Forest Ecology and Management, 63 ( 1 9 9 4 ) 2 4 7 - 3 0 0
247
Elsevier Science B.V.
The disturbance of forest ecosystems: the
ecological basis for conservative management
Peter M. Attiwill
School of Botany, The University of Melbourne, Parkville, Vic. 3052, Australia
(Accepted 30 August 1993 )
Abstract
The extensive literature on natural disturbance in forests is reviewed in terms of the hypotheses:
( 1 ) that disturbance is a major force moulding the development, structure and function of forests;
and (b) that management of forests for all their benefits can be controlled so that the effects can be
contained within those which result from natural disturbance. The causal factors of natural disturbance are both endogenous and exogenous; there are major difficulties in the formal characterization
of disturbance and of recovery after disturbance. As to the latter, the acceptance of classical generalizations of the nature of succession has led to particular difficulties in the assessment and interpretation of recovery.
Tree fall, which creates gaps, is fundamental to the development of many forests, and has been most
intensively studied in tropical forests of Central America and the Amazon and in temperate forests of
North America. Tree fall is part of autogenic change; mechanisms of gap-filling and subsequent growth
and species composition vary widely with forest type and geography.
Disturbance by wind is particularly difficult to characterize. Wind varies along a continuum; the
blow-down of an individual tree may be mostly due to autogenic processes of ageing and decay, whereas
catastrophic hurricanes and cyclones may be defined as wholly exogenous. Nevertheless, the resilience
in terms of species diversity of tropical forests following catastrophic disturbance by hurricane is
remarkable. A number of studies support the view that the tropical forest in hurricane-prone areas is
not a stable steady-state ecosystem but rather that heterogeneity is maintained by catastrophe. The
ability to regenerate by suckers and the coincidence of regenerative space and gregarious flowering
are important components of the response of rainforest following disturbance.
For much of the world, 'fire is the dominant fact of forest history'. As examples, fire and its effects
are reviewed for the northern boreal forests, oak-pine forests and north-western sub-alpine forests of
North America. The effect of fire on species composition varies with intensity and frequency. That,
together with the popular view of fire as unnatural and therefore unacceptable, places great demands
on management of forests for all of their benefits, including national parks and reserves. These difficulties also affect management of other ecosystems, such as Mediterranean-type shrublands and
heathlands where species diversity, productivity and cycles of regeneration and degradation are governed by fire as a natural disturbance.
Shifting agriculture is a traditional form of agriculture used by at least 240 million people in the
humid tropics. Shifting agriculture, together with wind, lightning and fire, is an exogenous disturbance which has little effect on soil fertility and on structure and composition of the rainforest which
re-establishes after abandonment. As the intensity of disturbance of rainforest increases, resilience of
the forest decreases and the current problems of extensive clearing for improved pasture and of uncontrolled logging are resulting in degraded ecosystems.
Regeneration follows the often extensive death of trees caused by outbreaks of insects in many
coniferous forests of northern America. This disturbance by herbivory halts increasing stagnation (as
© 1994 Elsevier Science B.V. All rights reserved 0 3 7 8 - 1 1 2 7 / 9 4 / $ 0 7 . 0 0
SSDI 0 3 7 8 - 1 1 27 ( 9 3 ) 0 3 2 8 9 - U
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P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
measured by decreasing rates of ecosystem production and nutrient cycling) and reinitiates succession. Other disturbances to forests occur through damage from ice-storms, snow avalanches, erosional
and earthquake landslides, and volcanic activity; the development of N o t h o f a g u s forests in Chile and
New Zealand is determined by such catastrophic mass movements.
An extensive literature supports the hypothesis that natural disturbance is fundamental to the development of structure and function of forest ecosystems. It follows that our management of natural
forest should be based on an ecological understanding of the processes of natural disturbance. Whether
or not we want to do this, and the extent to which we want to derive all of the benefits from the forest,
including timber, depends on social attitudes. Whereas humanism may treat conservation as the wise
husbanding of forests in the interests of social traditions and harmony, animism may give nature
unalienable rights. The conclusion from this review is that the ecological framework of natural disturbance and the knowledge of its component processes and effects provides the basis on which we can
manage our forests as a renewable resource which can be utilized so that the forests 'retain their diversity and richness for mankind's continuing benefit'. Nowhere is this management more desperately
needed than for the protection of the world's tropical forests, its peoples and their cultures.
Introduction
The literature on disturbance and on the associated concepts of stability,
susceptibility and the steady state of ecological systems is very large. Much of
this literature was synthesized in the multi-authored book The Ecology of NaturalDisturbance andPatch Dynamics (Pickett and White, 1985a). Since 1985,
there has been an increase in the number of studies of disturbance and there
has been a particular concentration of work on forested ecosystems, much of
which has been summarized in detail by Oliver and Larson ( 1990 ).
Pickett and White (1985b) state in their conclusion that: 'There can be no
doubt that disturbance is an important and widespread p h e n o m e n o n in nature... Disturbance is c o m m o n to many different systems. It functions or has
functioned at all temporal and spatial scales and levels of organization of ecological and evolutionary interest'.
The hypothesis proposed that harvesting a forest for all of its benefits, including its timber products, can be controlled so that it creates a disturbance,
the effects of which do not differ from those of natural disturbance. In this
first paper, disturbance of forests is reviewed with an emphasis on the recent
(post- 1985 ) literature. The second paper (Attiwill, 1994) reviews natural
disturbances in eucalypt forests in Australia with the aim of determining
whether or not the effects of forest harvesting can be contained within the
framework provided by the effects of natural disturbance.
Our judgement about the effects of human disturbance (or management)
will always have a level of uncertainty and the degree of uncertainty which
can be tolerated increases with the social acceptability of management and its
objectives (or politics; Walters and Holling, 1990). Thus, there is room for
great disparity between the degree of ecological uncertainty and social objec-
P.M. Attiwill /Forest Ecology and Management 63 (1994) 24 7-300
249
tives so that decreasing the degree of scientific uncertainty may not of itself
influence social acceptability. The rapidly changing social acceptability of
forest harvesting in Australia is based on a number of considerations. Most
of them are political in nature (e.g. what is the correct balance between conservation values and timber values?; do we want to export wood-chips?; what
are the employment opportunities in forestry?; is there inherent value in oldgrowth forest?; are industries running on subsidies from the state forest owners?; should we aim to get all of our timber only from plantations?) and science to a large degree is secondary. Nevertheless, science is brought to the
forefront where the various proponents see that they might use it to validate
their primary social objectives.
This paper does not embrace the proposition that all human disturbance
produces effects which differ from those of natural disturbance. Rather, I argue that there is a scale of human disturbance as there is a scale of natural
disturbance. At points on these scales, we could judge that the effects of human disturbance on ecosystem structure and function are not significantly
different from those of natural disturbance. The hypothesis is illustrated by
the effects of human disturbance on rainforest in eastern Amazonia (Fig. 1,
from Buschbacher et al., 1988; and see also Jordan, 1985; Uhl et al., 1988a).
Traditional slash-and-burn agriculture leads to secondary succession and the
recovery of diversity and productivity of the mature forest. The greater the
Mature rain
forest " ~ , , ~
., I
\
Pasture
I{.eco~er~ :
uloclerate'use
/
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Mechanized
fertilizer
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stein
Fig. 1. Degradation and recovery following various intensities of deforestation and modification of forest ecosystems in eastern Amazonia (from Buschbacher et al., 1988, with the permission of Blackwell's, Oxford).
2 50
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
disturbance (bulldozing, ploughing, fertilizers, herbicides) the less is the
probability that succession after abandonment will effect the recovery of diversity and productivity.
Disturbance, recovery and stability of forested ecosystems
The nature of disturbance
Terrestrial plant communities throughout the world are affected by a variety of major, natural disturbances. Major, natural disturbances listed by White
(1979 ) and White and Pickett (1985 ) include fire; hurricanes, windstorms
and gap dynamics; ice storms, ice push, cryogenesis and freeze damage; landslides, avalanches and other earth movements, including coastal erosion and
dune movement; coastal flooding; lava flows; karst processes; droughts, flash
floods, rare rainstorms, fluctuating water levels, alluvial processes and salinity changes; biotic disturbances including insect attack, fungal disease, browsing and burrowing animals, invasion by plants (weeds); and disturbance
caused by man.
Natural disturbance is difficult to characterize because the causal factors of
disturbance are both endogenous and exogenous and because they operate
over wide ranges of size, frequency, predictability, timing (season of the year)
and magnitude (or intensity) of impact (Webb et al., 1972; Grime, 1979;
White, 1979; Bormann and Likens, 1979a; Pickett and White, 1985b; Waring
and Schlesinger, 1985; Hopkins, 1990 ). Rykiel ( 1985 ) proposed a set of formal definitions of the various terms disturbance, stress and perturbation in
an ecological context; most authors, however, have used the terms in a general
rather than formally defined sense, an approach which I adopt in this paper.
A few examples show the difficulties of formal characterization.
Wind
Bormann and Likens (1979a) point out that 'most agents of disturbance
occur as a continuum'; wind ranges from zero to hurricane force. Where wind
blows over or breaks trees that have been weakened by growth stresses and
decay (autogenic development ), the disturbance is undoubtedly endogenous,
'an integral part of the developmental process of an ecosystem' (Bormann
and Likens, 1979a). At some intensity, where wind destroys healthy as well
as weakened trees, the disturbance might be considered to be exogenous and
catastrophic.
Fire
The literature on the ecological effects of fire on plant communities
throughout the world is enormous and it has been regularly reviewed (e.g.
Ahlgren and Ahlgren, 1960; Kozlowski and Ahlgren, 1974; Spurr and Barnes,
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
2 51
1980; Rundel, 1981; Gill et al., 1981; Mooney and Godron, 1983; Attiwill,
1985; Oliver and Larson, 1990). Fires range widely both in intensity and frequency. They are mostly caused by lightning or by humans, and occasionally
by spontaneous combustion and volcanic activity. There is a well-established
history of fires in pre-European settlement times in both North America and
Australia. In Australia, for example, there is a marked and significant increase in charcoal in lake deposits dating from 120 000 years ago (Singh et
al., 1981 ). The spread and intensity of fire depends not only on prevailing
climate (moisture, temperature and wind) but on endogenous factors, such
as the quantity of fuel and its combustibility. 'Fire is not an entirely exogenous factor and may be as much a result of community structure and composition as of environment' (White, 1979).
Insect attack
Plants produce a wide variety of compounds, some of which are more or
less nutritious, and some more or less repellant, to insects (Harborne, 1988;
and for general discussion see Waring and Schlesinger, 1985; Aber and Melillo, 1991 ). Recent studies indicate that the intensity of insect attack is related both to the availability of nitrogen in the foliage (Ohmart et al., 1985 )
and to the concentration of free amino acids (White, 1984; Cockfield, 1988 ).
A number of changes in biochemistry, which affect both nutritional quality
and palatability, can result from changes in environment. There is evidence
that nitrogenous solutes which act as osmotica accumulate during periods of
water stress, and that this accumulation is associated with intensive insect
attack (Poljakoff-Mayber et al., 1987 ). Dr. M.A. Adams and I observed total
defoliation of Eucalyptus sideroxylon forest in Victoria, Australia, in the middle of the worst drought on record (1982-1983); we observed the trees recovering by epicormic shoots in the crown later in the drought, in midsummer. We measured dawn water potentials of - 6 MPa at that time.
Insect attack varies from normal and generally low levels of herbivory (e.g.
in most eucalypt forests) (Attiwill, 1979; Ohmart et al., 1983; Lamb, 1985 );
to moderate to severe herbivory, particularly where trees are placed under
stress (e.g. as remnant patches or individuals after forest clearing) (Dreistadt
et al., 1990; Landsberg et al., 1990; Adams and Atkinson, 1991 ); to severe,
episodic outbreaks epitomized by the spruce budworm Choristoneura fumiferana in North America (e.g. MacLean, 1980; Ostaff and MacLean, 1989).
To what extent is insect attack a part of the ecosystem (part of autogenesis)
and to what extent is insect attack an exogenous catastrophe?
These simple examples illustrate the difficulties of rigorous classification of
natural disturbance. White (1979) summarized the position: 'There would
seem to be a continuum from factors that are relatively endogenous to those
that are relatively exogenous in vegetation composition, and, in addition, a
continuum from chronic, normal environmental factors to acute, cata-
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P.M. A ttiwill/Forest Ecology and Management 63 (1994) 247-300
strophic ones. Classifications posed for phenomena along these continua...
lend an artificiality to vegetation concepts that depend on them'.
