ECOSYSTEMS OF DISTURBED GROUND - CanopyMeg - Official Website

Reprinted from:

ECOSYSTEMS OF THE WORLD 16

ECOSYSTEMS OF DISTURBED GROUND

Edited by

Lawrence R. Walker Department of Biological Sciences, University of Nevada, Las Vegas, 4505 Maryland Parkway, Box 454004, Las Vegas, NV 89154-4004, USA

1999 ELSEVIER Amsterdam - Lausanne - New York - Oxford - Shannon - Singapore - Tokyo

Chapter 9

FOREST HERBIVORY: INSECTS

T.D. SCHOWALTER and M.D. LOWMAN

INTRODUCTION

Herbivory is the process of feeding on any plant parts, lncludlng foliage, stems, roots, flowers, fruits, or seeds. It is important in that it reduces the density of plants or plant materials, opens the canopy, transfers mass and nutrients to the forest floor, stimulates plant growth under certain conditions, and affects habitat and resource conditions for other organisms. Many studies have addressed the effects of disturbance, including ecosystem-management practices, on herbivorous insect species that are econon~ically important (e.g., Schowalter et al., 1986; Mattson and Haack, 1987; Paine and Baker, 1993) and the effects of herbivore outbreaks on plant growth and mortality (reviewed by Schowalter et al., 1986). However, measurements of the magnitude of herbivory in different ecosystems and under different environmental conditions have used various non-comparable techniques, which hinders interpretation (reviewed by Lowman, 1995). Few studies have assessed the relationship between herbivory and disturbances or environmental changes, or the effects of herbivory on other ecosystem processes.

Although loss of plant material through herbivory generally is negligible, or at least inconspicuous, periodic outbreaks of herbivorous insects can denude or kill plants of selected species over many square kilometers. This capacity to alter ecosystem structure dramatically has led to the widely-held view that herbivorous insects are biotic agents of disturbance (e.g., Veblen et al., 1994; D'Antonio et al., Chapter 17, this volume). However, insect herbivores (unlike abiotic disturbances) are an integral component of the ecosystem and respond (as do other organisms) to change in environmental conditions. Schowalter (1985) and Willig and McGinley (Chapter 27, this volume) have argued that integral ecosystem effects (e.g., stimulated

growth of non-host plants) should not be considered disturbance.

We recognize that both perspectives contribute to understanding herbivory and its relationship to disturbance. Our challenge is to represent herbivores both as integral components of communities that adapt and respond to environmental change, and as agents of change that affect ecosystem structure and function in ways similar to abiotic disturbances. At issue is the threshold at which herbivory shifts from a normal trophic process to a disturbance. Normal levels of herbivory should not be considered disturbance, but, at some undefined levels of intensity. scale, and frequency, outbreaks have effects on ecosystem structure and function similar to those of fire, storm, drought, or flood. A key aspect of herbivory is its selectivity with respect to plant species affected. Such specificity ensures a diversity of indirect effects on other components of ecosystems, in the same way that various species show different tolerances to other disturbances or environmental changes. Outbreaks typically are triggered by changing environmental conditions, especially disturbances (or their suppression), that alter the physiological condition or abundance of host plants or predators. However, outbreak populations can spread to neighboring patches, altering community composition, canopy coverage, and biogeochemical cycling processes. These herbivore effects may function as a regulatory mechanism to stabilize ecosystem processes following disturbances or during environmental changes (Schowalter, 1985).

This chapter addresses relationships between disturbance and herbivory in forest ecosystems. We review measurements of herbivory, attributes of herbivore outbreaks as agents of disturbance, disturbance-related factors that trigger growth of herbivore populations, and the effects of herbivory on forest structure and

254 T.D. SCHOWALTER and M.D. LOWMAN

ecosystem processes, including promotion of future disturbances. Comparison of data from temperate and tropical forests suggests general relationships between disturbance and herbivory. We note that our discussion focuses on herbivore relationships to discrete disturbances such as fire, storm, drought, or flood, rather than to longer-term environmental changes such as atmospheric pollutants or climate change, which are beyond the scope of this volume. Nevertheless, where appropriate, we consider these long-term environmental changes as they affect herbivore outbreaks. We also do not address exotic herbivores separately. Outbreaks of native or exotic herbivores depend on suitable environmental conditions (i.e., abundant and suitable hosts, and inadequate regulation by predators) and can have similar impact on ecosystem structure and function. Of course, exotic herbivores show a greater capacity to decimate their hosts, at least until their host- specific predators are introduced.

TYPES AND PATTERNS OF HERBIVORY

Types of herbivory

Plants are complex structures. They include nutritious parts (cytoplasm and fluids) surrounded by cell walls composed of lignin and cellulose, which insects cannot digest directly. Exploitation of plant resources is inhibited further by various biochemical defenses that are toxic or interfere with digestion. Herbivorous insects have evolved morphological, physiological, behavioral, symbiotic, and other adaptations to counter these barriers. However, these adaptations typically restrict herbivorous insects to one host (monophagy) or a few species of hosts (polyphagy) with similar defensive characteristics. Herbivores affect ecosystem structure and function to varying degrees depending on the degree of specificity and the dominance status of hosts.

