Biol. Assoc. UK 18,
435–476.
Connell, J.H., 1961. The influence of interspecific competition and other factors on the distribution of the
barnacle Chthamalus stellatus. Ecology 42, 710–723.
Connell, J.H., 1972. Community interactions on marine rocky intertidal shores. Ann. Rev. Ecol. Syst. 3,
169–192.
Connell, J.H., 1975. Some mechanisms producing structure in natural communities: a model and evidence
from field experiments. In: Cody, M.S., Diamond, J.M. (Eds.), Ecology and Evolution of Communities.
Harvard University Press, Cambridge, MA, pp. 460–490.
Connell, J.H., 1978. Diversity in tropical rain forests and coral reefs. Science 199, 1302–1310.
Connell, J.H., 1983. On the prevalence and relative importance of interspecific competition: evidence from
field experiments. Am. Nat. 122, 661–696.
Connell, J.H., 1985. The consequences of variation in initial settlement versus post-settlement mortality in
rocky intertidal communities. J. Exp. Mar. Biol. Ecol. 93, 11–46.
Connell, J.H., Sousa, W.P., 1983. On the evidence needed to judge ecological stability or persistence. Am. Nat.
121, 789–824.
Creese, R.G., Underwood, A.J., 1982. Analysis of inter- and intra-specific competition amongst limpets with
different methods of feeding. Oecologia (Berl.) 53, 337–346.
Crowe, T., 1996. Different effects of microhabitat fragmentation on patterns of dispersal of an intertidal
gastropod in two habitats. J. Exp. Mar. Biol. Ecol. 206, 83–107.
Darwin, C., 1859. On the Origin of Species. Murray, London.
Dayton, P.K., 1971. Competition, disturbance and community organization: the provision and subsequent
utilization of space in a rocky intertidal community. Ecol. Monogr. 41, 351–389.
˜ and patch stability in a southern California
Dayton, P.K., Tegner, M.J., 1984. Catastrophic storms, El Nino,
kelp community. Science 224, 283–285.
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
71
Duffy, J.E., Hay, M.E., 1991. Food and shelter as determinants of food choice by an herbivorous marine
amphipod. Ecology 72, 1286–1298.
Eckman, J.E., 1996. Closing the larval loop: linking larval ecology to the population dynamics of marine
benthic invertebrates. J. Exp. Mar. Biol. Ecol. 200, 207–237.
Elner, R.W., Vadas, R.L., 1990. Inference in ecology: the sea urchin phenomenon in the Northwestern Atlantic.
Am. Nat. 136, 108–125.
Estes, J.A., Duggins, D.O., 1995. Sea otters and kelp forests in Alaska: generality and variation in a
community ecological paradigm. Ecol. Monogr. 65, 75–100.
Fairweather, P.G., 1988. Predation creates haloes of bare space among prey on rocky shores in New South
Wales. Aust. J. Ecology 13, 401–409.
Fairweather, P.G., Underwood, A.J., 1983. The apparent diet of predators and biases due to different handling
times of their prey. Oecologia (Berl.) 56, 169–179.
Fairweather, P.G., Underwood, A.J., 1991. Experimental removals of a rocky intertidal predator: variations
within two habitats in the effects on prey. J. Exp. Mar. Biol. Ecol. 154, 29–75.
Fairweather, P.G., Underwood, A.J., Moran, M.J., 1984. Preliminary investigations of predation by the whelk
Morula marginalba. Mar. Ecol. Prog. Ser. 17, 143–156.
´
ˆ
´
Fischer-Piette, E., 1936. Etudes
sur la biogeographie
intercotidale
des deux rives de la Manche. J. Linn. Soc.
Lond. 40, 181–272.
Foster, M.S., 1982. Factors controlling the intertidal zonation of Iridaea flaccida (Rhodophyta). J. Phycol. 18,
285–294.
Foster, M.S., 1990. Organization of macroalgal assemblages in the Northeast Pacific: the assumption of
homogeneity and the illusion of generality. Hydrobiologia 192, 21–34.