Timber harvesting and acid rain are examples of disturbances which are
clearly exogenous. Fire started by lightning is presumably endogenous and
fire started with malicious intent or which has unintentionally escaped is presumably exogenous. It is less easy to assign a place on the endogenous-exogenous continuum of fire, often of catastrophic scale and intensity, as an integral part of h u m a n development over 10 000 years in the Americas (Schule,
1990), over tens of thousands of years in Australia (Singh et al., 1981; Singh
and Geissler, 1985 ) and over 1.5 million years in Africa (Schule, 1990).
Recovery after disturbance
The recovery of an ecosystem after disturbance can be described in terms
of resilience 'the persistence of relationships within a system' and 'a measure
of the ability (of the system) to absorb changes of state variables, driving
variables, and parameters, and still persist' and stability 'the ability of a system to return to an equilibrium state after a temporary disturbance' (Holling,
1973; Denslow, 1985 ). Webster et al. ( 1975 ) defined resistance as the extent
to which the ecosystem is displaced (equivalent to resilience sensu Holling,
1973 ) and resilience as the rate of recovery of the ecosystem (equivalent to
stability sensu Holling, 1973 ). These terms and their definitions (Fig. 2 and
Table 1, from Swank and Waide, 1980) are essentially conceptual and their
quantification in forest ecosystems is difficult for at least two reasons. First,
ighResilience
_.E
-q
~
IDislurbance
a
~ / -
n
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e
.
LowResilience
Time
Fig. 2. Concepts of resistance and resilience following a disturbance (from Swank and Waide,
1980, with the permission of Oregon State University Press). Several responses are shown, each
differing in magnitude of deviation from the pre-disturbance level of function (resistance) and
in the rate of recovery, or return to pre-disturbance condition (resilience).
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
2 53
Table 1
Definitions of resistance and resilience, two components of ecosystem stability (from Swank and
Waide, 1980, with the permission of Oregon State University Press)
Component
Analogous terms
Resistance
Inertia
Rigidity
Resilience
Meaning
1 Inversely proportional to the amount by which some functional
property or measure of the state of the ecosystem deviates from
a nominal level in response to a specific disturbance
2 Directly proportional to the extent of disturbance required to
produce a certain magnitude of deviation in some ecosystem
functional property
1 Directly proportional to the rate at which some functional
Elasticity
measure of the state of the ecosystem recovers or returns to a
Restoration time
predisturbance or nominal level
Classical stability
2 Inversely proportional to the time required for ecosystem
recovery following disturbance
they require definition of a reference, or pre-disturbance condition. Secondly,
since forest ecosystems are diverse, complex and long lived, their quantification is without limit.
Westman ( 1986 ) defined inertia as 'ecosystem resistance to change under
stress' (equivalent to resistance as defined by Webster et al., 1975 ) and resilience as the 'degree, manner and pace of restoration of initial structure and
function in an ecosystem after disturbance' essentially the same concept as
that of Webster et al. ( 1975 ). Westman (1986) proposed that resilience has
components of elasticity (the rapidity of return to pre-disturbance), amplitude (the extent of displacement ), hysteresis (recovery under disturbance vs
recovery when not under stress ), and malleability (how similar the recovered
state is to the pre-disturbance state). Other workers have used some or all of
these terms in various ways (see Table 1 in Westman, 1986), and this again
is probably due to the difficulties of rigorous measurement and quantification. In my view, the simple, conceptual framework of resistance and resilience (Webster et al., 1975) provides a sufficiently complex yet comprehensible framework to aim for in studies on the recovery of forest ecosystems.
The conceptual model of Webster et al. ( 1975 ) embraced the range of ecosystems from streams and oceans to grasslands and forests. They concluded
that resistance 'results from the accumulated structure of the ecosystem' and
is associated with large biomass storage, large amounts of recycling and long
turnover times while resilience 'reflects dissipative forces inherent in the ecosystem' and is associated with low storage, fast recycling and short turnover
times. Neither term, it seems, is definitive. Harwell et al. (1977) reanalysed
the Webster et al. ( 1975 ) model and reached quite different conclusions, especially in the ranking of resistance as it is related to biomass storage.
There is, however, surprisingly little empirical work on these themes in plant
254
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
communities. Recovery after disturbance is most often assessed by indices of
c o m m u n i t y structure and species diversity (e.g. Ewel, 1983; Denslow, 1985;
Halpern, 1988; Olsen and Lamb, 1988; U n w i n et al. 1988 ) and diversity profiles (Lewis et al., 1988), and there is some, limited evidence that resistance
(sensu Webster et al., 1975) increases with increasing diversity (McNaughton, 1977; Walker, 1982; Frank and McNaughton, 1991 ). At the process level of nitrogen cycling in forests and the potential for losses after disturbance, resistance (sensu Webster et al., 1975 ) is characteristic of less fertile
sites and resilience is characteristic of more fertile sites (Vitousek et al., 1982;
Polglase et al., 1986; Weston and Attiwill, 1990).
In summary, both disturbance and recovery after disturbance are difficult
to classify, conceptually and quantitatively, with precision. Fragility is often
used descriptively, but it rarely appears in the ecological literature and appears not to have a strict ecological definition. It is implicit in all of these
terms that there is a succession leading toward a defined reference.
Disturbance, succession and climax
Ecosystem development - the northern hardwoods model
The model for ecosystem d e v e l o p m e n t following disturbance proposed by
B o r m a n n and Likens (1979a,b) is based on the rate of accumulation of biomass (living and dead). The model has four phases (Fig. 3 ). After disturI
I
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Clear Cut
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Phase
Transilion
Phase
Steady State
Phase
D
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,~
Fig. 3. Phases of ecosystem development after clear-felling a second-growth hardwood forest,
north eastern United States (from Bormann and Likens, 1979a,b, with the permission of Springer-Verlag, New York). There is no exogenous disturbance after clear-felling.
P.M. A ttiwill / Forest Ecology and Management 63 (1994) 247-300
255
bance, there is an immediate decrease in the total amount of biomass, the
reorganization phase, where increased rates of respiration and decomposition
follow increases in water availability and temperature. The rate of biomass
accumulation by the regenerating forest is then m a x i m u m during the aggradation phase, declines during the transition phase and eventually reaches an
irregula~ shifting mosaic steady-state phase. The model was developed to describe changes following clear-felling of the mixed hardwood forest at the
Hubbard Brook Experimental Forest, north-eastern USA. The forest therefore starts as even-aged and remains so to the end of the aggradation phase,
at which time it is probably described as old-age (or virgin, climax, pristine
or steady-state, depending on one's point of view; Bormann and Likens,
1979a). This even-aged forest is dominated by a few old trees per unit area;
as these trees die, gaps are created within which new vegetation develops to
become dominant, thereby creating a shifting mosaic of ages and resulting in
a decrease in total biomass averaged over the whole. Thus 'an individual plot
will never exhibit a prolonged steady-state in which biomass accumulation
will approximate zero' (Bormann and Likens, 1979a).
Ecosystem development in mountain ash (Eucalyptus regnans)
The 'northern hardwood model' (Fig. 3 ) is driven essentially by autogenic
processes; gaps are created by trees falling over or being blown over, regeneration in gaps eventually producing an all-aged forest. In many forests, however, regeneration depends wholly on disturbance. Australian eucalypts of wet
sclerophyll forest (Beadle and Costin, 1952) or tall open-forest (Specht,
1970), such as E. regnans (mountain ash), are outstanding examples of such
a dependency (Ashton, 1981 a,b). These forests regenerate naturally only after
major fire, the result being extensive tracts of even-aged forest. The forest at
250 years is dominated by perhaps 20 trees ha-1, and gaps created by death
or windblow of these trees are filled only by the relatively low understorey.
The even-aged structure of the dominant trees is maintained; multi-aged forest is restricted to no more than two or three age classes, generally in small
edge-patches where subsequent, non-lethal ground-fires have intruded. We
therefore expect that the relative decrease in biomass, averaged over the whole,
as the eucalypt tall open-forest progresses from aggrading to steady-state, will
be greater than for the northern hardwood forest (Fig. 3 ).
The pattern of water yield or streamflow from mountain ash forest with age
(Fig. 4, from Kuczera, 1987; see also Attiwill, 1991; O'Shaughnessy and Jayasuriya, 1991 ) implies a high rate of evapotranspiration in the aggrading forest and a decreasing rate in the maturing to old-age forest. A likely explanation of this pattern with age is that water use depends on sapwood area. In the
young, aggrading forest, all of the wood is sapwood. As the forest ages, the
area of sapwood decreases ( D u n n and Connor, 1991 ). The decrease in water
256
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
b
r-
:~
12
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d
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120
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160
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200
FOREST AGE (years)
Fig. 4. Relationships between water yield and stand age for Eucalyptus regnans forested catchments (from Kuczera, 1987 ). The idealized response curve is for E. regnans in the high rainfall
region of the Central Highlands of Victoria, Australia, annual rainfall in the range 1800-2000
ram. (See Attiwill, 1991. )
use of the old-aged forest (at about 160 years, Fig. 4 ) relative to the aggrading
forest is very large.
We could accept that the phases of reorganization, aggradation and transition of the northern hardwood model (Fig. 3 ) are universally applicable, but
that the steady state phase is entirely forest specific. The northern hardwoods
of New Hampshire, USA, are relatively fire-resistant, to the extent that they
have been called 'asbestos forests' (Bormann and Likens, 1979a); historically, large-scale disturbances by fire (Russell, 1983) and windthrow (Seischab and Orwig, 1991 ) were rare. There is a mixture of dominant species,
shade tolerance is high, and regeneration or release of suppressed trees follows the creation of gaps caused by endogenous disturbance. The shifting
mosaic steady state therefore involves shifts in both diversity and net productivity in time and space. The frequency of catastrophic exogenous disturbance such as large-scale winds or fire which would initiate a new reorganization phase over extensive areas is relatively low. It follows, that if an
exogenous disturbance such as clear-felling (followed by regeneration ) is imposed on such a system at a frequency which differs markedly from the frequency of catastrophic disturbance, the 'asynchronous' age structure of the
shifting mosaic steady state will be regulated.
P.M. Attiwill / Forest Ecology and Management 63 (1994) 24 7-300
257
In contrast, Eucalyptus regnans (mountain ash ) in south-eastern Australia
is a truly fire climax forest (Ashton, 1981 a,b ); its natural presence is due to a
major fire in the past. At its best development, away from ecotones, it grows
as pure, even-aged stands. It is shade intolerant, so that mortality in the aggrading phase is high. Trees never remain suppressed, to be liberated by the
creation of a gap. Seedlings do not develop in gaps, nor do trees re-grow from
coppice shoots or root suckers. The 'steady state' phase (perhaps from about
160 years, Fig. 4) is therefore neither a steady state nor a shifting mosaic; it
is a progression from domination of the site by trees to domination by understorey species of shrubs (understorey trees such as Acacia themselves being
dependent on fire for their regeneration). Given time, species characteristic
of cool temperate rainforest may become established. To a major degree,
however, this is speculation, since the frequency of catastrophic fire which
regenerates the forest is high enough to preclude the gathering of empirical
evidence. It follows that the effect of an exogenous disturbance such as clearfelling (followed by regeneration) will affect the age structure of the mountain ash forest to the extent that the frequency is related to the stochastic
frequency of catastrophic fire.
Succession and climax
Successional change in community structure and composition is often obvious, but the end-product of succession cannot be predicted with the certainty implicit in the classical view of the stable and self-perpetuating climax
that prevailed earlier this century (see discussions and reviews by White, 1979;
Bormann and Likens, 1979a; Denslow, 1980; Noble and Slatyer, 1980; Spurr
and Barnes, 1980; Oliver, 1981; Oliver and Larson, 1990; Hopkins, 1990).
Rather, there is very great variation, both in time and space, in the forces
which influence and direct the development of the structure, composition and
functioning of plant communities (Whittaker, 1953, 1956; Egler, 1954; Drury
and Nisbet, 1973; Horn, 1974; Connell and Slatyer, 1977; White, 1979; Shugart, 1984). Disturbance is one of these forces, to the extent that 'disturbance
gradients may parallel physical environmental gradients (and so ) the two may
be difficult to distinguish' (Harmon et al., 1983). Some disturbances (especially fire) have had such profound evolutionary influence on the survival
and regeneration of plant communities that periodic disturbance is essential
if diversity is to be maintained, and a delay in the recurrence of disturbance
leads to a reduction in diversity (e.g. Loucks, 1970; Whelan and Main, 1979;
Walker, 1982 ).