Insects can be classified into different feeding guilds, or functional groups, based on their mode of exploiting plants for food. Specific groups of plant- feeders include chewers (consumers of foliage, stems, flowers, pollen, seeds, and roots); miners and borers (which eat wood, bark, or one or more of the tissue layers between the intact upper and lower epidermis of leaves); gall-jormers (which reside and feed within the plant and induce the production of abnormal growth reactions by the plant tissues); sap-suckers (which have

specialized mouthparts and survive by tapping into plant fluids); and seed-eaters and jruit-eaters (which consume the reproductive parts of plants) (Romoser and Stoffolano Jr, 1994). Only seed eaters, seedling- eaters, and some tree-killing bark beetles are true plant predators; most herbivores are more correctly classified as parasites because they do not kill their hosts but feed on the living plant (Price, 1980). These different modes of consumption affect plants in different ways. For example, folivores (species that feed on foliage) directly reduce the area of photosynthetic tissue, whereas sap-sucking insects affect the flow of fluids and nutrients throughout the plant.

Folivory is the best-studied aspect of herbivory. In fact, the term herbivory often is used even when folivory alone is measured, because loss of foliage is the most obvious and most easily measured aspect of herbivory. Folivory represents the direct consumption of photosynthetically active material. Consequently, the loss of leaf area by a tree can be used as a relative term to indicate the severity of an herbivore outbreak. In contrast, other herbivores such as sap-suckers or root- borers cause less conspicuous damage to trees which is more difficult to measure, and also may have longer- term impacts (e.g., by disease transmission).

Patterns of herbivory

All plant species support characteristic associated insect herbivores, although some host a greater diversity of herbivores and suffer higher levels of herbivory than do others. Some plants can sustain high levels of herbivory and survive, whereas other species suffer mortality at significantly lower levels. The consequences of herbivory vary significantly, not just among plant species, but also in conjunction with different spatial and temporal factors (Huntly, 1991). For example, the combination of drought and herbivory or the effects of forest fragmentation on herbivore activity can significantly affect the ability of the host plant to respond. The relative timing of herbivory and the intervals between attacks also have important effects on the overall impact on the forest ecosystem.

Measurement of herbivory is usually expressed in temporal units (e.g., daily or annual rates), and ranges from negligible to several times the standing-crop biomass of foliage (Table 9.1), depending on forest type, environmental condition, and regrowth capabil- ity of the plant (Lowman, 1995). Lowman (1992)

FOREST HERBIVORY INSECTS

Table 9.1 Herbivory measured in temperate and troplcal forests (including understory) ' Location Forest type Level of grazing ~ e c h n i ~ u e ~ Source

understory only

tropical rainforest

Eopic.ul

Costa Rica tropical rainforest 7.5? b (new leaves)

tropical rainforest 309 b (old leaves)

Pananla tropical rainforest 13%

Barro Colorado Island tropical rainforest 8% (6% insect; 1-2% (Panama) vertebrates)

15%

21% (but up to 190%)

Puerto Rico 7.8%

5.5-16.1%

2-6%

2-13%

0. 1-2.2n/o

9-12?b

1 5-300? b

14.69 h

8-1 29 h

Vene~uela understory only

New Guinea tropical rainforest

Australia sclerophyll

subtropical rainforest

Cameroon (Afrlka) tropical rainforest

Tempel-utc,

North America temperate deciduous

coniferous

Australia

Europe

sclerophyll

dry sclerophyll

montane or cloud rainforest

warm temperate rainforest

deciduous

Stanton ( 1975)

Stanton 11 975)

Wint (1983)

Leigh and Smythe (1978)

Lelgh and Windsor (1982)

Coley (1983)

Odum and Ruiz Reyes (1970)

Benedict (1976)

Schowalter (1994)

Schowalter and G a n ~ o (1999)

Golley (1977)

Wint (1983)

Lowman and Heatwole ( 1992)

Lowman ( 1 984)

Lowman et al. (1993)

Bray (1963)

Relchle et al. (1973)

Schowalter et al. (198 1)

Schowalter (1989)

Schowalter (1995)

Lowman and Heatwole (1992)

Fox and Morrow (1986)

Ohmart et al. (1983)

Lowman ( 1984)

Lowman (1 984)

Nielsen 11 978)

' Updated from Lowman (1995) by permission from Academic Press. Techniques: I, visual ranking; 2, litter trap; 3, graph paper or template squares; 4, leaf area meter: 5, other calculations; 6, insect frass;

7, estimation of missing or truncated needles.

compared herbivory on five tree species in Australian rain forests, and found highly variable spatial and temporal patterns. For example, Dendrocnide excelsu, typical of early-successional species, had relatively soft, short-lived leaves devoid of tannins or phenolics (Lowman and Box, 1983) and averaged 42% loss of leaf area, due primarily to a host-specific chrysomelid beetle (Hoplostines uiridipennis; Fig. 9.1 ). In con-

trast, Doryphora sassufius, typical of late-successional species, had relatively tough, long-lived leaves with high concentrations of tannins and phenolics, and averaged only 15% loss of leaf area. Dor?phora sussafras had significantly less insect damage to its sun leaves in the upper crown (13%) than to its shade leaves in the lower crown (16%; Fig. 9.1), and suffered significantly more herbivory in warm-