Foster, M.S., Schiel, D.R., 1988. Kelp communities and sea otters: keystone species or just another brick in the
wall. In: Van Blaricom, G.R., Estes, J.A. (Eds.), The Community Ecology of Sea Otters. Springer-Verlag,
Berlin, pp. 93–115.
Fretwell, S.D., 1987. Food chain dynamics: the central foraging theory of ecology? Oikos 50, 291–301.
Gallardo, M.H., Carrasco, J.I., 1996. Genetic cohesiveness among populations of Concholepas concholepas
(Gastropoda Muricidae) in Southern Chile. J. Exp. Mar. Biol. Ecol. 197, 237–249.
Glasby, T.M., 1997. Analysing data from post-impact studies using asymmetrical analyses of variance: a case
study of epibiota on marinas. Aust. J. Ecol. 22, 448–459.
Gosling, E.M., McGrath, D., 1990. Genetic variability in exposed-shore mussels, Mytilus spp., along an
environmental gradient. Mar. Biol. 104, 104–113.
Gray, D.R., Naylor, E., 1996. Foraging and homing behaviour of the limpet, Patella vulgata: a geographical
comparison. J. Moll. Stud. 62, 121–124.
Green, R.H., 1979. Sampling Design and Statistical Methods for Environmental Biologists. Wiley, Chichester.
Grosberg, R.K., Quinn, J.F., 1986. The genetic control and consequences of kin recognition by the larvae of a
colonial marine invertebrate. Nature 322, 456–459.
Gurevitch, J., Morrow, L.L., Wallace, A., Walsh, J.S., 1992. A meta-analysis of competition in field
experiments. Am. Nat. 140, 539–572.
Hall, D.J., Cooper, W.E., Werner, E.E., 1970. An experimental approach to the production dynamics and
structure of freshwater communities. Limnol. Oceanogr. 15, 839–928.
Hartnoll, R.G., Wright, J.R., 1977. Foraging movements and homing in the limpet Patella vulgata L. Anim.
Behav. 25, 806–810.
Hatton, H., 1932. Quelques observations sur le peuplement en Fucus vesiculosus des surfaces rochenses
denudees. Bull. Lab. Mar. Mus. Hist. Nat. 9, 1–6.
Hedgecock, D., 1979. Biochemical genetic variation and evidence of speciation in Chthamalus barnacles of the
tropical eastern Pacific Ocean. Mar. Biol. 54, 207–214.
Hjort, J., 1914. Fluctuations in the great fisheries of northern Europe viewed in the light of biological research.
´
Rapp. Proc.-Verb. Reun.
Cons. Int. Explor. Mer 20, 1–228.
Hughes, T.P., 1990. Recruitment limitation, mortality and population regulation in open systems: a case study.
Ecology 71, 21–26.
Hughes, R.N., Elner, R.W., 1979. Tactics of a predator, Carcinus maenas, and morphological responses of the
prey, Nucella lapillus. J. Anim. Ecol. 48, 65–78.
72
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
Hughes, R.N., Seed, R., 1995. Behavioural mechanisms of prey selection in crabs. J. Exp. Mar. Biol. Ecol.
193, 225–238.
Huston, M.A., 1997. Hidden treatments in ecological experiments: re-evaluating the ecosystem function of
biodiversity. Oecologia (Berl.) 110, 449–460.
Jackson, J.B.C., 1981. Interspecific competition and species’ distributions: the ghost of theories and data past.
Am. Zool. 21, 889–902.
Jernakoff, P., 1985. An experimental evaluation of the influence of barnacles, crevices and seasonal patterns of
grazing on the algal diversity and cover in an intertidal barnacle zone. J. Exp. Mar. Biol. Ecol. 88, 287–302.
Johnson, C.R., Mann, K.H., 1988. Diversity, patterns of adaptation, and stability of Nova Scotia kelp bed.
Ecol. Monogr. 58, 129–154.
Kastendiek, J., 1982. Competitor-mediated coexistence: interactions among three species of benthic macroalgae. J. Exp. Mar. Biol. Ecol. 62, 201–210.