These various themes are encompassed by Shugart and West ( 1981 ): 'Forests often represent a tranquillity associated with the apparent changelessness
of large trees. Ironically, this tranquillity is largely a product of the human
perception of time... Forest ecosystems are dynamic entities that may not be
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P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
static in either time or space. Many of our concepts of ecosystem management
are based on the hope that if a forest is left alone it will gradually return to its
natural state. 'Succession', 'wilderness', 'virgin forests' and 'climax forests'
are all concepts that appeal to the basic notion that forest systems should approach some equilibrium state with time.'
Loucks (1970) proposed the hypothesis that m a x i m u m diversity and productivity of forest ecosystems are maintained by random periodic disturbance, a hypothesis which is directly applicable to the mountain ash forest, as
well as to other forests which depend on fire for their renewal (Bormann and
Likens, 1979a). There is no autogenically derived steady state, and the system is entirely dependent on exogenous and catastrophic disturbance for its
recycling. This disturbance, random in time, is therefore fundamental to the
stability of the ecosystem (Loucks, 1970). Loucks ended his paper in the
strongest way, concluding that the elimination of disturbance by modern humans 'will be the greatest upset of the ecosystem of all time... It is an upset
which is moving us unalterably toward decreased diversity and decreased
productivity at a time when we can least afford it, and least expect it'.
This system, in which stability is achieved through random disturbance, is
found in many plant communities including heathlands of Mediterranean regions (Specht, 1979, 1981c). Coastal heaths in south-eastern Australia include a large number of species (e.g. Allocasuarina paradoxa, Banksia marginata, Hakea sericea and Leptospermum myrsinoides) which regenerate
either (or both) vegetatively from shoots and suckers or (and) from seed
held in bradysporous fruit (Specht, 198 la,b). Like the chaparral in California (Force, 1981 ), diversity is highest immediately after disturbance by fire.
If fire is excluded from heathland, diversity of both plants and animals quickly
decreases (Force, 1981; Walker, 1982 ) and dominance is exerted by one or a
few original or invading (Burrell, 1981 ) species. The capacity of the heath to
respond to a disturbance when it eventually occurs is then greatly diminished
(Walker, 1982). Not only would it be practically difficult to regenerate the
original set of species but there would be a decrease in diversity, m a x i m u m
diversity being maintained at the historic frequency of disturbance (Denslow, 1980).
The autogenically derived steady state for the northern hardwood forest
(Fig. 3 ) is predictable for the whole, but less predictable for a space (or plot )
within the whole. Whether or not the mountain ash forest and the heathland
reach a steady state, and what this steady state will be for a given space, are
for the most part matters for conjecture. Furthermore, I suggest that steady
state is an inappropriate term for an ecological system, since it implies both
predictability and stability at a level we can neither expect nor assess.
It is apparent then that the process of succession depends both on the behaviour and responses of species (e.g. Drury and Nisbet, 1973; Westman and
O'Leary, 1986) and on the regime of disturbance. Noble and Slatyer (1980)
P.M. A ttiwill / Forest Ecology and Management 63 (1994) 247-300
2 59
identified three sets of attributes of plants which are vital to their role in a
developing plant community: ( 1 ) the method of arrival or persistence of the
species at the site during and after a disturbance; (2) the ability to establish
and grow to maturity in the developing community; and (3) the time taken
for the species to reach critical life stages. However, at the level of the plot,
the unit on which we must assess resistance and resilience, Halpern ( 1988 )
concluded that 'variation in the long-term response of communities reflects
complex interactions between species life history, disturbance intensity, and
chance, suggesting that both deterministic and stochastic factors must be considered in evaluating community stability and response to disturbance'. While
the approach through vital attributes and life histories is wholly rational, autecological data even for common understorey species is mostly limited, as
Halpern (1989) noted for Pseudotsuga forests. The problem in tropical forests is much greater. G6mez-Pompa and V~zquez-Yanes (1981 ) acknowledged the 'enormous importance' of understanding life cycles, but concluded
that: 'There exists an enormous diversity of tropical ecosystems in the world...
Of the milli.ons of species of plants and animals in the tropics, we know with
some detail (although superficially) the life cycles of only a few. Until our
knowledge increases, it will not be possible to arrive at a general concept of
succession in the tropics'.
Natural disturbances in forests
Tree fall and gap dynamics
The topic of tree fall and gap dynamics has been developed principally as
the study of autogenic changes in species composition and community structure and function at a given site (the dynamics of tree crowns of existing
dominants and their occupation of the space created by the fall ofa neighbour
has been given less attention). Gap dynamics is therefore the essence of the
shifting mosaic steady state (Fig. 3 ); it is concerned with the development of
species which require high levels of resources and grow rapidly (the terms
'large gap specialists' (Pickett, 1983 ) and 'pioneer' species (Whitmore, 1989 )
appear to be synonyms) and of species which require low levels of resources
and grow slowly ('small gap specialists' or 'non-pioneer' species). These latter, shade tolerant species vary widely in their tolerance of shade and hence
in their responses to gaps (Canham, 1989; Whitmore, 1989 ).
The foundation for tree fall and gap dynamics is 'pattern and process' (Watt,
1947 ), and much of the current work is based on Watt's defined phases, 'gap',
'building' and 'mature'. The literature to 1985 has been intensively reviewed:
by Runkle ( 1985 ) for temperate forests; by Brokaw ( 1985 ) for tropical forests; and by Veblen (1985a) for Nothofagus forests in Chile. Since that time,
a special feature on treefall gaps has been published in Ecology (Platt and
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P.M. Attiwill /Forest Ecology and Management 63 (1994) 247-300
Strong, 1989 ) and a collection of papers from a symposium was published in
Canadian Journal of Forest Research (Denslow and Spies, 1990 ).
It seems that the diversity of results support Grubb's (1977) conclusions:
tree fall creates a wide diversity of gaps, and different species which might fill
the gaps have different requirements. For example, Denslow and Spies ( 1990 )
concluded that the importance of gaps depends on the peculiarities of the
system and its component species, whether or not there are established understorey shrubs which might compete with regenerating trees, whether or not
there are established seedlings and if not, what is the pattern and composition
of seedfall?
Most of the work on tree fall and gap dynamics has been in tropical forests
and in temperate forests of the USA. The mean proportion of mature temperate forest opened to gaps each year is within the range 0.005-0.02 (Runkle,
1985 ) so that the turnover time for the whole forest is 50-200 years. Data for
tropical forest are within this range (e.g. 80-138 years (Hartshorn, 1978 );
60-135 years (Brokaw, 1985, with the exclusion of one much longer time);
137 years (Lang and Knight, 1983 ); 47 years (Martinez-Ramos et al., 1988 ) ).
Both frequency and intensity of treefall in tropical rainforest vary widely with
age, the proportion opened to gaps in individual years varying from < 0.01 to
about 0.09 (Martinez-Ramos et al., 1988). This proportion is significantly
and positively correlated with annual rainfall and Martinez-Ramos et al.
( 1988 ) suggest that it is this temporal and spatial variance of treefall disturbance which gives greater diversity in tropical forest than in temperate forest.
Strong (1977) c o m m e n t e d on the increase in the number of vines and epiphytes in wet lowland forest toward the equator; he suggested that the weight
of vines and epiphytes in the canopy would lead to greater disturbance, a sort
o f ' d o m i n o ' effect. As far as I am aware, however, there has been no detailed
analysis of the relationship between species diversity in tropical forest and
frequency and intensity of treefall along a latitudinal gradient.
The size of gaps ranges typically from 50 to 200 m 2 for single tree falls
(Hartshorn, 1978; Brokaw, 1985; Arriaga, 1988; Howe, 1990; Runkle, 1990)
up to 300-500 m 2 for multi-tree falls (Canham et al., 1990). The light regime
within the gap depends on tree height (Canham et al., 1990) and on the nature of canopy damage caused by the fall of trees and limbs (Lawton, 1990 ).
Light penetration into the gap decreases with decreasing latitude but even in
gaps as large as 1000 m 2, direct sun reaches the forest floor for less than 4 h
day-1 (Canham et al., 1990). The light regime in gaps is therefore more heterogeneous and changes more significantly with size in temperate forest than
in tropical forest (Poulson and Platt, 1989). Poulson and Platt (1989) proposed that, because of this heterogeneity and variability, variation in growth
rates of individual species will be greater in temperate forest than in tropical
forest. This relationship has been developed in a general model (Fig. 5, from
Howe, 1990) in which size of the gap has increasing influence on the availa-
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P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
1O0
.0~0~0 ~ 0 d~
~O ~O~O~
-75
•
~-~
"'em~
•
o~O ~O
/
/
50
~o
/
/
•
•
~o
d=10
~o
d:14
•~•
o~
=
0
d=6
/
•
•
•
•
/ ./"
•
/
I
I
I
I
I
l0
20
30
40
50
Gap Radius (m)
Fig. 5. Habitat available to colonizing tree species as a function of gap size, represented by the
radius of a circular gap. The variable d is the distance (m) into the gap required by different
species to avoid suppression. Shade-tolerant species have smaller values of d than light-demanding species (from Howe, 1990, with the permission of Oikos).
bility of space as shade tolerance decreases. Thus shade-tolerant species establish more densely in small gaps than in large gaps in tropical cloud forest
in Costa Rica, while shade-intolerant species establish more densely in large
gaps (Lawton and Putz, 1988 ), and in tropical lowland forest on B arro Colorado Island, gap size has a lesser and longer lasting effect on species composition of shade-tolerant than shade-intolerant species (Brokaw and Scheiner,
1989).
Most studies of gap dynamics in tropical forest have been based in Middle
America (Uhl et al., 1988b) and have been mainly concerned with post-gap
regeneration in the tropical environment where seedfall is plentiful and diverse and the soil seedbank is rich (Young et al., 1987; Brandani et al., 1988;
Alvarez-Buylla and Martinez-Ramos, 1990; Denslow and G6mez Diaz, 1990).
In contrast, however, Uhl et al. (1988b) found that advanced growth of nonpioneer species rather than post-gap regeneration from seed was wholly dominant (advanced growth accounted for 97% of trees in single-treefall gaps) in
gaps in Amazonian rainforest at San Carlos de Rio Negro. While gap size
markedly affected rates of growth, it had little effect on plant density and
mortality. Thus for the Amazonian forest, 'the abundance of advance regeneration and (its) high survivorship.., minimizes size and microhabitat effects' (Uhl et al., 1988b).
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It is clear that there are many strategies involved in gap dynamics. In tropical cloud forest at Costa Rica, seedlings germinating from buried seed and
developing from epiphytic seedlings in the crowns of fallen trees are more
important in gap dynamics than are seedlings germinating from post-gap
seedfall (Lawton and Putz, 1988). Sprouters appear to be as important as
seeders in gap dynamics in beech (Fagus crenata) forest in Japan (Hara,
1987) and Putz and Brokaw (1989) found a strong survival of sprouts from
broken trees in tropical forest at Barro Colorado Island. While sprouts were
produced by a number of species after treefall in beech-hemlock forest in
eastern USA, their survival was poor (Peterson and Pickett, 1991 ).
A feature of treefall in temperate forest is soil disturbance in the form of
pits and mounds which may last for decades or centuries (e.g. Hutnik, 1952;
Stephens, 1956; Raup, 1981; review by Schaetzl et al., 1989) and which may
be of significance in soil development (Armson and Fessenden, 1973 ). This
microsite variation associated with treefall is of considerable importance in
maintaining species diversity in temperate forest (Hutnik, 1952; Nakashizuka, 1989; Peterson et al., 1990) and boreal forest (Jonsson and Esseen,
1990). In contrast, mounds and pits formed by treefall in tropical forests level
out rapidly and are therefore not prominent (Putz, 1983 ). In the Amazonian
tropical forest at San Carlos de Rio Negro, microhabitat within gaps had no
measurable effect either on fine root biomass (Sanford, 1990) or on nutrient
loss (Uhl et al., 1988b). Understorey shrubs planted across gaps in tropical
forest at Costa Rica responded to the changing light environment, but not to
the addition of fertilizers (Denslow et al., 1990). It seems, therefore, that the
effect of microsite in tropical forest is less important than in temperate forest,
in terms both of soil upheaval and of light penetration (Canham et al., 1990 ).
I have not seen detailed studies on the effects of disturbance due to soil upheaval on species diversity in tropical forest apart from Barro Colorado Island where Putz (1983) reported that the uprooting of buried seed and the
baring of mineral soil favoured the establishment of pioneer species.