T.D. SCHOWALTER and M.D. LOWMAN

Doryphora sassafras

KJ \ SHADE

2-

SEPT. I OCT. I NOV. I DEC. I JAN. I FEB. I MAR. I APR. I MAY. I JUN. I JUL. I AUG. 1,

Dendrocnide excelsa

Fig. 9.1. Herhivory averaged over a period of 3 years for neighboring trees of Dendrocnide excelsa and Doryphora sassufi-us (open circles, sun; solid circles, shade) in Australian rain forest. Dendrocnide exce1.s~ 1s a colonizer that grew in a disturbed patch, and Doryphor-a sussufi-us is a slow-growing, shade-tolerant species in undisturbed conditions. The standard error common to all the observations in the graph is indicated in the lower right corner (site: complex notophyll vine forest in Dorrigo National Park, New South Wales).

temperate rainforest stands (23%) than subtropical cult [reviewed by Lowman (1995) and Lowman and rainforest stands (1 5%; Lowman, 1992). Wittman (1996)l. Eucalypt forests have shown annual

Levels of herbivory are variable between forest foliage losses of more than 300% of the foliage stands and among ecosystem types (Table 9.1), al- standing crop (Lowman and Heatwole, 1992). Although though researchers have used a variety of methods, some studies suggest that tropical forests sustain higher which makes interpretation and comparisons dim- levels of herbivory than do temperate forests (Coley


and Aide, 1991), the data in Table 9.1 indicate considerable variability in herbivory among forest types, even when comparable methods are used. Most methods are short-term snapshots of herbivory and do not provide information on deviations in environmental conditions, plant chemistry, or herbivore densities from long-term means. Long-term studies using standardized techniques are necessary to compare rates of herbivory among forest types.

Temporal and spatial variability in herbivory has recently become a subject of debate in terms of how best to quantify this process, as biologists are becoming more capable of sampling leaves in the canopy. The methodological problems in sampling herbivory are twofold: 1) logistic problems of access to leaves in tall trees; and 2) selection of a method of measurement that is both accurate and efficient in use of time. The levels of accuracy among methods of measurement must take into account the specific hypotheses tested. For example, measuring the percentage of leaf area missing at a point in time may be an appropriate measure of the effect of herbivory on photosynthetic capacity or canopy-atmosphere interactions, but is an inaccurate representation of consumption by herbivores because holes expand as leaves expand, and completely consumed or prematurely abscissed foliage cannot be assessed (Lowman, 1984; Risley and Crossley, 1993).

Changes in herbivore abundance affect patterns and rates of herbivory. However, few studies have addressed changes in herbivore abundance or herbivory as a result of disturbance or succession. Schowalter et al. (1981) (see also Schowalter, 1994, 1995) used the same methodologies to compare canopy herbivore abundances and folivory in replicate disturbed (by harvest or hurricane) and undisturbed patches of temperate deciduous, temperate coniferous, and tropical evergreen forests. In all three forest types, disturbance resulted in a shift in dominance from folivores to sap-suckers. However, individual species within each of these functional groups showed different responses to disturbance; some folivores became more abundant and some sap-suckers became less abundant after disturbance, even on the same plant species. These different responses indicate that species belonging to a particular functional group are not redundant but maintain distinct functional roles under different environmental conditions. Continued measurements in these forests (Schowalter and Ganio, 1999) indicate that dominance shifts back to folivores within five years in the tropical forest.

INSECT HERBIVORE RESPONSES TO ENVIRONMENTAL CHANGE

Although outbreaks of herbivorous insects may re- semble other disturbances, in terms of their effect on ecosystem structure and function, they differ from other disturbances in being a biotic response to changing environmental conditions. Herbivore populat~ons are sensitive to changes in environmental factors, such as abiotic conditions, host plant abundance and condition, and predation, as these influence herblvore growth, survival, and reproduction. Effects of these abiotic and biotic factors on herbivorous insects are summarized below.

Abiotic effects

Disturbances affect many herbivore populations directly, by creating lethal conditions. Populations of many species can be greatly reduced by severe disturbances, and rare species may be eliminated (Torres, 1992; Schowalter, 1994, 1995). For example, Willig and Camilo (1991) reported the virtual disappear- ance of two species of walkingsticks (Agamemnon iphimedeia and Lamponius portoricensis) from forests of tabonuco (Dacryodes excelsa) in Puerto Rico following Hurricane Hugo in 1989. Miller and Wagner (1984) reported that pandora moth (Coloradia pandora) in forests of ponderosa pine (Pinus ponderosa) in western North America pupates preferentially in bare soil where it is more likely to survive fire, a frequent disturbance in these forests. Similarly, sensitive species may be decimated by toxic effects of atmospheric pollutants, such as ozone or acid precipitation (Alstad et al., 1982). In contrast, Torres (1988) reviewed cases of insect herbivores being transported into new areas by hurricane winds, such as African desert locusts (Schistocerca gregaria) deposited on Caribbean islands.

Disturbances also have indirect effects on forest herbivores through alteration of light, temperature, and moisture. Altered light penetration through the canopy can affect visual orientation of some insect species toward necessary resources. For example, aphids orient toward green and yellow wavelengths, which are characteristic of young succulent leaves (Matthews and Matthews, 1978), young leaves are more visible and available in canopy zones exposed to light. Light intensity may also affect abundances of predators that regulate herbivore abundances. For example, Oboyski


(1995) found that the guild of webspinning spiders (trapping by non-visual means) was significantly more abundant in darker canopy zones whereas the guild of jumping spiders (hunting visually) was significantly more abundant in sunlit canopy zones. Vertical profiles of temperature and moisture typically mirror each other in closed-canopy forests; highest temperatures and lowest relative humidities generally occur in the upper canopy, whereas lower temperatures and higher relative humidities occur in lower canopy layers (Parker, 1995). In contrast, open canopies show more even profiles of temperature and relative humidity. lnsect respiration, feeding, and developmental rates generally increase with temperature, within tolerance ranges that are species-specific. Insect species also vary in their tolerances to desiccation. Hence, the distribution of various species within the forest re- flects gradients in temperature and relative humidity. Temperature and moisture also affect insect herbivores through the increased susceptibility of stressed plants to herbivores, disruption of mate attraction as pheromones are convected out of warm or open-canopied forests, and increased abundance and virulence of pathogenic viruses, bacteria, fungi, and nematodes in cooler, moister forests (Fares et al., 1980; Mattson and Haack, 1987).