Keough, M.J., Butler, A.J., 1983. Temporal changes in species number in an assemblage of sessile marine
invertebrates. J. Biogeog. 10, 317–330.
King James Bible. Mathew 13, 3.
Kuhn, T., 1970. In: The Structure of Scientific Revolutions, 2nd Edition. University of Chicago Press,
Chicago.
Lawton, J.H., Hassell, M.P., 1981. Asymmetrical competition in insects. Nature 289, 793–795.
Levin, S.A., Paine, R.T., 1974. Disturbance, patch formation, and community structure. Proc. Natl. Acad. Sci.
USA 71, 2744–2747.
Lewin, R., 1986. Supply-side ecology. Science 234, 25–27.
Lewis, J.R., 1964. The Ecology of Rocky Shores. The English Universities Press, London.
Little, C., Morritt, D., Paterson, D.M., Stirling, P., Williams, G.A., 1990. Preliminary observations on factors
affecting foraging activity in the limpet Patella vulgata. J. Mar. Biol. Assoc. UK 70, 181–195.
Lohse, D.P., 1993. The importance of secondary substratum in a rocky intertidal community. J. Exp. Mar. Biol.
Ecol. 166, 1–17.
Loosanoff, V.L., 1964. Variations in time and intensity of setting of the starfish Asterias forbesi in Long Island
Sound during a twenty-five year period. Biol. Bull. 126, 423–439.
Loosanoff, V.L., 1966. Time and intensity of setting of the oyster, Crassostrea virginica, in Long Island Sound.
Biol. Bull. 130, 211–227.
Lubchenco, J., 1978. Plant species diversity in a marine intertidal community: importance of herbivore food
preference and algal competitive abilities. Am. Nat. 112, 23–39.
Mann, K.H., 1982. Kelp, sea urchins and predators: a review of strong interactions in rocky subtidal systems of
eastern Canada, 1970–1980. Neth. J. Sea Res. 16, 414–423.
Mann, K.H., Breen, P.A., 1972. The relationship between lobster abundance, sea urchins and kelp beds. J. Fish.
Res. Bd Canada 29, 603–605.
McDonald, L.L., Erickson, W.P., Strickland, M.D., 1993. Survey design, statistical analysis, and basis for
statistical inference in coastal habitat injury assessment: Exxon Valdez oil spill. In: Wells, P.G., Butler, J.N.,
Hughes, J.S. (Eds.), Exxon Valdez Oil Spill: Fate and Effects in Alaskan Waters. American Society for
Testing and Materials, Philadelphia, PA, pp. 296–311.
McGuinness, K.A., 1987a. Disturbance and organisms on boulders. I. Patterns in the environment and the
community. Oecologia (Berl.) 71, 409–419.
McGuinness, K.A., 1987b. Disturbance and organisms on boulders. II. Causes of patterns in diversity and
abundance. Oecologia (Berl.) 71, 420–430.
Menge, B.A., 1976. Organization of the New England rocky intertidal community: role of predation,
competition and environmental heterogeneity. Ecol. Monogr. 46, 335–393.
Menge, B.A., 1992. Community regulation: under what conditions are bottom-up factors important on rocky
shores? Ecology 73, 755–765.
Menge, B.A., 1997. Detection of direct versus indirect effects: were experiments long enough? Am. Nat. 149,
801–823.
Menge, B.A., 2000. Top-down and bottom-up community regulation in marine habitats. J. Exp. Mar. Biol.
Ecol. 250, 257–289.
Menge, B.A., Berlow, E.L., Blanchette, C.A., Navarrete, S.A., Yamada, S.B., 1994. The keystone species
concept: variation in interaction strength in a rocky intertidal habitat. Ecol. Monogr. 64, 249–286.
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
73
Menge, B.A., Lubchenco, J., Ashkenas, L.R., Ramsey, F., 1986. Experimental separation of effects of
consumers on sessile prey in the low zone of a rocky shore in the Bay of Panama: direct and indirect
consequences of food web complexity. J. Exp. Mar. Biol. Ecol. 100, 225–270.
´ M.E., Doak, D.F., 1993. The keystone species concept in ecology and conservation.