There are many other examples of gap creation having an important influence on species composition, the co-existence of Abies lasiocarpa and Picea
engelmannii in subalpine forest in the Rocky Mountains, Colorado (Veblen,
1986); the persistence of Picea sitchensis in coastal forests of northwestern
USA where Tsuga heterophylla is the putative climax (Taylor, 1990 ) and the
regeneration of the putative climax Tsuga canadensis in forests of Pseudotsuga menziesii (Spies et al., 1990) in western Oregon and Washington (to
choose but a few recent examples). Pseudotsuga menziesii regenerates only
on mineral soil and so it requires large-scale disturbance, a response which is
typical of many species which grow in even-aged stands (e.g. Eucalyptus reghans in Australia (Ashton, 1981 a); Nothofagus spp. in South America (Ve-
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
263
blen, 1989) ). For these forests, the relevance of gap dynamics depends on
'the kind and heterogeneity of seedbeds, structures and resources' (Spies et
al., 1990).
In summary, gaps are a fundamental part of the life of forests. The mechanisms of gap filling and the responses of growth and species composition vary
widely with forest type and geography. Lieberman et al. ( 1989 ) warn of the
dangers of the gap paradigm, of treating gaps as holes in an otherwise homogeneous forest: 'considering forests as a Swiss cheese of gaps and non-gaps
does not even begin to do justice to the daunting complexity of real forests.
In fact, non-gaps are as heterogeneous as gaps'.
Lieberman et al. ( 1989 ) place their emphasis on the forest canopy, so that
the response of individual species can be tested against the continuum of canopy closure and thereby be extrapolated to predict behaviour over the range
of canopy openings. Brokaw and Scheiner ( 1989 ) developed a similar theme:
'More work is needed on dynamic processes determining the composition and
structure of the understorey beneath a closed canopy. This 'subcommunity'
covers much more area than do gaps (and) contains more individuals... Creation of a gap in the canopy acts on patterns already established in the understorey of the closed phase'.
Large-scale disturbance by wind
Disturbance by wind occurs over a continuum, and the nature of disturbance may therefore be difficult to characterize. The blow-down or snapping
of an individual tree is at one extreme, which might be defined as wholly
endogenous (or the result of autogenic processes). At the other extreme, catastrophic hurricanes may be defined as wholly exogenous (an allogenic force
on the community); on the other hand, there is a current view (see Boucher,
1990 ) that the composition of some rainforests is maintained by catastrophic
disturbance, that periodic disturbances of rainforest preclude an equilibrium
condition ever being reached.
The regeneration of balsam fir (Abies balsarnea ) forest which grows at high
altitudes in the north-eastern USA is a nice example of the difficulty of characterization. The trees have a life-span of 60-80 years and at senescence, local
disturbance of the weakened trees expands windward as a wave (Sprugel and
Bormann, 1981 ). Regeneration of the disturbed areas follows, resulting in
wind-induced waves of death and regeneration which move through the forest at a rate of 0.75-3.3 m year-~ in a windward direction. The maintenance
of this forest therefore depends on disturbance by wind (Sprugel and Bormann, 1981 ).
In a sub-alpine forest of the Rocky Mountains, Colorado, disturbance by
hurricane-force winds has actually accelerated succession (Veblen et al.,
1989). The forest was dominated by lodgepole pine (Pinus contorta var. latifolia ) which had regenerated after fire with a sub-canopy ofAbies lasiocarpa
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and Picea engelmannii. Blow-down of the pine has released the sub-canopy
species, thereby accelerating successional development.
There has been a long and sustained interest in the role of disturbance in
north-central and north-eastern USA (see, for example, early accounts of disturbance in the pre-settlement forest in Stearns (1949) and Mclntosh
(1961) ). Return times for catastrophic windthrows ( > 1 ha) in pre-settlement forests are in the range 1000-2000 years (as judged from original survey records; Canham and Loucks ( 1984 ); Seischab and Orwig ( 1991 ) ); most
windthrows were in the size range 1-50 ha with a lesser number in the size
range 50-1000 ha (from Canham and Loucks, 1984; Seischab and Orwig,
1991 ). Hurricanes strike central New England every 20-40 years, and major,
catastrophic storms occur every 100-150 years (Foster 1988a,b).
The area devastated by severe winds and hurricanes can be very large. A
severe windstorm (gusts in excess of 150 km h - ~) in north-western Ontario,
Canada, uprooted or snapped up to 50% of trees in valley bottoms and 100%
of trees on hilltops over an area of 350 km 2 (Schindler et al., 1980) and the
1938 hurricane in New England, USA (gusts in excess of 200 km h - l ) , blew
down more than half the trees on 43% of the area of the Harvard forest (Spurr,
1956). Damage and subsequent recovery from the 1938 hurricane to the
Hubbard Brook Experimental Forest in New Hampshire, USA, was assessed
from 1942 and 1978 aerial photographs (Peart et al., 1992). Damage was
patchy, and the most heavily damaged area (72% of the area open) had substantially recovered 40 years later (9% open); '40 years of canopy development effectively masked the patchy effects of hurricane damage' (Peart et al.,
1992).
The susceptibility of a forest to damage from severe windstorms depends
on species, on stand age and structure (older and more open forests are more
susceptible than younger, denser forests), on the characteristics of the site
including slope and soil depth, on the physiography of the region, on the surrounding forest structure, and on the nature of the prevailing storm (Canham
and Loucks, 1984; Foster, 1988a,b; Webb, 1989).
Canham and Loucks (1984) painted the broad picture of pre-settlement
disturbance in forests of the northern USA, low levels of disturbance giving
more or less steady state communities in the north-east (the northern hardwoods model of Bormann and Likens, 1979a,b, Fig. 3), intermediate levels
of disturbance by severe winds, thunderstorms and fire in the less mesic forests (Brewer and Merritt, 1978; Frelich and Lorimer, 1991 ) toward the northcentral region, and catastrophic disturbance in the fire-dependent coniferous
forests in the north-west. For the central part of this scenario, the mixed forests of northern Wisconsin, Stearns (1949) wrote: 'There are many instances
of natural catastrophe noted in the surveyors' field notes and they are still
occurring in the residual stands. Windfall alone or with the other agents of
mortality, fire, drought, glaze storm, insect or fungus infestation and senes-
P.M. A ttiwill / Forest Ecology and Management 63 (1994) 247-300
26 5
cence keeps the forest in a constant state of flux which results in continuous
variation in composition, both in space and time, within the limits of the
species present'.
In tropical latitudes, catastrophic disturbance by windstorm is frequent,
with trees being extensively damaged either by uprooting or snapping (Putz
et al., 1983). Tropical cyclone Winifred (1 February 1986) passed through
the coastal area around Innisfail, Queensland, with winds peaking at 185 km
h - l ; some 1500-2000 km 2 of rainforest was damaged, an area of 300-700
km 2 being moderately to severely (boles or crowns of most trees broken,
smashed or windthrown) damaged. Within 40 days, structural resilience was
demonstrated by 'an overwhelming mode of short term recovery'; virtually
all tree species were resprouting from leaf shoots or coppice shoots or both
(Unwin et al., 1988). Webb (1958) concluded that the rainforests of north
Queensland lowlands and foothills suffer severe or general damage every 340 years: 'Present observations in north Queensland suggest that cyclones are
a potent ecological factor which regularly upsets forest equilibrium, with farreaching consequences for the regeneration, suppression and reproduction of
species. The heterogeneity of the tropical biological environment, preserved
by cyclones, has important implications for speciation' (Webb, 1958 ).
This view is re-stated for the tropical rainforests of Puerto Rico in a study
of the effects of Hurricane Hugo, 1989: 'Hurricanes strike Puerto Rico every
10-30 years... Such a high frequency of hurricanes overrides other ecological
factors and creates a cyclic steady state in the forest ecosystem... A stable,
steady state ecosystem in which there is no net change in total biomass over
time is not possible' (Basnet et al., 1992; and compare with the model for
northern hardwoods, Fig. 3 ).
Hurricane Hugo was the largest hurricane to strike the Luquillo Experimental Forest in Puerto Rico since 1932. Basnet et al. (1992) report that
every tree was damaged, and some 85% were severely damaged, the largest
trees being the most susceptible to damage. It is interesting that there was
greater damage in the valleys than on ridges and slopes; the possible explanations are that the valley soils are deep and poorly drained and hence unstable, that rocks on slopes provide anchorage for trees, and that valley trees
suffer damage from tree and branchfall on the slopes (Basnet et al., 1992 ).
Hurricane David (August 1979) passed through the Carribean island of
Dominica at a speed of 22.5 km h-1, with wind velocities at its centre reaching 241 k m h -~ (Lugo et al., 1983). Some quarter of a million hectares of
tropical forest, one-third of the island, was heavily damaged. Regrowth of
seedlings and saplings was rapid after the hurricane, with densities up to 20
plants m -2 within 1 month, and the number of species in the regrowth was
greater than in forest that had not been damaged. Recovery time, to predisturbance condition, was estimated at 50 + years (Lugo et al., 1983).
In 1988, Hurricane Joan struck the east coast of Nicaragua, and winds up
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to 250 km h-1 toppled some half a million hectares of rainforest. The broadleaved forest before the hurricane was very rich, with a total of 79 species of
dicotyledonous plants and the ten most abundant species making up only 43%
of the total (Vandermeer et al., 1990; Yih et al., 1991 ). After the hurricane,
the forest regenerated profusely and predominantly by re-sprouting from the
damaged trees, together with seedlings which had probably established before
the hurricane. It is of great significance that there was no reduction in the
diversity of plant species; that is, regeneration both of sprouts and seedlings
was characteristic of primary forest, rather than of secondary pioneers. Such
a result would not be anticipated from the conventional sequence coming from
gap dynamics (Fig. 5 ).
Yih et al. ( 1991 ) conclude that: 'If disturbances like Hurricane Joan have
occurred repeatedly in the history of Central American rainforests (on a scale
of centuries ), we can no longer view them as stable equilibrium communities.
As with many other ecosystems, their structure depends on patterns of natural damage and recovery. The appropriate metaphor for natural ecosystems
is not eternal constancy, but rather cycles of death and resurrection'.
Lugo et al. ( 1983 ) developed a similar theme on the problems of interpreting succession, climax and the steady state: 'With such high frequencies, intense hurricanes could easily override the usual ecological factors that impinge on an ecosystem. Thus stable steady-state ecosystems probably cannot
develop in hurricane-prone regions'.
It is worth recalling the views of some earlier writers. In 1958, Webb strongly
questioned the concept of a 'stable tropical forest climax'. Webb quoted Richards (1952) who outlined the mosaic theory of regeneration developed in
1938 by A. Aubrrville. The theory holds that species composition at a given
site of mixed tropical forest is variable both in space and time. No set of species is more stable than is any other, and none is in permanent equilibrium
with the environment; the forest as a whole is a constantly changing mosaic.
Richards gives a translation from Aubrrville; it is a delightful passage to include here: 'To borrow a comparison from algebra, one could say that the
conditions of equilibrium between the determining factors of the habitat and
the numerous independent variables constituted by the characteristic species
of the community form a number of equations much fewer than the number
of independent variables. All kinds of solutions are thus possible satisfying
the equations of equilibrium' (Richards, 1952, pp. 49-50 ).
An interesting feature of rainforest and its response to major disturbance is
gregarious flowering (Richards, 1952 ). Hopkins and Graham ( 1987 ) counted
54 species of shrubs and trees flowering a m o n t h after cyclone damage to tropical lowland rainforest in northern Queensland. Twelve of these 54 species
had already flowered and fruited before the cyclone. Levey (1988) noted a
higher intensity of fruit production in gaps in lowland rainforest in Costa Rica,
the intensity increasing with gap size; it is possible that this too was due to
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
267
gregarious flowering. Lugo et al. (1983) noted 'conspicuous' flowering of
Chimarrhis cymosa 1 m o n t h after very heavy hurricane damage to tropical
forest on the island of Dominica. The coincidence of regenerative space and
an enhanced seed supply due to gregarious flowering may be an important
component of response of rainforest following disturbance (Hopkins and
Graham, 1987 ).
Large-scale disturbance by fire
Disturbance by fire in forests of North America
Spurr and Barnes (1980) wrote that 'fire is the dominant fact of forest history': 'The great majority of the forests of the world - excepting only the perpetually wet rain forest, such as that of southeastern Alaska, the coast of
northwestern Europe and the wettest belts of the tropics - have been burned
over at more or less frequent intervals for many thousands of years. Even
under present-day conditions, marked by a great awareness of forest fires, the
separation of forest tracts by intervening tracts of farmland and settlement,
and the crossing of the forest by many roads and trails, fire continues to be a
major disturbing factor in much of the North American forest' (pp. 421-422 ).