Biological effects

Host condition Healthy plants have a variety of chemical, physical

and phenological defenses against herbivorous insects (see Harborne, 1982; Coley et al., 1985; Aide, 1993). Feeny (1976) and Rhoades and Cates (1 976) suggested that plant defenses should differ by successional stage: short-lived, rapidly growing early-successional species may be expected to produce highly toxic antiherbivore defenses to prevent damage, whereas long- lived later-successional plants should primarily produce feeding deterrents to reduce herbivory. While this hypothesis may explain general patterns in temperate forests, Coley and Aide (1991) compiled data from a number of studies indicating that later-successional tropical species with longer foliar life-spans may have higher concentrations and toxicity of defensive compounds than do species of temperate forests or early-successional species in tropical forests These differences correspond to the higher rates of herbivory generally reported in tropical forests (although they acknowledged a number of biases in the source data).

Disturbances, or other environmental changes, promote herbivore population growth on stressed plants or poorly defended plant species which replace better- defended species following disturbance. Disturbances injure surviving plants and alter environmental conditions, thus affecting uptake and allocation of water and nutrients by these and early-successional plants (Schowalter, 1985). Wound repair and replacement of lost foliage or root tissues to meet metabolic demands require redirection of carbohydrates and nitrogen from energetically expensive defensive compounds, such as phenols, terpenes, and alkaloids (e.g., Coley et a]., 1985; Schowalter, 1985).

If herbivorous insects have adapted to particular plants by developing mechanisms to detoxify or avoid their defenses, they feed selectively on these plants. Be- cause these adaptations also place demands on energy resources, herbivorous insects face an evolutionary trade-off between specific adaptations that tailor (and restrict) them to develop and survive on a specific host plant and more general adaptations that allow a wider host range but sacrifice survival under conditions favorable to some host plants. As stressed plants reduce production of defensive compounds, they become more general biochemical resources suitable for generalist, as well as specialist, herbivores. Consequently, the likelihood of herbivore outbreaks increases following widespread disturbances, or chronic environmental change such as atmospheric pollution (Alstad et al., 1982). For example, outbreaks of bark beetles and root- feeding beetles typically follow tree-injuring disturbances (Witcosky et al., 1986; Paine and Baker, 1993).

Host abundance Disturbances change the abundance of plant species

through selective mortality. Severe windthrow may destroy much of the overstory, leaving mainly understory vegetation. Fire, depending on its severity, may eliminate understory vegetation, may destroy both understory and overstory, or may target particularly intolerant species. Heavy rains can saturate certain types of tree trunks, leading to selective tree-fall of specific species in a forest stand (Lowman, unpubl. data). Plant species abundances subsequently change during recovery (succession), as various species respond to modified abiotic conditions and increasing competition. These changes affect vegetation diversity, as well as the proximity and apparency (likelihood of being perceived) of particular plant species to host- seeking herbivores (Schowalter et al., 1986).

Herbivorous insects are highly sensitive to host

FOREST HERBIVORY: INSECTS 259

spacing and to confusion by the presence of non- host plant species (Risch, 1980; Kareiva, 1983; Visser, 1986; Schowalter and Turchin, 1993). Different plant species produce different blends of volatile chemicals, which insect herbivores use as host cues (Visser, 1986). Volatiles from non-host plant species are non-attractive, or even repellent, to a given insect. Spacing affects the ability of insects to perceive host cues, to reach a potential host with limited time and energy reserves, and to avoid the attention of predators while searching for hosts (Kareiva, 1983; Schowalter, 1985). Hence, little search effort is needed to colonize closely-spaced hosts in monocultures, compared to sparse hosts or hosts mixed with non-hosts. Dense hosts also are more likely to be stressed (and susceptible to insects) as a result of competition for limited resources or of injury by falling neighbors. A particular plant species may become more apparent to its associated herbivores if disturbance selectively reduces the abundance of non- hosts (Visser, 1986). Succession following disturbance often produces temporarily dense monocultures of rapidly reproducing ruderal plant species, which gen- erate herbivore outbreaks. For example, Torres (1992) reported outbreaks of several species of Lepidoptera on early-successional grasses, herbs, and vines that became abundant in the Luquillo Mountains of Puerto Rico following Hurricane Hugo.