Mills, L.S., Soule,
Bioscience 43, 219–224.
Moran, M.J., 1985. Distribution and dispersion of the predatory intertidal gastropod, Morula marginalba. Mar.
Ecol. Prog. Ser. 22, 41–52.
Moran, M.J., Fairweather, P.G., Underwood, A.J., 1984. Growth and mortality of the predatory intertidal whelk
Morula marginalba Blainville (Muricidae): the effects of different species of prey. J. Exp. Mar. Biol. Ecol.
75, 1–17.
Moreno, C.A., Sutherland, J.P., Jara, F.H., 1984. Man as a predator in the intertidal zone of central Chile.
Oikos 42, 155–160.
Morton, J.E., Miller, M., 1968. The New Zealand Seashore. Collins, London.
Naeem, S., Thompson, L.J., Lawler, S.P., Lawton, J.H., Woodfin, R.M., 1995. Empirical evidence that
declining species diversity may alter the performance of terrestrial ecosystems. Phil. Trans. R. Soc., Ser. B
347, 249–262.
Newell, R., 1976. Adaptations to intertidal life. In: Newell, R.C. (Ed.), Adaptations to Environment: Essays on
the Physiology of Marine Animals. Butterworth’s, London, pp. 1–82.
Orton, J.H., 1937. Some interrelations between bivalve spatfalls, hydrography and fisheries. Nature 140, 505.
Paine, R.T., 1966. Food web complexity and species diversity. Am. Nat. 100, 65–75.
Paine, R.T., 1969. The Pisaster-Tegula interaction: prey patches, predator food preference, and intertidal
community structure. Ecology 50, 950–962.
Paine, R.T., 1974. Intertidal community structure: experimental studies on the relationship between a dominant
competitor and its principal predator. Oecologia (Berl.) 15, 93–120.
Paine, R.T., 1977. Controlled manipulations in the marine intertidal zone, and their contributions to ecological
theory. Acad. Sci. Spec. Pub. 12, 245–270.
Paine, R.T., 1991. Between Scylla and Charybdis: do some kinds of criticism merit a response? Oikos 62,
90–92.
Paine, R.T., Castilla, J.C., Cancino, J., 1985. Perturbation and recovery patterns of starfish-dominated intertidal
assemblages in Chile, New Zealand, and Washington State. Am. Nat. 125, 679–691.
Paine, R.T., Levin, S.A., 1981. Intertidal landscapes: disturbance and the dynamics of pattern. Ecol. Monogr.
51, 145–178.
Paine, R.T., Vadas, R.L., 1969. The effects of grazing by sea urchins, Strongylocentrotus spp. on benthic algal
populations. Limnol. Oceanogr. 14, 710–719.
Peters, R.H., 1991. A Critique for Ecology. Cambridge University Press, Cambridge.
Petraitis, P.S., Dunham, A.E., Niewiarowski, P.H., 1996. Inferring multiple causality: the limitations of path
analysis. Funct. Ecol. 10, 421–431.
Pielou, E.C., 1974. Population and Community Ecology: Principles and Methods. Gordon and Breach, New
York.
Popper, K.R., 1968. The Logic of Scientific Discovery. Hutchinson, London.
Power, M.E., 1990. Top-down and bottom-up forces in food-webs: do plants have primacy? Ecology 73,
733–746.
Root, R.B., 1967. The niche exploitation pattern of the blue-gray gnatcatcher. Ecol. Monogr. 37, 317–350.
Roughgarden, J., Iwasa, Y., Baxter, C., 1985. Demographic theory for an open marine population with
space-limited recruitment. Ecology 66, 54–67.
Scheltema, R.S., 1971. Larval dispersal as a means of genetic exchange between geographically separated
populations of shallow-water benthic marine gastropods. Biol. Bull. 140, 284–322.
Schoener, T.W., 1983. Field experiments on intraspecific competition. Am. Nat. 122, 240–285.
Searle, S.R., Casella, G., McCulloch, C.E., 1992. Variance Components. John Wiley and Sons, New York.