Over the past 50 years or so, the recognition of the importance of fire as an
integral component of forest ecosystems has been increasing, rapidly so over
the past 20 years. The role of fire in coniferous forests of North America is
well documented: for the redwood (Sequoia sempervirens) and the giant sequoia (Sequoiadendron giganteum ) in California (e.g. Kilgore, 1973; Weaver,
1974; Kilgore and Taylor, 1979 ); for ponderosa pine (Pinus ponderosa ) and
for Douglas-fir (Pseudotsuga menziesii) in western and north-western USA
(e.g. Weaver, 1951, 1974; Cooper, 1960, 1961; Agee, 1991 ); for longleafpine
(Pinus palustris) in south-eastern USA (e.g. Heyward, 1939; Garren, 1943;
Komarek, 1974); for pitch pine (Pinus rigida ) communities in northeastern
USA (Forman and Boerner, 1981; Seischab and Bernard, 1991 ) to choose
but a few examples.
In the following sections, the recent literature on disturbance by fire of selected forest communities in North America and of Mediterranean-type communities is reviewed. Disturbance by fire of eucalypt forests in Australia will
be reviewed in a second paper (Attiwill, 1994).
Northern boreal forest. Boreal coniferous forest of fir-pine-spruce in northern America forms a band as wide as 1000 km (Larsen, 1980; Van Cleve and
Dyrness, 1983 ) from Newfoundland to Alaska, a vast area greater than half
the forest area of North America (Heinselman, 1981 ). There is a high frequency of lightning strikes, with fire return times of 100-300 years (Rowe
and Scotter, 1973 ). In the main boreal forest, fires are high-intensity crown
fires or severe surface fires of large size, often greater than 10 000 ha and
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sometimes greater than 40 000 ha (Heinselman, 1981 ) or 200 000 ha (Dyrness et al., 1986 ). In extreme fire years, the total area burned in Alaska ranges
from 1 million to 2 million hectares (Dyrness et al., 1986 ). The main boreal
zone is a typical fire-dominated community (Zasada, 1986 ). The vegetation
is a mosaic of even aged patches of forest; there are no communities at steadystate, no unidirectional succession of c o m m u n i t y development, diversity decreases with age, and the random perturbation by fire restarts succession, rejuvenates the growing stock and reprimes the cycling of carbon and nitrogen
(Dix and Swan, 1971; Shaft and Yarranton, 1973; Heinselman, 1981, Bonan,
1990).
The dominance of fire in boreal forest is well-illustrated by a study of forest
in northern Sweden dominated by Picea abies, Pinus sylvestris, Populus tremula and Betula spp. with a ground layer of Vaccinium spp. and Calluna vulgaris. Using tree-ring chronology up to a m a x i m u m age of 600 years, Zackrisson (1977) showed that the mean return time of fire has doubled to 155 years
since fire suppression started in the 19th century. Zackrisson (1977 ) argues
that: 'The vegetational diversity maintained by forest fires in the past was
probably a pre-requisite for the long-term stability of boreal forest ecosystems
as implied by paleoecological records... In the past the forest was in a state of
continuous change, and climax stages in the classical sense can hardly have
existed anywhere... Due to efficient fire suppression during the past two centuries, fire is no longer a rejuvenating factor in the forests'.
Black spruce (Picea mariana) forms extensive, almost monospecific stands
in north-eastern Canada. The age of the trees increases along a northward
gradient, this increase being associated with a decrease in fire return time
(Sirois and Payette, 1989). Black spruce has semi-serotinous cones and is
dependent on fire for regeneration (Dyrness et al., 1986 ). Seed germination
and seedling survival is almost exclusive to severely burned sites where the
mineral soil has been bared (Zasada et al., 1983 ).
Sirois and Payette ( 1989 ) cite evidence that treeless areas in the sub-arctic
boreal-tundra zone originated from fire-induced deforestation over the last
3000 years. A number of destructive fires can therefore lead to exclusion of
trees in the subarctic zone, and present evidence suggests that tundra is expanding into the upper boreal zone (Sirois and Payette, 1991 ).
In the northern areas of the boreal zone of North America, fires have not
been (and are not capable of being) controlled to the extent that they have
been in the south. At the southern transition of the boreal zone, 'near boreal
forest' (e.g. in the Boundary Waters Canoe Area in northern Minnesota,
Heinselman, 1973), the mean return time of fire over the past 1000 years is
60-70 years with a range o f 2 0 - 1 0 0 years (Swain, 1973 ). Heinselman ( 1973 )
writes of this forest: 'In much ecological literature as well as the popular press,
fire has been viewed as unnatural, man-caused, and a 'disturbance' somehow
not part of the ecosystem... It must now be seen as just a perturbation within
P.M. A ttiwill / Forest Ecology and Management 63 (1994) 247-300
2 69
the system... (as) an essential factor in maintaining.., long-term stability and
diversity'.
The stultifying effects of earlier ecological concepts of succession and climax are evident in the literature on boreal forest. Larsen (1980) concluded
that each successional pathway may have its own development and progression according to the environmental conditions it creates; 'the fortunes of
succession.., make most hypotheses very general indeed'. Larsen's (1980)
comments on the role of fire are most relevant: 'Fire is without question the
most important cause of the initiation of pioneer stages of succession. It is so
ubiquitous that Dix and Swan (1971) could make no conjecture as to the
nature of the forest that might develop were the area free of fire for a protracted period of time. They point out: 'For Candle Lake forest as a whole,
succession does not seem to be important. It seems more important that the
area has probably undergone an infinite number of vegetational readjustments.' They cite Rowe ( 1961 ) and confirm his opinion that the western boreal forest is a disturbance forest, one that does not fit classic concepts: 'In
view of these observations, any attempt to fit the vegetation into the mold of
a climax concept would be unreal and, in our opinion, unjustified.' (Larsen,
1980, p. 317).
Heinselman (1973) and others (e.g. Viereck, 1973) throw out the challenge for the introduction of a programme of prescribed fires and monitored
lightning fires aimed at restoring and maintaining community structure and
diversity. Rowe and Scotter (1973) placed the problem in context for the
land manager: 'The management of national parks, nature reserves, and wilderness areas poses many questions about the use of fire. The near exclusion
of wildfires in such places has had profound effects. If the major goal of such
areas is to perpetuate samples of as many landscapes as possible - with the
recognition that fire is an inseparable part and natural agent in the ecology of
many ecosystems - then land managers must 'unsell' the false impression that
all fires are bad and be prepared to use both prescribed fires and natural lightning fires in landscape management'.
Oak-pine forests of eastern USA. The history and development of forests in
eastern USA dominated by oak (Quercus alba and Quercus velutina) has been
documented by Abrams ( 1992 ) and Abrams and Nowacki ( 1992 ). Oak forests occupied much of the country (except for the north-east), and a change
from pine (Pinus strobus) domination to oak domination some 9000-10 000
years ago is associated with an increase in charcoal abundance. Abrams ( 1992 )
suggests that fire return times of 50-100 years promoted dominance and stability of oak, and that 'regions of eastern North America that lacked oak domination before European settlement probably burned too frequently (tallgrass
prairie) or too infrequently (hemlock-northern hardwoods) for oak to prosper'. Abrams concludes: 'Fire played a vital role in maintaining oak domi-
270
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
nance before European settlement. If in the current oak forest factors antagonistic to oak regeneration, such as a lack of fire, persist into the twenty-first
century, a retrenchment of oak dominance would seem inevitable'.
Subalpineforests of north-western USA. The vegetation sequence of mountain ranges of the inland north-western USA (e.g. the Grand Teton National
Park, Loope and Gruell ( 1973 ) and see detailed example, Fig. 6, from Romme
and Knight ( 1981 ) ) is Douglas-fir (Pseudotsuga menziesii) at lower elevations (to about 2000 m or more on south-facing slopes) followed by lodgepole pine (Pinus contorta) at intermediate elevations ( 1800-2400 m), both
occurring as essentially even-aged stands from fire origin. Subalpine fir (Abies
lasiocarpa) grows from 1850 m to the timberline at about 3050 m, and Engelmann spruce (Picea engelmannii) mixes with fir in the upper part of this
range. The evidence for the role of fire in maintaining the diversity of these
communities over thousands of years has been clearly recognized. For example, Taylor ( 1973 ) showed that numbers of species of plants, birds and small
mammals, and number of individuals of birds and small mammals, was max-
3000
PINE-SPRUCE-FIR
2900
28O0
SPRUCE-FIR
FOROR
EST
F
2700
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~"
[.I.]
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k~
EE , R J
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LODGEPOLE
PINE FOREST
N
FOREST
2600
2500
~
MEADOW
COMMUN,T
~
Z
O
WILLO~
-ALDER MESIC
COMM- ASPEN
UNITY FOREST
2400
DOUGI.AS /
FIR
FOREST
DOUGLAS
FIR
FOREST
SAGEBRUSH
COMMUNITY
2300
2250
NORTll
SLOPE
MOST MESIC 9
TOPOGRAPHIC
EW
SOUTH
SLOPE
SLOPE
IRIDGETOPI
~ MOST XERIC
POSITION
Fig. 6. Vegetation of the Savage Run Watershed (Medicine Bow Mountains, Wyoming) in relation to elevation and topography (from Romme and Knight, 1981, with the permission of the
Ecological Society of America).
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
271
i m u m in the first 25 years and decreased thereafter. In alpine krummholz and
heath communities of the North Cascades, Washington, fire results in a distinctive and persistent flora which 'makes a substantial contribution to the
diversity of alpine areas' (Douglas and Ballard, 1971 ). Taylor ( 1973 ) concluded that: 'It is necessary that fire be accepted as one of the natural and
important environmental factors of the Park (Yellowstone National Park).
Older lodgepole pine forests must be periodically burned to perpetuate natural plant and animal c o m m u n i t y life cycles and to promote greater biotic diversity. This could be done by allowing certain lightning-caused fires to burn
and by controlling only those that endanger Park facilities and human lives.
A system of zoning that would allow quick decisions on which fires to control
and which to allow to undergo natural development'.
A policy of halting fire suppression and allowing natural fires to run their
course in selected areas of Yellowstone National Park was made in 1972
(Houston, 1973; Habeck and Mutch, 1973; Loope and Gruell, 1973) and
there is now intensive work on formulating strategies to manage fire as a natural ecosystem process (Keane et al., 1990 ). This is a difficult task since both
fuel loadings to support fire and regeneration strategies after fire vary greatly
in the steep mountainous region and are therefore to a large extent unpredictable (e.g. Lunan and Habeck, 1973; Agee and Smith, 1984). An example of
variability comes from a study of western larch (Larix occidentalis) and lodgepole pine forests in the Glacier National Park, Montana (Barrett et al.,
1991 ). There are two distinct fire regimes: short intervals of 25-75 years between fires of variable intensity on dry sites and longer intervals of 120-350
years between stand-replacing fires on moist sites. Fire suppression over the
last 50 years has resulted in a major reduction in fire frequency on dry sites
but there has been no change on moist sites. However, 50 years is a short time
within the primeval fire interval for stand-replacing fires, and Barrett et al.
( 1991 ) cite evidence to suggest that, given sufficient time for the accumulation of fire-supporting fuels, climate is the ultimate controlling factor in the
occurrence of large, stand-replacing fires. Further to the north in the Glacier
National Park, British Columbia, weather patterns produce the combination
of intense drying of fuel, high winds and lightning and the potential for fire
and its rapid spread is therefore very great. From an assessment of historical
fire frequencies, Johnson et al. (1990) concluded that neither fire suppression policies nor the accidental starting of fires following European settlement have significantly altered the fire frequency. Barrett et al. ( 1991 ) conclude: '(Glacier National Park) contains one of North America's last expanses
of unlogged larch-lodgepole pine forest, and large fires will continue to be
important perturbations. Still, managers have tried to extinguish most fires
to protect park visitors and facilities, evidently with varying degrees of success. Fire history reveals the inherent irony and futility of this approach, and
consequently the need for new management strategies'.
272
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
The complexities of management strategies are shown by Romme's ( 1982 )
reconstructions of vegetation mosaics in a 73 km 2 watershed of Yellowstone
National Park, Wyoming (Fig. 7 ). The diversity of landscape was maximum
in the early 1800s following major fires; total exclusion of fire is predicted to
increase evenness and to reduce diversity (Fig. 7 ). A carefully planned programme of prescribed burning may increase diversity but it may also have
negative effects and Romme ( 1982 ) suggests 'that such an experiment would
be wholly inappropriate in a natural area such as Yellowstone National Park'.