Landscape effects The growth of herbivore populations triggered by

changes in host-plant condition or abundance may be restricted to disturbed patches in otherwise unsuitable landscapes, or may propagate into adjacent areas where conditions permit population spread (Schowalter and Turchin, 1993; Schowalter, 1995). Widespread disturbances may create even-aged, low-diversity forests, which promote the growth of herbivore populations over large areas of susceptible hosts. For example, the large-scale fires characteristic of the southern United States and arid western North America historically resulted in the development of low-density pine forests over millions of square kilometers. Populations of bark beetles (Dendroctonus spp., Ips spp., Scolytus spp.) in these forests were restricted to scattered injured or diseased trees, but likely reached levels capable of killing some trees in isolated thickets (Goheen and Hansen, 1993; Schowalter and Turchin, 1993). Anthropogenic suppression of fire and conversion of forests to more rapidly growing, commercially valuable species on a regional scale resulted in dense

monocultures of susceptible trees, which now support widespread outbreaks of bark beetles and associated pathogens (Goheen and Hansen, 1993; Hadley and Veblen, 1993). Abundant early-successional species in tropical forests may also increase herbivory, as found in the Cecropia schreberiana thickets that developed in the forests of Puerto Rico after Hurricane Hugo (Schowalter, 1994; and unpubl. data).

Forest fragmentation may also affect herbivory in forests. Kruess and Tscharntke (1994) and Schowalter (1 994, 1995) reported that diversity and abundance of predator species were greatly reduced in disturbed or fragmented habitats, releasing herbivore populations from control by predation. Roland (1993) reported increased duration of a defoliator in fragmented forests. Herbivore outbreaks across landscapes dominated by susceptible patches could threaten areas within these landscapes that are designated for preservation. Other- wise resistant patches of forest could be overwhelmed by constant inundation by herbivore populations from adjoining patches. For example, Futuyma and Wasser- man (1980) reported severe defoliation of scattered white oaks (Qtlerctu alba; an ordinarily resistant host) by fall cankerworm (Alsophila pornetaria) in stands dominated by this insect's primary host, scarlet oak (Qtlercus coccinea). In contrast, forest fragmentation can interrupt spread of some bark beetles, such as Dendroctonus frontalis (Schowalter, 1985).

An obvious issue regarding preservation of particular patches within landscapes is the minimum critical size needed to protect ecosystem processes and prevent disruption by herbivores from neighboring patches (Lovejoy and Bierregaard, 1990). Few natural exper- iments have addressed this issue. Shure and Phillips (1991) suggested that the significant effects of the size of a clear-cut area on arthropod diversity, abundance of functional groups, and biomass might reflect the extent of environmental difference between the cleared patch and surrounding forest. In no situation have insect pests been manipulated experimentally between fragments of different size. Such an ambitious project would require overwhelming logistic and financial support. Nevertheless, controlled experimental studies have indicated that herbivore damage is more extensive in low-diversity stands than in stands with greater distance between conspecifics (e.g., Schowalter and Turchin, 1993), just as in agricultural monocultures versus polycultures (Risch, 1980; Kareiva, 1983). In natural forests, herbivory of Doryphora sassafras was


greater in stands where the species was more common than where it was relatively rare (Lowman, 1992).

In reviews of forest-dieback situations (Old et al., 1981; Huettl and Mueller-Dombois, 1993), insect epidemics commonly were implicated as major causes of mortality, and were usually most severe, in situations where human disturbance was part of land-use history. In contrast, outbreaks are considered less frequent in natural forest ecosystems where human disturbance has not been a major factor (but see Wolda and Foster, 1978; Wong et al., 1990). The relationships between insect populations, forests, and human impacts are the subject of extensive global speculation, as awareness of the importance of forests in regulation of global climate increases (e.g., Salati, 1987).

INSECT HERBIVORES AS AGENTS OF DISTURBANCE

Although debate may continue over whether or not insect herbivores should be regarded as agents of disturbance (Veblen et al., 1994; see also D'Antonio et al., Chapter 17, this volume), rather than simply as ecosystem components that respond to environmental change (Schowalter, 1985; see also Willig and McGin- ley, Chapter 27, this volume), herbivore outbreaks can dramatically alter ecosystem structure and function in much the same way as other disturbances. Effects of herbivore outbreaks, as for other disturbances, depend on the type, intensity, severity, scale, frequency, and regularity of the outbreak, as described below.

Different herbivore species and functional groups (e.g., foliage chewers versus sap-suckers, described above) determine the type of disturbance - that is, the plant species and part(s) affected. Outbreaks of any insect species affect some plant species more than others, depending on host preferences. Even the gypsy moth (Lymantria dispar), an exemplar of polyphagy, shows a distinct preference for tree species producing mainly phenolic defenses, and avoids species with non-phenolic defenses (Miller and Hansen, 1989). Monophagous insects can virtually eliminate a particular plant species from the outbreak area. For example, southern pine beetle (Dendroctonus jiontalis) can kill all pine trees (Pinus spp.) over relatively large areas, resulting in replacement by hardwood species (e.g.. Schowalter, 1985); hemlock woolly adelgid (Adelges tsugae) has eliminated eastern hemlock (Tsuga canadensis) from large areas of the

northeastern United States (McClure, 199 1). Foliage chewers open the canopy and transfer solid canopy material to the forest floor, in much the same manner as a storm, whereas sap-suckers siphon and transfer liquid materials from the plant.

lntensity and severity of herbivory depend on the population level of the herbivore and the susceptibility of hosts. Most herbivore species never reach densities capable of causing serious injury to hosts. However, as noted above, different plants and plant species respond differently to the same intensity of herbivory, resulting in differing severity of injury. Deciduous trees often are capable of repeated refoliation following chronic complete defoliation, as a result in part of adaptations to seasonal senescence and refoliation, whereas many evergreen species are killed by a single complete defoliation (Schowalter et al., 1986).