Shanks, A.L., 1995. Orientated swimming by megalopae of several eastern North Pacific crab species and its
potential role in their onshore migration. J. Exp. Mar. Biol. Ecol. 186, 1–16.
Shanks, A.L., Wright, W.G., 1987. Internal wave-mediated shoreward transport of cyprids, megalopae, and
gammarids and correlated longshore differences in the settling rate of intertidal barnacles. J. Exp. Mar. Biol.
Ecol. 114, 1–14.
74
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
Shkedy, Y., Roughgarden, J., 1997. Barnacle recruitment and population dynamics predicted from coastal
upwelling. Oikos 80, 487–498.
Simberloff, D., 1983. Competition theory, hypothesis testing, and other community ecological buzzwords. Am.
Nat. 122, 626–635.
Slocum, C.J., 1980. Differential susceptibility to grazers in two phases of an intertidal alga: advantages of
heteromorphic generations. J. Exp. Mar. Biol. Ecol. 46, 99–110.
´ M.E., Simberloff, D., 1986. What do genetics and ecology tell us about the design of nature reserves?
Soule,
Biol. Cons. 35, 19–40.
Sousa, W., 1979a. Disturbance in marine intertidal boulder fields: the nonequilibrium maintenance of species
diversity. Ecology 60, 1225–1239.
Sousa, W.P., 1979b. Experimental investigations of disturbance and ecological succession in a rocky intertidal
algal community. Ecol. Monogr. 49, 227–254.
Sousa, W.P., 1980. The responses of a community to disturbance: the importance of successional age and
species life histories. Oecologia (Berl.) 45, 72–81.
Southward, A.J., 1967. Recent changes in abundance of intertidal barnacles in south-west England: a possible
effect of climatic deterioration. J. Mar. Biol. Assoc. UK 47, 81–95.
Southward, A.J., 1991. 40 years of changes in species composition and population density of barnacles on a
rocky shore near Plymouth. J. Mar. Biol. Assoc. UK 71, 495–514.
Southward, A.J., Crisp, D.J., 1954. Recent changes in the distribution of the intertidal barnacles Chthamalus
stellatus Poli and Balanus balanoides L. in the British Isles. J. Anim. Ecol. 23, 163–177.
Southward, A.J., Hawkins, S.J., Burrows, M.T., 1995. Seventy years observations of changes in distribution
and abundance of zooplankton and intertidal organisms in the Western English Channel in relation to rising
sea temperature. Therm. Biol. 20, 127–155.
Spight, T.M., 1974. Sizes of populations of a marine snail. Ecology 55, 712–729.
Stephenson, T.A., Stephenson, A., 1972. Life Between Tidemarks on Rocky Shores. W.H. Freeman and Co,
San Francisco.
Stewart-Oaten, A., Murdoch, W.M., Parker, K.R., 1986. Environmental impact assessment: ‘pseudoreplication’
in time? Ecology 67, 929–940.
Stimson, J., 1970. Territorial behavior of the owl limpet Lottia gigantea. Ecology 51, 113–118.
Stimson, J., 1973. The role of the territory in the ecology of the intertidal limpet Lottia gigantea (Gray).
Ecology 54, 1020–1030.
Sutherland, J.P., 1990. Perturbations, resistance, and alternative views of the existence of multiple stable points
in nature. Am. Nat. 136, 270–275.
Tansley, A.G., Adamson, R.S., 1925. Studies on the vegetation of the English chalk. III. The chalk grasslands
of the Hampshire-Sussex border. J. Ecol. 13, 177–223.
´ M.E. (Ed.), Conservation
Terborgh, J., 1986. Keystone plant resources in the tropical forest. In: Soule,
Biology: the Science of Scarcity and Diversity. Sinauer Press, Sunderland, MA, pp. 330–344.
Thorson, G., 1946. Reproduction and larval development of Danish marine bottom invertebrates. Medd.
Komm. Danm. Fisk. Havunders., Serie: Plank 4, No. 1, 523 pp.
Thorson, G., 1950. Reproductive and larval ecology of marine bottom invertebrates. Biol. Rev. 25, 1–45.