Romme ( 1982 ) concluded that: 'The current fire management plan probably
will be effective in maintaining natural landscape patterns in the subalpine
zone if most lightning-caused fires are allowed to burn naturally, including
the very large fires... Managers should expect, and allow, an occasional fire
(a) Richness
L(R)
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81/"
EXCLUSION
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.
.
.
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70
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ted
e¢
b-(c) Patchiness
Z
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Z
O
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15-
NATURAL
14
% Watershed Area Burned Under Natural Fire Regime
4(/-
20-
i 9
17i5
1834
19115
1949
1778 1798 18'18 1838 1858 18~]8 1898 19'18 1938 1958 1978
YEAR
Fig. 7. Reconstructions of diversity in a watershed of subalpine forest, Yellowstone National
Park in relation to natural and imposed fire regimes (from Romme, 1982, with the permission
of the Ecological Society of America). Relative richness, relative evenness and relative patchiness are reconstructed since 1778 under three fire regimes: (a) natural, all fires allowed to burn;
(b) selective, large fires excluded, small fires allowed to burn; and (c) exclusion, total fire
exclusion.
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
273
covering many square kilometres; these are the fires that will exert a predominant influence on landscape composition and diversity for many decades to
follow. Such large fires should not be viewed as unusual events occurring because of unusually high fuel accumulations'.
These views were put fiercely to the test in 1988 when 562 000 ha of the
greater Yellowstone area, including 396 000 ha within Yellowstone National
Park, were burned (SchuUery, 1989 ). However, Romme and Despain ( 1989 )
consider that, since the reconstructed fire history shows fires of similar magnitude in the early 1700s, the 1988 fires should not be viewed as an abnormal
event. But the views of the American people in general and ecologists and
foresters in particular vary. Elfring (1989) quotes Romme as saying that
Americans think of parks as static curiosities rather than as dynamic ecosystems, and concludes that people see the role of national parks as 'preserving
pretty places' rather than 'protecting remnants of functioning natural systems'.
On this theme, the use of prescribed fires to reduce the hazards of wildfire,
for regeneration, for enhancement of wildlife habitat, for insect and disease
control and for conservation of diversity of forested ecosystems in Canada
has been recently reviewed (Weber and Taylor, 1992). They concluded that
prescribed burning is both ecologically compatible and cost effective; it requires a 'vigorous public awareness campaign', particularly where it is to be
used in parks and reserves, so that people are more aware of the dynamic
nature of ecology and of the ecological goals of management.
Clark ( 1990a,b ) investigated the highly complex problems of determining
fire regimes and the effects of fire suppression in the face of climatic change.
The evidence is that fire frequency would have increased by 20-40% this century because the climate has become warmer and drier. Indeed, 'direct effects
of climate change on forest production and decomposition may well be
swamped by the more drastic effects of climate change on fire regime' (Clark,
1990a). However, fire suppression in northwestern Minnesota has resulted
in significant increases in organic matter storage in litter and coarse woody
debris, as well as in significant compositional changes.
Disturbance by fire in Mediterranean-type ecosystems
Mediterranean-type ecosystems (sensu DiCastri, 1981; Margaris, 1981;
Trabaud, 1981; Specht and Moll, 1983) are found in the Mediterranean
(monte bajo and tomillares in Spain, maquis and garrigue in France, macchia
and gariga in Italy, choresh and batha in Israel, xerovuni and phrygana in
Greece, gatha nabati in Syria and Lebanon), in California (chaparral and
coastal sage ), in Chile (espinal and matorral ), in South Africa (renosterveld,
strandveld and fynbos) and in Australia (mallee or mallee open-scrub and
heathland). The fire dependence of Mediterranean-type vegetation has long
been recognized (e.g. Biswell, 1974 ). For the Mediterranean region itself, Na-
274
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
veh ( 1975 ) wrote less than 20 years ago: 'Due to the destructive combination
of overgrazing and land abuse with burning, wildland fires have been regarded in the Mediterranean Region as a wholly condemnable, man-made
element. This might be the reason for the few, systematic studies of fire effects
and the neglect of their evolutionary role in Mediterranean vegetation. In previous studies, fire has been mentioned only briefly as a destructive factor or
in connection with man-induced regression stages, leading away from the
'maquiclimax'. Only (recently has fire been recognized) as one of the major
ecological factors which shaped the Mediterranean landscape and affected its
mosaic-like pattern of regeneration and degradation stages' (see also Naveh,
1974).
This recognition is now extensively documented; fires at frequencies in the
general ranges 10-25 years or 20-50 years are fundamental to the maintenance of composition and structure of Mediterranean-type shrublands (e.g.
Keeley, 1986) and heathlands (e.g. see papers in Specht 198 l c). Mediterranean-type ecosystems regenerate rapidly after fire (Margaris, 1981; Hanes,
1971 ). In Australian heathlands, for example (Specht, 1981a,b) plants are
characterized by sclerophyllous and often resiniferous leaves, heat-resistant
bark, lignotubers, protected buds which can produce epicormic shoots, and
woody (bradysporous) fruits which retain viable seed for many years. In wetter communities, plants regenerate mainly from sprouts whereas in more open
drier communities, plants established from seed can withstand the competition from the sprouters (the strategies of maquis (or wetter) ecosystems and
phryganic (or drier) ecosystems, Margaris ( 1981 ) ). The oils and resins, together with an accumulation of litter, makes the vegetation extremely freprone, but regenerative adaptations to fire result 'in a wealth of regenerating
plants and seedlings ... thus repairing the catastrophic effect of the fire' (Specht
198 l a). In modelling these responses for the Californian chaparral, Hilbert
(1987 ) suggests that resprouting might be the dominant strategy where the
fire interval is short ( < 5 years) or long ( > 60 years), and seed production
and seed survival would be dominant at intermediate intervals, and both
would be equal in the frequency range 40-60 years.
Change in diversity of plant species in Mediterranean-type ecosystems following fire (Fig. 8, from Kruger, 1983 ) is not dramatic. Diversity is generally
greatest in the first few years after fire (see also Keeley et al., 1981 ) and then
declines slowly (Fig. 8). Diversity of small mammals in Australian heathland, however, appears to be entirely dependent on a mosaic of disturbance:
'The members of the small mammal community have evolved by adapting to
changes in the vegetation, which provides their food and shelter, in the same
way that the heath vegetation has evolved in response to specific fire regimes... The heathlands with the highest small-mammal diversity seem to be
in areas where this regime has not been greatly altered and other man-made
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
27 5
1O0
"'
.~
"~x Protea fynbos
80
- - ~ ~ ~
I
Heathland, Wilson's Promontory
o
40
o- . . . . . . . . . . . . . . . .
u
.....
Manzanita Chaparral
0
20
Z
"..
Healhland, Dark Island
''-
" "b Chamise Chaparral
t
!
i
i
10
2( )
30
40
T i m e s i n c e fire, yr.
Fig. 8. Trends in plant diversity in Mediterranean-typecommunitieswith time after fire (from
Kruger, 1983, with the permission of Springer-Verlag,Berlin).
disturbances have been minimal. Where man's impact has altered the frequency of burning the range of different seral stages is reduced' (Fox, 1983 ).
The Californian chaparral is dominated by Adenostoma fasciculatum
(chamise) which regenerates after fire both by seeding and resprouting
(Rundel et al., 1987 ), it is interesting that the seed-bank of chamise increases
with time since fire, an unusual strategy for older, resprouting dominants in
Mediterranean-type ecosystems ( Z a m m i t and Zedler, 1988). Short-lived or
'temporary' cover appears to be important to ecosystem stability immediately
after fire (Horton and Kraebel, 1955 ) and species diversity is greatest in the
first few years and decreases rapidly thereafter (Parsons, 1976; and see limited data for chamise chaparral, Fig. 8 ). Growth is most rapid in the earliest
years (Hanes, 1971; Rundel and Parsons, 1979). Chaparral stands older than
60 years are 'decadent' (Hanes, 1971 ), characterized by little growth and an
accumulation of dead material, both on the shrubs and in the litter layer. There
may be an accumulation of phytotoxic materials - materials which have an
allelopathic effect and thereby inhibit regeneration (Hanes, 1971; Christensen and Muller, 1975; Keeley et al., 1985). The foliage of younger stands of
chaparral is relatively rich in nitrogen and phosphorus; Rundel and Parsons
(1980) suggest that fire is of critical importance in recycling nutrients from
nutrient-poor 'senescent' stands to nutrient-rich regenerating stands, and that
276
P.M. A ttiwill / Forest Ecology and Management 63 (1994) 247-300
this may well be an important factor in determining the quality and quantity
of foliage for vertebrate herbivores in the chaparral.
The role of allelopathy in the Californian chaparral (as in a number of other
fire-dependent communities) remains uncertain. The hypothesis is that phytotoxins are produced by plants which inhibit germination and which are
denatured by fire. However, as Keeley et al. ( 1985 ) have argued, much of the
evidence for allelopathy is circumstantial and lacks rigour; their data for the
chaparral suggest that species have a diversity of life histories and germination cues. Fire appears not to have determined the germination response of
some species, so that species diversity is enhanced by maintaining a variable
frequency of fire within a given area. (Keeley, 1987 ).
The maintenance of diversity in Mediterranean-type ecosystems presents a
problem for the land-manager. The accumulation of large amounts of slowly
decomposing dead material (e.g. in chaparral (Hanes, 1971 ); in fynbos (Van
Wilgen, 1982)) together with hot, dry summers provides the conditions for
high inflammability and for fires of high intensity. Mediterranean-type
shrublands and heathlands form a mosaic with other vegetation types which
may differ in their fire dependency (Wells, 1962 ). They are therefore at some
risk from surrounding land and Zedler et al. ( 1983 ) documented the case of
major modification of chaparral after two fires in less than one year, leading
to near elimination of some species. On the other hand, fire suppression in
national parks causes a different set of problems; that theme will be developed later in this paper with reference to heathlands in national parks in southeastern Australia.
Shifting agriculture in the tropics
The distribution of natural woody vegetation in 1980 in tropical latitudes
(rounded areas in millions of hectares, from Lanly, 1982) is given in
Table 2.
The present rate of deforestation of tropical moist forest (as defined by
Myers ( 1991 ), roughly equivalent to closed broadleaved forest in Table 2 ) is
13.86X 106 ha year -1, almost double the rate in 1979. If the same rate of
deforestation applies to open forest, the total rate is about 22 X 106 ha year(Houghton, 1990 ). The rate of deforestation in tropical America is 7.68 X 106
ha year- 1, more than half the total rate (Houghton, 1990; Myers, 1991 ). The
original extent of tropical moist forest was 1360 X 106 ha and the present area
is 778X 106 ha of which 532X 106 ha is primary forest (Myers, 1991 ). The
latter figure is 20% less than the area of 668 X 106 ha of undisturbed and productive closed broadleaved forest in 1980 (Lanly, 1982). The tropical moist
forests of Brazil represent 28% of the world total, and the annual amount
deforested in Brazil is more than one-third of the annual deforestation of the
world's tropical moist forests (Myers, 1991 ). On a global basis, it has been
P.M. AttiwiU / Forest Ecology and Management 63 (1994) 247-300
277
Table 2
Distribution of natural woody vegetation in tropical latitudes (based on data in Lanly, 1982 )
Formation
Vegetated area
Fallow area
Closed broadleaved forest
Coniferous forest
Bamboo forest
Open broadleaved forest
1160
34
6
734
228
10
1
170
Total forest
1934
409
Shrub
Total woody vegetation
624
2558
Closed broadleaved forest includes 15 X 106 ha of mangrove forest. Fallow areas include the complex
of shifting cultivation and secondary forest growth within closed broadleaved forest (almost 20% of
the area) and coniferous forest which has been degraded by fire, over-exploitation and overgrazing.
estimated that 1 X 10 6 ha of land developed after forest clearing for agriculture is abandoned each year (Salati and Vose, 1983 ).
The mean rate of destruction of the world's tropical moist forest is 1.8%
per year (Myers, 1991 ). Rates in Ecuador, Madagascar, Mexico, Philippines,
Thailand and Vietnam are in the range 4-10% per year, and on the Ivory
Coast and in Nigeria, 14-16% per year. The relatively small area of tropical
forest on Haiti was almost completely destroyed by the end of the last century
(Lewis and Coffey, 1985) and continuing rates of forest clearing on other
islands is giving cause for concern (Lugo et al., 1981 ) (3.3% per year rate of
deforestation from 1980 to 1986 in Jamaica (Eyre, 1987) ).
For Latin America, 370 × 106 ha (range 313-412 × 106 ha) were replaced
by some other ecosystem over the period 1850-1985 (Houghton et al., 1991 ).