Herbivory can affect forests over a wide range of spatial scales. In some cases, a single tree may be targeted, as in the case of bark-beetle attraction to individual lightning-struck trees (Paine and Baker, 1993). In other cases, dense patches of hosts within an otherwise resistant forest may be affected. In the most extreme cases, herbivory can affect thousands of square kilometers. Furniss and Carolin (1977) have given examples of defoliator outbreaks in western North America that have covered 3000-200 000 km2. Veblen et al. (1994) reported that, since -1633, 39% of a montane forest landscape in Colorado (U.S.A.) has been affected by outbreaks of spruce beetle (Dendroctonus rufipennis), compared to 9% by snow avalanches and 59% by fire.

Outbreaks of many herbivores occur infrequently and irregularly. Other herbivores show regular cycles of outbreak and decline. For example, outbreaks of the Douglas-fir tussock moth (Orgyia pseudotsugata) in western North America occur at intervals of approximately 9-10 years and generally last 2-3 years (Mason and Luck, 1978). Veblen et al. (1994), in a montane forest landscape in Colorado, reported an average return interval for spruce beetle outbreaks of 1 17 years, compared to 202 years for fire. Such patterns may correspond to climate cycles or to forest recovery time (re-establishment of dense hosts or host resources) following disturbance. Fuel accumulation resulting from outbreaks of some bark beetles and defoliators at particular successional stages ensures regular fire return intervals, thereby maintaining relatively even- aged forests dominated by host species (Schowalter, 1985). lncreasing the reliability of return interval for

FOREST HERBIVORY INSECTS 26 1

herbivory or other disturbance should increase the rate of directional selection for tolerance.

EFFECTS OF INSECT HERBIVORES ON ECOSYSTEM PROCESSES

Herbivory affects ecosystem processes through its influence on plant reproduction, growth, survival, and turnover. Processes affected include primary productivity, community dynamics, biogeochemical cycling, and canopy-atmosphere-soil interactions, as discussed below.

Primary productivity

Traditionally, herbivory has been viewed solely as a process that reduces primary productivity. As described above. herbivory can remove several times the standing crop of foliage, alter the growth form of trees, or kill all trees of selected species over large areas during severe outbreaks. However, several recent studies indicate more complex effects of herbivory.

Low or moderate levels of herbivory can stimu- late plant growth (e.g., Carroll and Hoffman, 1980; Lowman, 1982; Trumble et al., 1993), whereas severe defoliation usually results in mortality or decreased fitness (Marquis, 1984). Trumble et al. (1993) reviewed literature demonstrating that compensatory growth (replacement of consumed tissues) following low to moderate levels of herbivory is a widespread type of response. Wickman (1980) and Alfaro and Shepard (1991) reported that short-term growth losses by defoliated conifers were followed by several years, or even decades. of growth rates that exceeded pre- defoliation rates. Such compensatory growth likely depends on suitable growing conditions. Lovett and Tobiessen (1993) reported that defoliation resulted in elevated photosynthetic rates of seedlings of red oak (Quercus rubra) grown under conditions both of low and of high nitrogen availability, but that high- nitrogen seedlings were able to maintain the high photosynthetic rates for a longer period of time. Hence, resource-limited trees are more likely to succumb to herbivores. Trees stressed by multiple factors, including herbivory, may lack resources or ability to compensate, and often succumb to mortality agents such as bark beetles (Schowalter, 1985). High levels of herbivory can exceed a plant's ability to compensate and lead to growth reduction and mortality.

Although herbivory may reduce productivity of targeted plants, other plant species may compensate as a result of reduced competition for resources. For example, demise of pines during outbreaks of Dendroctonus frontalis in the southern United States typically stimulates growth of understory hardwoods (Schowalter, 1985); Wickman ( 1 980) reported increased growth of ponderosa pine (Pinus ponderosa) following defoliation of white fir (Abies concolor) by the Douglas-fir tussock moth (Orgyia pseudotsugata) in the northwestern United States.

Community dynamics

Selective herbivory among plant species provides space and other resources to non-targeted plant species, resulting in altered plant community composition (e.g., Schowalter et al., 1986; Davidson, 1993). Herbivores also affect other animals and micro-organisms through changes in food resources and in plant and soil conditions.

Depending on various conditions, including the successional stage of the vegetation and the particular herbivores involved, succession may be advanced or reversed (Davidson, 1993). For example, in coniferous forests of interior western North America, pine forest represents an earlier successional stage and fir forest a later successional stage. Where moisture is adequate and fire return intervals are long (riparian corridors and high elevations), bark beetles (Dendroctonus ponderosae) advance succession by accelerating the gradual replacement of host pines by the more shade- tolerant, fire-intolerant firs. However, where moisture is inadequate and fire return intervals are short (lower elevations), defoliators and bark beetles associated with the water-stressed firs reverse succession, and help to maintain the earlier-successional pine forest.