Tilman, D., 1997. Distinguishing between the effects of species diversity and species composition. Oikos 80,
185.
Tilman, D., Downing, J.A., 1994. Biodiversity and stability in grasslands. Nature 367, 363–365.
Underwood, A.J., 1978a. An experimental evaluation of competition between three species of intertidal
prosobranch gastropods. Oecologia (Berl.) 33, 185–208.
Underwood, A.J., 1978b. A refutation of critical tidal levels as determinants of the structure of intertidal
communities on British shores. J. Exp. Mar. Biol. Ecol. 33, 261–276.
Underwood, A.J., 1979. The ecology of intertidal gastropods. Adv. Mar. Biol. 16, 111–210.
Underwood, A.J., 1984. Vertical and seasonal patterns in competition for microalgae between intertidal
gastropods. Oecologia (Berl.) 64, 211–222.
Underwood, A.J., 1986a. The analysis of competition by field experiments. In: Kikkawa, J., Anderson, D.J.
(Eds.), Community Ecology: Pattern and Process. Blackwells, Melbourne, pp. 240–268.
Underwood, A.J., 1986b. What is a community. In: Raup, D.M., Jablonski, D. (Eds.), Patterns and Processes in
the History of Life. Dahlem Konferenzen, Springer-Verlag, Berlin, pp. 351–367.
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
75
Underwood, A.J., 1988. Design and analysis of field experiments on competitive interactions affecting
behaviour of intertidal animals. In: Chelazzi, G., Vannini, M. (Eds.), Behavioural Adaptation to Intertidal
Life. Plenum Press, New York, pp. 333–358.
Underwood, A.J., 1989. The analysis of stress in natural populations. Biol. J. Linn. Soc. 37, 51–78.
Underwood, A.J., 1990. Experiments in ecology and management: their logics, functions and interpretations.
Aust. J. Ecol. 15, 365–389.
Underwood, A.J., 1992a. Competition in marine plant–animal interactions. In: John, D.M., Hawkins, S.J.,
Price, J.H. (Eds.), Plant Animal Interactions in the Marine Benthos. Clarendon Press, Oxford, pp. 443–475.
Underwood, A.J., 1992b. Beyond BACI: the detection of environmental impact on populations in the real, but
variable, world. J. Exp. Mar. Biol. Ecol. 161, 145–178.
Underwood, A.J., 1993. The mechanics of spatially replicated sampling programmes to detect environmental
impacts in a variable world. Aust. J. Ecol. 18, 99–116.
Underwood, A.J., 1994. On beyond BACI: sampling designs that might reliably detect environmental
disturbances. Ecol. Appl. 4, 3–15.
Underwood, A.J., 1996a. Detection, interpretation, prediction and management of environmental disturbances:
some roles for experimental marine ecology. J. Exp. Mar. Biol. Ecol. 200, 1–27.
Underwood, A.J., 1996b. Spatial patterns of variance in densities of intertidal populations. In: Floyd, R.B.,
Sheppard, A.W., De Barro, P.J. (Eds.), Frontiers of Population Ecology. CSIRO Publishing, Melbourne, pp.
369–389.
Underwood, A.J., 1997. Experiments in Ecology: Their Logical Design and Interpretation Using Analysis of
Variance. Cambridge University Press, Cambridge.
Underwood, A.J., 1998a. Design, implementation and analysis of ecological experiments: pitfalls in the
maintenance of logic structures. In: Resetarits, W.J., Bernardo, J. (Eds.), Issues in Experimental Ecology.
Oxford University Press, Oxford, pp. 325–349.
Underwood, A.J., 1998b. Grazing and disturbance: an experimental analysis of patchiness in recovery from a
severe storm by the intertidal alga Hormosira banksii on rocky shores in New South Wales. J. Exp. Mar.
Biol. Ecol. 231, 291–306.
Underwood, A.J., 1999a. Physical disturbances and their direct effects on an indirect effect: responses of an
intertidal assemblage to a severe storm. J. Exp. Mar. Biol. Ecol. 232, 125–140.