This area is 28% of the forest area in 1850; 44% of the 370× 10 6 ha has been
converted to pastures, 25% to croplands, 20% is degraded land, and shifting
cultivation accounts for 10%. The rate of conversion of forests to managed
lands in Latin America was 5.6 × 106 ha in 1980, increasing to 6.6 × 106 ha in
1985 (Houghton et al., 1991 ).
Shifting cultivation (or swidden, bush fallowing or slash-and-burn ) is a traditional form of agriculture used by at least 240 million people in the humid
tropics (Padoch and Vayda, 1983 ). A small area of forest (0.5-2 ha) is felled
and the debris is burnt providing an ash bed. A variety of plants is then grown
until their productivity declines after 2-4 years; that area is then abandoned
to the encroaching forest, and agriculture shifts to a new forest clearing. In
the Amazon Territory of Venezuela, for example, shifting cultivation has been
practised for more than 3000 years (Jordan, 1987). The Venezuelan studies
(Jordan, 1987, 1989; Buschbacher et al., 1988 ) support the long-standing work
of Nye and Greenland (1960) that 'the soil under shifting cultivation' suffers
278
P.M. Auiwill / Forest Ecology and Management 63 (1994) 247-300
little, nutritionally or physically, under traditional practices. Jordan (1987)
concludes that 'in contrast to crop plants, native successional species invading an old conuco do not appear to suffer from low availability of nutrients.
The wild species are adapted to take up nutrients relatively unavailable to
crop species'.
Nor does the general pattern of recovery of forest structure and composition suffer under traditional shifting cultivation; Saldarriaga et al. (1988)
state that shifting agriculture has influenced the Amazon basin for millenia
and that, together with wind, lightning and fire, is part of the disturbance
regime - presumably a regime of endogenous disturbance. Similarly, there is
no clear distinction between patterns of recovery of tropical forest in Southeast Asia after disturbance by shifting cultivation or by other natural disturbances (Kartawinata et al., 1989). The studies based on Venezuelan Amazonia are outstanding (see Fig. 1; Herrera et al., 1981; Uhl et al., 1981, 1982,
1988a,b; Jordan, 1985, 1987; Saldarriaga, 1987; Uhl, 1987; Buschbacher et
al. 1987, 1988; Saldarriaga et al., 1988; Saldarriaga and Uhl, 1991 ) in that
they clearly define, in terms of biomass, productivity, nutrient cycling and
species composition, the decreasing capacity of the tropical forest to recover
with increasing intensity of disturbance. With low-level disturbance (shifting
cultivation or slash-and-burn, Fig. 1 ) return to the mature primary forest is
possible with a time-scale of about 190 years. The greater the input into maintaining the clearing in a forest-free state (bulldozing, ploughing, applying fertilizers and herbicides, sowing grasses for pasture) the less certain it becomes
that forest will re-establish (Fig. 1 ) and the time-scale for possible recovery
increases to a thousand or more years. Under traditional resource-use the
tropical forests recover from disturbance, whereas under modern, fossil fuelusing humans, the forests are destroyed (Padoch and Vayda, 1983 ). The threat
of catastrophic and exogenous disturbance of tropical forest for agriculture
has been eloquently summarized: 'The period of the great cattle ranches was
only the beginning of the exploitation of the Amazonian nature for the benefit
of foreign, non-Amazonian regions. With it started a true 'second wave of the
conquista' for the Amazonian part of South America... The second wave now
rolls unhindered over the interior of Amazonia and devours one of the very
last reserves on earth of still almost intact nature. The aim is the exportation
of all natural resources of Amazonia that can be turned to money in other
countries, other continents, i.e. to the so-called 'highly developed', industrialized countries in North America and Europe, as well as Japan, without any
benefit for the Amazonian region and the native inhabitants of its interior'
(Sioli, 1991 ).
It is not only the large-scale operations of forest clearing and pasture development which are the cause for great concern; forestry operations, the increasing intensity of shifting cultivation, and the increasing risk of fire are all
contributing to serious changes in forest structure and composition.
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
279
The rate of logging without any substantial measures for regeneration in
the tropics has been estimated as 12 X l 0 6 ha year-~ (G6mez-Pompa and
Burley, 1991 ). Bruenig ( 1985 ) makes the valid point that it is often implied
that logging invariably causes deforestation, which of course is not true. However, management is the key to successful logging and subsequent regeneration of the forest; 'the arguments for intensive management (of the tropical
forest) have never been stronger' (Wadsworth, 1983). Bruenig (1989) concluded that the obstacles to conservation and rational, sustained-use management of tropical forests are political: 'applications of knowledge, implementation of plans and improvement of forestry practice are not so much problems
of scientific knowledge and technical practicability, as of public awareness,
and political responsibility and goodwill'. Whitmore ( 1991 ) makes the same
point: 'Nor should we forget that rain forest management fails, as it usually
does, not because the silvicultural system is biologically unsound, but because
of social, economic and political forces'. The deficiencies of management are
vividly described for the Paragominas area of the Eastern Amazon: 'Taken
alone, selective forest cutting is not a severe disturbance or cause for concern,
but in the Paragominas region, several factors interact to make timber harvesting activities much more detrimental than they first appear. Since cutting
is being done by people who do not own the land, there is no interest in doing
the work carefully. Many more trees are cut than are actually harvested. During skidding operations when boles with good shape are dragged to spur roads,
many young saplings are crushed, and dead trees and branches are scattered
about. The end result is thousands of square kilometres of cut-up forest crisscrossed with soil exposed by machinery, and laden with dead slash. The extensive canopy openings and the dry slash on the forest floor turn a normally
fire-resistant ecosystem into a fire-susceptible ecosystem. Fires that are set in
nearby pastures to control weeds easily spread into the logged forests. The
logging roads provide a route for rapid spread of fire into the forest interior.
Few forests in the Paragominas area have escaped fire damage.' (Buschbacher et al., 1987).
Traditional shifting agriculture can be viewed as an endogenous disturbance, as outlined above. However, as the intensity of shifting agriculture increases, degradation of the ecosystem follows (Fig. 1; and see Herrera et al.,
1981 ). The problem is intensified with greater accessibility, and this often
follows the construction of roads for logging. Clear-felling from intensive agriculture is then associated with timber removal, and with major disturbance
by machinery as the openings become larger and site preparation becomes
more intensive. This pattern of intensified shifting agriculture with increased
accessibility has been recorded in Costa Rica (Ewel et al., 1981 ), in Brazil
(Malingreau and Tucker, 1988; Uhl et al., 1991 ), in Thailand (Kunstadter,
1990) and in South-east Asia (Kartawinata et al., 1989).
Fire in tropical rainforest has been widespread (as shown by the c o m m o n
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P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
occurrence of charcoal in Amazonian soils) but it is relatively rare (Uhl and
Kauffman, 1990). When the canopy is opened, however, there is an increase
in the amount of debris, an increase in temperature and a decrease in moisture content of potential fuel. The risk of spread of fire therefore increases
substantially, and evidence for the Amazon Basin is that fire is now occurring
at an unprecedented rate (Kauffman and Uhl, 1990).
Major disturbance by insects
Much of the ecological literature on major disturbance by insects is from
North American studies on the eastern spruce budworm (Choristoneura fumiferana) and the spruce beetle (Dendroctonus rufipennis) in north-eastern
forests of balsam fir (Abies balsamea) and spruce (Picea spp. ), on the western spruce budworm (Choritoneura occidentalis) in north-western forests of
Douglas-fir (Pseudotsuga menziesii) and other conifers, and on the spruce
beetle and the mountain pine beetle (Dendroctonous ponderosae) in northwestern forests of Engelmann spruce (Picea engelmannii), subalpine fir (Abies
lasiocarpa ) and lodgepole pine (Pinus contorta ).
The eastern spruce budworm attacks balsam fir causing high mortality. Ost a f f a n d MacLean (1989) cite up to 23 million ha of dead and dying trees in
a severe outbreak in 1983. In an unusually severe spruce budworm outbreak
on Cape Breton Island, Nova Scotia, from 1974 to 1981, attack was first most
prominent in less vigorous trees, but later was distributed equally over all size
classes (MacLean and Ostaff, 1989). Budworm caused the death of 78-89%
of the balsam fir and 27% of the spruce. A further 39% of the spruce was killed
by spruce beetle (Ostaff and MacLean, 1989). MacLean and Ostaff (1989)
found that mortality of fir due to defoliation by budworm was highly correlated with basal area. In contrast, Piene (1989a,b) found that mortality was
significantly greater in spaced stands of fir than in unspaced stands.
There has been a great deal of debate over the years on the eastern spruce
budworm (see reviews by Ghent et al., 1957; Baskerville, 1975; MacLean,
1980). Has it become worse with post-settlement logging and fire control?
Blais ( 1983 ) describes spruce budworm outbreaks as 'a natural p h e n o m e n o n
the result of the coevolution of a plant-insect system'. However, outbreaks
are becoming more frequent and larger in size, the 1970 outbreak covering
some 55 million ha. Blais (1983) links these increases with fire suppression,
clear-felling for pulpwood and the application of insecticides, all practices
which favour the growth and survival of fir, the species most preferred by the
budworm. After budworm attack, there is 'an extraordinary accumulation of
debris.., which offers no obstacle to the rooting of balsam' (de Gryse, 1944,
cited in Ghent et al., 1957). Ostaff and MacLean (1989) concluded that
'spruce budworm outbreaks are periodic natural occurrences that.., recycle
older fir stands to regenerating fir stands'.
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2 81
In the sub-alpine forest of the Rocky Mountains, succession is dominated
by disturbance (Veblen et al., 1991 ). After severe fire, Engelmann spruce and
lodgepole pine dominate the seedling regeneration. The shade-tolerant subalpine fir (Abies lasiocarpa) establishes on forest litter and slowly increases
in abundance. Severe spruce beetle attacks on pine and spruce then shift dominance to sub-alpine fir (a non-host species). As the fir matures, it is attacked
by fungi and dies, releasing sub-canopy fir and spruce; the stand then becomes
increasingly susceptible to beetle attack and blowdown in wave-like oscillations. 'Disturbance by spruce beetle outbreaks is wide-spread in the subalpine
forests of the southern Rockies, and perhaps is as significant as fire in forest
development' (Veblen et al., 1991 ).
The westem spruce budworm attacks Douglas-fir (Pseudotsuga rnenziesii)
forests of the Pacific north-western USA and British Columbia (Van Sickle
et al., 1983 ) and other conifers. McCune ( 1983 ) estimated that fire suppression in mesic coniferous forest of the Bitterroot Range of the Rocky Mountains has increased the fire return time from 60 years pre-settlement to 7000
years between 1910 and 1980. McCune (1983 ) suggests that recent increases
in intensity of budworm outbreaks are linked to fire exclusion; fire exclusion
favours the shade-tolerant and budworm-susceptible species (Abies grandis,
A. lasiocarpa and Picea engelmannii).
The mountain pine beetle kills extensive areas of lodgepole pine with a frequency of build-up of infestation within a given area of forest from 20 to 40
years ( A m m a n and Schmitz, 1988): 'Mountain pine beetle-lodgepole pine
interactions appear to favour survival of both species... The beetles' continued role in the seral stands depends upon the presence of fire... The large
buildup of fuel as a result of mountain pine beetle infestation sets the stage
for a stand replacement fire that eliminates competitive tree species such as
Douglas-fir, the true firs, and spruces. Following fire, lodgepole pine usually
seeds-in abundantly. Mountain pine beetle outbreaks occur again close to the
time when lodgepole pine matures, i.e. when adequate seed is available to
regenerate the site to lodgepole, but before excessive amounts of seed are
available that would cause an overstocked stagnated stand'.
Episodic regeneration patterns of lodgepole pine are associated with intensity and frequency of disturbances by both fire and the mountain pine beetle
(Stuart et al., 1989). The forests therefore range from even-aged (regeneration after fire) to multi-aged (regeneration following tree mortality caused by
endemic beetles and low fire intensity). It is of great interest to note that
beetle outbreaks follow a period of decrease in the rate of growth of lodgepole
pine, and that growth increases following disturbance (Stuart et al., 1989 ).
Aber and Melillo (1991) sum up these important disturbances: 'Disturbance by fire, defoliation, or other agents is an intrinsic and necessary part of
the function of most terrestrial ecosystems - a mechanism for reversing declining rates of nutrient cycling or relieving stand stagnation. The require-
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ment for fire to reverse soil organic matter accumulation and increase nutrient cycling has been known for some time... In the cases of the budworm
and the mountain pine beetle, stand break up, the reinitiation of succession,
and the reversal of stand stagnation are facilitated by herbivory rather than
fire'.