Changes in plant community composition and structure affect habitat and food for other animals and micro-organisms. Changes in abundance of particular foliage, fruits, or seeds affect abundances of animals that use those resources. Populations of animals that require or prefer nesting cavities in dead trees may be increased by tree mortality resulting from herbivore outbreaks. Changes in the rate or quality of litter fall and in the litter microclimate resulting from folivory can affect litter communities. For example, Schowalter and Sabin (1991) reported that three taxa of litter arthropods were significantly more abundant under Douglas-fir (Pseudotsuga menziesii) saplings


after defoliation. Insect herbivores themselves consti- tute highly nutritious resources for insectivores. The concentrations of essential nutrients in caterpillars, especially, are several orders of magnitude greater than those in foliage tissues (e.g., Schowalter and Crossley, 1983). Abundance of insectivorous birds and mammals may increase in patches experiencing outbreaks of insect herbivores (Barbosa and Wagner, 1989). The accidental invasion of insect herbivores such as the gypsy moth in the northeastern region of the United States has had widespread biological and economic impacts on forest ecosystems. A native of Europe, the gypsy moth preferentially consumes beech (Fagus spp.), but large populations will consume foliage of other trees. In the absence of predators, gypsy moth populations have had repeated outbreaks, resulting in the absence of oaks (Quercus spp.) from urban regions, forest fragments, and even large hillsides of deciduous woodlands. Other examples of introduced insect pests that have significantly altered the forest community include the elm bark beetles Hylurgopinus rujipes and Scolytus rnultistriatus, carriers of Dutch elm disease caused by the fungus Ceratocystis ulrni (U.S. Department of Agriculture, 1979) and the hemlock woolly adelgid (see above, p. 260).

Biogeochemical cycling

Insect herbivores affect biogeochemical cycling in a number of ways. Altered vegetation composition, discussed above, changes patterns of acquisition and turnover of various nutrients by the vegetation. For example, insects (such as bark beetles) that affect the relative proportions of pines and hardwoods in the southern United States or of Pseudotsuga rnenziesii and Thuja plicata in the northwestern United States, indirectly affect calcium dynamics and soil pH, since soil under hardwoods or Thuja has more calcium and a higher pH than under pines or Pseudotsuga (Kiilsgaard et al., 1987).

Herbivory can affect biogeochemical cycling directly by changing the seasonal timing and amount of nutrient flows from plants to litter or soil. Kimmins (1972), Seastedt et al. (1983) and Schowalter et al. (1991) found that the leaching of nutrients was greatly increased following chewing by herbivores. Risley and Crossley (1993) reported that herbivory also causes premature abscission of damaged foliage. Whereas litter-fall in the absence of herbivory may be highly seasonal - for instance, concentrated at the onset of

cold or dry conditions - herbivory increases litter-fall and nutrient turnover during more productive seasons when nutrient capture, by soil micro-organisms as well as the trees, may be more efficient. However, Swank et al. (1981), working in the southern Appalachians of the United States, reported that defoliation of hardwood forests resulted in increased nitrate export in streamflow.

Herbivory also affects the form of nutrients in litter and the rate of litter mineralization (Lovett and Ruesink, 1995). Senescent foliage, foliage fragments lost via herbivory, and foliage passed through herbivore digestive systems differ in the amount and form of nitrogen and carbon compounds, and in the degree of microbial pre-conditioning of this organic material. Where defoliators are dominant among herbivores, as is normally the case in undisturbed forests, material flow is dominated by solid fragments with long turnover times. Disturbance in both temperate and tropical forests, however, causes dominance among herbivores to shift to sap-suckers (Schowalter et al., 198 1 ; Schowalter, 1994, 1995), with the consequence that the material flow is dominated by a rain of soluble materials with short turnover times.

Reduced metabolic demands by pruned or defoliated plants can reduce water and nutrient uptake (Webb, 1978) and contribute to plant survival during drought periods (Parks, 1993). Since herbivore outbreaks often are associated with drought (Schowalter et al., 1986), future studies should address the extent to which herbivores contribute to ecosystem stability under these conditions.

Canopy-atmosphere-soil interactions

Herbivory affects abiotic conditions and the likelihood of future disturbance, especially through canopy opening and fuel accumulation. Canopy opening as a result of herbivore activity has immediate effects on ecosystem conditions similar to those resulting from fire or storm disturbances. Opened canopies are subject to increased penetration of light, precipitation, and wind. These factors affect erosion, soil moisture, and soil fertility. Evapotranspiration and interception of precipitation are reduced (Parker, 1983) and oro- graphic precipitation may be affected in montane or tropical regions (Salati, 1987). Increased leaf fall and accumulation of woody litter increase the likelihood and severity of fire, especially in forests where these materials decompose slowly or where lightning strikes

Fig. 9.2. Dicback of El~crrlyj~ttrs noun-unglicu in New South Wales, Australia, illustrating the intensity of insect outbreaks that occur as a result of ecosystem disturbance and result in widespread tree mortality throughout the rural landscape.

are frequent (Schowalter, 1985). In high-diversity forests, intermittent herbivory may result in patches of open and closed canopy. The patchiness created may affect the understory by altering light levels and also variable canopy heights will affect the flow-through patterns of rainfall (see Herwitz et al., 1994). Similarly, Herwitz et al. (1994) found that variation in canopy heights and aspects led to significant differences in moisture at a small scale.

CASE STUDY - EUCALYPT DIEBACK IN AUSTRALIA

The defoliation of Australian eucalypts (Fig. 9.2) is considered one of the most severe tree declines in the world (Heatwole and Lowman, 1986; Mueller- Dombois, 1990191) and has been linked directly to ecosystem disturbance. This syndrome has been recorded for over a century (Norton, 1886), but only during the past three decades have the episodes of dieback become more frequent and severe. Millions of eucalyptus are estimated to have suffered mortality

from the dieback, although its causes are not well understood. The loss of trees on the Australian landscape is of great significance to agriculture, to the water table, to tourism, and to the sustainability of Australian ecosystems.