Underwood, A.J., 1999b. History and recruitment in the structure of intertidal assemblages on rocky shores: an
introduction to problems for interpretation of natural change. In: Whitfield, M. (Ed.), Aquatic Life Cycle
Strategies. Institute of Biology, London, pp. 79–96.
Underwood, A.J., 1999c. Trying to detect impacts in marine habitats: comparisons with suitable reference
areas. In: Sparks, T. (Ed.), Statistics in Ecotoxicology. Wiley and Sons, Chichester, pp. 279–308.
Underwood, A.J., Barrett, G., 1990. Experiments on the influence of oysters on the distribution, abundance and
sizes of the gastropod Bembicium auratum in a mangrove swamp in New South Wales. J. Exp. Mar. Biol.
Ecol. 137, 25–45.
Underwood, A.J., Chapman, M.G., 1992. Experiments on topographic influences on density and dispersion of
Littorina unifasciata in New South Wales. In: Grahame, J., Mill, P.J., Reid, D.G. (Eds.), Proceedings of the
Third International Symposium on Littorinid Biology. The Malacological Society of London, London, pp.
181–195.
Underwood, A.J., Chapman, M.G., 1996. Scales of spatial patterns of distribution of intertidal invertebrates.
Oecologia (Berl.) 107, 212–224.
Underwood, A.J., Chapman, M.G., 1998. A method for analysing spatial scales of variation in composition of
assemblages. Oecologia (Berl.) 117, 570–578.
Underwood, A.J., Denley, E.J., 1984. Paradigms, explanations and generalizations in models for the structure
of intertidal communities on rocky shores. In: Strong, D.R., Simberloff, D., Abele, L.G., Thistle, A. (Eds.),
Ecological Communities: Conceptual Issues and the Evidence. Princeton University Press, New Jersey, pp.
151–180.
Underwood, A.J., Denley, E.J., Moran, M.J., 1983. Experimental analyses of the structure and dynamics of
mid-shore rocky intertidal communities in New South Wales. Oecologia (Berl.) 56, 202–219.
Underwood, A.J., Fairweather, P.G., 1989. Supply-side ecology and benthic marine assemblages. TREE 4,
16–20.
76
A. J. Underwood / J. Exp. Mar. Biol. Ecol. 250 (2000) 51 – 76
Underwood, A.J., Petraitis, P.S., 1993. Structure of intertidal assemblages in different locations: how can local
processes be compared. In: Ricklefs, R.E., Schluter, D. (Eds.), Species Diversity in Ecological Communities: Historical and Geographical Perspectives. University of Chicago Press, Chicago, pp. 38–51.
Wethey, D.S., 1984a. Effects of crowding on fecundity in barnacles Semibalanus ( Balanus) balanoides,
Balanus glandula and Chthamalus dalli. Can. J. Zool. 62, 1788–1795.
Wethey, D.S., 1984b. Spatial pattern in barnacle settlement: day to day changes during the settlement season. J.
Mar. Biol. Assoc. UK 64, 687–698.
Whittaker, R.H., 1956. Vegetation of the Great Smoky mountains. Ecol. Monogr. 26, 1–80.
Winer, B.J., Brown, D.R., Michels, K.M., 1991. Statistical Principles in Experimental Design, Third Edition.
McGraw-Hill, New York.
Wolcott, T.G., 1973. Physiological ecology and intertidal zonation in limpets (Acmaea): a critical look at
‘limiting factors’. Biol. Bull. 145, 389–422.
Wootton, J.T., 1993. Indirect effects and habitat use in an intertidal community: interaction chains and
interaction modifications. Am. Nat. 141, 71–89.
Wootton, J.T., 1994a. The nature and consequences of indirect effects in ecological communities. Ann. Rev.
Ecol. Syst. 25, 443–466.
Wootton, J.T., 1994b. Predicting direct and indirect effects: an integrated approach using experiments and path
analysis. Ecology 75, 151–165.
Yamada, S.B., 1987. Geographic variation in the growth rates of Littorina littorea and L. saxatilis. Mar. Biol.
96, 529–534.
Young, C.M., 1987. Novelty of ‘supply-side ecology’. Science 235, 415–416.