Other natural disturbances in forests
Disturbance other than that caused by windfall, catastrophic wind, fire and
insects has not been extensively reported. A brief account of some recently
reported disturbances is included here.
Ice storms
Despite a plea for 'improved and standardized glaze observations' some 30
years ago (Lemon, 1961 ), there is little work on the effects of ice. Lemon
( 1961 ) suggested that the more susceptible species belong to earlier successional stages, so that the effect of ice storms is to accelerate forest succession.
Such a result was found in a forest in south-eastern Wisconsin (De Steven et
al., 1991 ) dominated by sugar maple (Acer saccherum) with beech (Fagus
grandiflora), basswood (Tilia americana) and others as secondary dominants. Ice storms causing heavy ice accretions up to 12 cm diameter caused
some mortality of beech and other dominants and sub-dominants, promoting
the release of saplings of the shade-tolerant sugar maple. De Steven et al.
( 1991 ) concluded that disturbance by ice-storms accelerated succession toward increased cominance by sugar maple, together with the persistence of
beech and basswood through re-sprouting.
The pattern of disturbance by ice storm in the northern hardwoods appears
to be heterogeneous, and intensity varies largely with topography (Boerner et
al., 1988; De Steven et al., 1991 ). In contrast, ice storms in spruce-fir forests
in the Black Mountains, North Carolina, and in the Southern Appalachians
have a frequency of 2-4 years and severe ice storms 'may be frequent natural
disasters' (Nicholas and Zedaker, 1989). There is currently concern about
decline in red spruce (Picea rubens) forests. Nicholas and Zedaker (1989)
suggest that 'ice storms play a considerable role in southern high elevation
forest growth and dynamics and should be evaluated before anthropogenic
factors are implicated in a possible forest decline'.
Snow avalanche
There is very little work on the effects of avalanches. The probability that
trees will overtop shrubs obviously increases with increasing return time.
Johnson ( 1987 ) showed that return times were both shorter and less variable
at the top of the avalanche path, increasing as a double exponential function
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
2 83
with distance down the path. Avalanches therefore are an important local determinant of the timberline (Veblen et al., 1977).
Erosional landslides and earthflows
Landslides and earthflows have provided a range of age and size of disturbance which have been used in studies of plant succession (e.g. in the White
Mountains of New Hampshire (Flaccus, 1959); the large (1.7 × 0.4 k m )
Gothic Earthflow in Colorado (Langenheim, 1956)). In a recent study in
subtropical lower montane wet forest of Puerto Rico, Guariguata (1990)
found that species composition resembled pre-disturbance condition within
52 years, and suggested that landslides may be important in maintaining diversity of c o m m u n i t y structure.
Earthquake landslides and volcanic activity
Garwood et al. (1979) commented on the occasional, extensive disturbance of tropical forest by earthquakes; they cite denudation of 54 km 2 of
tropical forest in Panama and of 135 km 2 of tropical forest in New Guinea in
1935. Garwood et al. ( 1979 ) presented estimates of the rate of treefall and of
erosional landslides in tropical forest generally, and of earthquake landslides
in tropical forest in Panama and New Guinea. Return times are 62-125 years
for treefall, 3300 years for erosional landslide, and 620-1250 years for earthquake landslide in New Guinea and 5000 years for earthquake landslide in
Panama. If we allow 20 years for recovery after treefall, 50 years for recovery
after erosional landslide, and 100 years for recovery after earthquake (somewhat shorter times than those used by Garwood et al., 1979), then the total
disturbed area is 12-25% by treefall, 1.5% by erosional landslide, 8-16% by
earthquake landslide in New Guinea and 2% by earthquake landslide in Panama. Earthquakes produce a patchiness of landslides and are therefore an important determinant of vegetation mosaic and species diversity in tropical
forests.
Catastrophic disturbance (mass movement and volcanic eruptions) are the
major determinants of temperate rainforests dominated by Nothofagus spp.
in the Valdivian Andes of Chile, 'one of the most seismically-active parts of
the world' (Veblen and Ashton, 1978; Veblen, 1979, 1989; Veblen et al., 1981 ).
The mid-elevation forests are dominated by the evergreen and shade-intolerant N. dombeyi, together with sub-dominants of the deciduous species N. alpina and N. obliqua. There is no seedling regeneration in the old-growth stands.
Nothofagus is long-lived ( > 500 years) and the frequency of periodic catastrophic disturbance by earthquake landslides, by deposition of volcanic ash,
and by large-scale windthrows ensures that 'extensive Nothofagus-free stands
do not occur in the Valdivian Andes' (Veblen et al., 1981; Veblen, 1985a,b,
1989).
In the Antillanca area of Chile, the evergreen N. betuloides and the deci-
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duous N. pumilio grow together on the lower slopes ( 1000 m to 1100 m or
1300 m, depending on aspect ). Above this, N. pumilio forms a narrow band,
giving way to krummholz of N. antarctica and N. pumilio. The former, because of its prostrate form and its ability to produce adventitious roots from
its branches, is maintained by its ability to survive the effects of scoria deposition and avalanche damage (Veblen et al., 1977). The timberline is dynamic and is determined by catastrophic events including strong winds and
snow avalanches and their interactions with topographical features, such as
aspect, steepness and stability of slope (Veblen et al., 1977 ).
Temperate rainforest dominated by Nothofagus in New Zealand has a similar regeneration strategy to that described above for mid-altitude Nothofagus
forests in Chile, with even-aged stands regenerating after major disturbances,
such as mass earth movement, snow damage or windthrow (Wardle, 1984;
Stewart, 1986 ). Nothofagus solandri var. cliffortioides (mountain beech) grows
in essentially pure, even-aged stands on the South Island. Regeneration is absent beneath the mature forest canopy but is prolific after disturbance.
Windthrows 10-100 ha are common, however: 'Wind damage is little cause
for concern as it must be viewed in a framework of short-term forest stability.
Periodic mortality in mountain beech forests can be seen as a regeneration
strategy of a light-demanding species, since it produces ideal conditions for
forest perpetuation. Forest collapse, followed by mass regeneration is thus an
effective competitive mechanism against a more shade-tolerant canopy species' (Jane, 1986).
This paper does not cover the extensive ecological literature on colonization and succession of new areas emerging from lava flows or glaciers. However, there is an interesting study of the effects of tephra (Antos and Zobel,
1985 ). The Mount St. Helens eruption, May 1980, destroyed some 500 km 2
of forest and deposited tephra over a much larger area. Antos and Zobel
( 1985 ) studied the effects oftephra deposits from 23 to 150 cm deep in Tsuga
heterophylla and Abies amibilis forests to the north-east of Mount St. Helens,
the area of m a x i m u m tephra deposition. Shrubs and trees were not greatly
damaged, even where the depth of tephra was 150 cm. Bryophytes ad herbs
were more affected, but few species were eliminated by any depth of tephra.
Antos and Zobel ( 1985 ) place great importance on refugia such as fallen logs
from which recolonization of the devastated area can spread.
The managed utilization of forests: perspectives in conclusion
Disturbance has always been recognized by ecologists and toresters, but it
has been traditionally treated as a superfluous influence (Oliver and Larson,
1990). In the last 20 or so years, however, world-wide, explicit and instant
communications increasingly document the magnitude, frequency and universality of disturbance. Oliver and Larson (1990) make the very strong
P.M. Attiwill / Forest Ecology and Management 63 (1994) 247-300
28 5
c o m m e n t that 'natural disturbances are so c o m m o n and so devastating that
there are probably few human disturbances for which a counterpart cannot
be found in nature'.
The literature on experimentation and observation of disturbance as a natural force in forest ecology reviewed here (in part) and elsewhere (Pickett
and White, 1985a; Oliver and Larson, 1990) is enormous, and the conclusions from this literature that development of most forests of the world is
moulded by disturbance and that their sustainability depends on disturbance
seem incontrovertible.
Since diversity, structure and function of forests are moulded by natural
disturbances, it follows (as Whitmore (1982) proposed) that management
of natural forests should be based on an ecological understanding of the processes of natural disturbance. I accept S.G. Taylor's (1990) concept of a 'natural' forest as one which is recognized as supporting native vegetation by simple field observation and the application of conventional phytosociological
criteria. In dealing with the term 'natural forest' in Australia, Taylor (1990)
states that: 'the present equilibrium vegetation (in Australia) has not been
'isolated in time' from the pre-Aboriginal native vegetation of the late Pleistocene. It has descended from this late Pleistocene native vegetation through
an unbroken sequence of autogenic and allogenic successional responses to
human-generated disturbance and other natural agents of landscape change'.
We are therefore part of the natural forest; all of our options such as fire
control policies, timber utilization plans, protection of old-growth values and
conservation of biodiversity involve management decisions. But we can go
little further since nature and its conservation are controversial. As examples
of this controversy, Passmore' s (1974) humanism, on the one hand, treats
conservation as the wise husbanding of nature in the interests of social traditions and harmony; on the other, Sheldrake's (1990) animism leads to a 'deep
ecology' which is 'life-centered rather than human-centered, biocentric rather
than anthropocentric; an ecology which recognizes the interconnectedness of
all life and sees humanity as part of a larger living whole'.
This paper concludes with some comments on conservation of tropical forests; the argument for planned management of forests is nowhere more powerful than for the forests of the tropics, nor is the destruction of forest anywhere greater.
It is clear that the greater the intensity of disturbance, the less likely it is
that the forest will recover (Fig. 1 ). For example, whereas traditional shifting
agriculture is sustainable, the intensive and extensive clearing of tropical forest for non-sustainable agriculture which followed the Green Revolution leads
to degradation (e.g. G6mez-Pompa et al., 1972). Selective logging which is
not properly planned and supervised results in only the best timber being extracted, leaving behind a fragmented and altered forest; the spirit of colonization (of opening up new frontiers, of building new roads) together with the
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perception that the land-use value of forests is small, leads further to the clearing of the selectively-logged forests for farms and pastures (Uhl et al., 1988a,
1991; Dykstra and Heinrich, 1992).
If the high-value timber of the tropical forests is to be utilized in a way
which is ecologically compatible and sustainable, selective silvicultural systems must be developed based on knowledge of size and frequency of natural
gap disturbance (e.g. Kasenene and Murphy, 1991 ). The appropriate silvicultural systems must be accompanied by detailed and intensive planning,
management and supervision and by appropriate revenue structures (Repetto, 1987; Kemp, 1992; Mok, 1992; Brown and Press, 1992; and see Healey's (1992) report on a conference preceding the Earth Summit (Rio de Janeiro, June 1992) on Wise Management of Tropical Forests). Sustainable
utilization of wood must be planned in concert with sustainable utilization of
non-wood products, the rich ethnobotanical yield of the tropical forest (Boom,
1985; Myers, 1988; Bennett, 1992). Utilization must be planned within appropriate systems of reservation (e.g. Nations and Komer, 1983; Katzman
and Cale, 1990) and supplemented by appropriate areas of plantation (Gladstone and Ledig, 1990). The planned and sustainable utilization of tropical
forests for all of their benefits demands that short-term, larger but non-sustainable profits are forsaken for the protection of the tropical forests, its peoples and their cultures.
Whitmore ( 1990 ) concluded his book on tropical forests with words which
are generally applicable to the management of forested lands and which are
therefore appropriate as a summary of this paper: '...tropical forests are a renewable resource which can be utilized and still retain their diversity and
richness for mankind's continuing benefit; but only if we care to learn enough
about how they work, and also if, as has been repeatedly stated, utilization
takes place within the limits of the forest's inherent dynamic processes'.
Acknowledgements
This review was funded by the Forest Protection Society and I am most
grateful to Robyn Loydell and Dominique LaFontaine for their support and
encouragement. Dr. Pauline Grierson and Isabelle Strachan assisted with library research. My interest and concern in this topic have been sustained over
recent years by research on the recovery of mountain ash forest following the
disastrous bushfire of Ash Wednesday 1983. Australian Research Council,
Rural Credits Development Fund, Ian Potter Foundation, William Buckland
Foundation, Department of Conservation and Natural Resources, Victoria,
and Victorian Association of Forests and Forest Industries funded various
aspects of that project. My interests in tropical forests were greatly enhanced
by time spent with Drs. Ernesto Medina and Elvira Cuevas, Instituto Vene-
P.M. A ttiwill / Forest Ecology and Management 63 (1994) 247-300
287
zolano de Investigaciones Cientificas, Caracas, Venezuela, and with Dr. Ariel
Lugo, Institute of Tropical Forestry, San Juan, Puerto Rico.
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