In Australia, dieback has been attributed to different factors according to geographic region (Old et al., 198 1; Heatwole and Lowman, 1986). In Western Aus- tralia, millions of jarrah trees (Eucalyptus mat-,oirzata) suffered primarily from a fungal pathogen that invaded the root systems of the trees. Phvtophthora cinnamomi was introduced to Australia inadvertently in the 1950s, when its spores were transported on tractor tires from soils in Indonesia. It quickly jnvaded the soils of Western Australia, and specifically targeted jarrah trees. Because jarrah was an important economic species, identifying the cause of its dieback was a particular concern. Extensive rehabilitation is under way.

In South Australia, the dieback syndrome does not have one major cause as in the case for Western Australia (Boardman, 1981). The increasing salinity of soils, as agricultural practices intensify, has been blamed for dieback in several regions; additions of non-


native grasses together with exotic grazers (e.g., sheep, cattle) have altered the natural balance of the Australian grassland and dry sclerophyll ecosystems. Fluctuating rainfall and temperature on these agricultural areas serve to exacerbate the conditions that lead to tree decline.

In New South Wales and southern Queensland, the dieback is even more complex and controversial (Carter et al., 1981; Lowman and Heatwole, 1993). Diebacks in these states are reputedly the most severe in Australia, and the causes are the most difficult to determine (D. Mueller-Dombois, pers. commun.). All of the factors (salt, weather, agriculture) in the South Australian dieback are implicated in the New South Wales diebacks, and have led to more extreme fluctuations in the populations of organisms, both native and non-native. Distributions of many native birds have been reduced with the decline of suitable trees for nesting (Ford, 1989). Many of these birds were important predators on insects. For some herbivorous insects existing in the soil as larvae conditions for survival have improved as a result of compaction of roots and soil by sheep and cattle; this has led to increased numbers of scarabs (Anoplognathes sp.) and other herbivorous beetles (Roberts et al., 1982). For many decades, farmers have reported intermittent epidemics of scarab beetles, sawfly larvae (Perga afinis), and other defoliators attacking eucalypts. The enhanced survival of herbivorous beetles, as well as other herbivores, is a consequence of human alteration or disturbance of the natural landscape.

Lowman and Heatwole (1992) evaluated the impact of the herbivore component in the New South Wales dieback by tagging leaves of eucalypts and the related genus Angophora, and measuring herbivory and mortality monthly over five years (1982-1986). The study involved sampling different heights on trees representing the three different types of stands in rural Australia: healthy trees in woodland stands; trees in pastures; and dying trees in pastures. Cumulative annual levels of herbivory ranged from 8% (Angophora floribunda in woodlands) to 88% (Eucalyptus blakelyi in isolated stands surrounded by disturbed pastures). Occasional outbreak situations resulted in some species being entirely defoliated repeatedly, and cumulative foliage losses of up to 300% (Eucalyptus nova- anglica, also in isolated stands surrounded by disturbed pastures). The highest levels of defoliation occurred in isolated trees or stands of trees that were surrounded by pastures, whereas the canopies within healthy

woodlands suffered only moderate herbivory (Fig. 9.3). Average defoliation for dieback stands in pastures was 60%, as compared to 32% and 14% for healthy trees in pastures and in woodlands, respectively.

The differences between healthy and dieback trees in pastures were linked to the history of disturbance of the surrounding pastureland. Changes in the pasture that accompanied intensive stocking of cattle and sheep included trampling of soil, consumption of seedlings by stock, clearing of trees, girdling of trees by cattle, aerial spraying of fertilizers (especially superphosphate), alterations of the water table, planting of non-native grasses for winter feed supplements, and plowing of pastures for crops. The changes in soils following these practices created conditions conducive to epidemics of various folivores. As a result of clearing, fewer trees remained standing as food sources. Subsequently, the beetles, which feed gregariously, defoliated the remaining trees. This scenario was repeated over several years, and in many cases has resulted in complete mortality of forest fragments.

As in many complex land-use situations, it is difficult to implicate insects as the major cause of Australian dieback. Which came first'? - the insect defoliation leading to tree decline, or the environmental stresses on trees leading to increased defoliation? In other regions, the depletion of soils has been a primary progenitor of tree diebacks (e.g., in Hawaii, Mueller-Dombois, 199019 1). The dietary quality of foliage also has been linked to the decline of tree health (Landsberg, 1990).

In addition to the stress of defoliation, insects may have other deleterious effects on native vegetation. Scarab beetle larvae (and other soil organisms) feed on tree roots, and create below-ground stresses on eucalypt trees (Lowman et al., 1987). This aspect of insect damage to trees is difficult to quantify because measurements require destructive sampling, and biomass estimates usually include only the plant parts present and cannot account for the portions consumed by insects or other herbivores.

Lowman et al. (1 987) examined insect damage, both above- and below-ground, to a healthy tree and a dieback tree (New England peppermint, Eucalyptus nova-anglica), growing in close proximity and similar in stature. Borers and termites occupied 19% of the branches of the dieback tree, but only 5% of the healthy tree. These are conservative estimates, because branches and roots entirely consumed by insects could not be detected. The healthy tree had 990 kg of woody biomass, versus 640 kg for the dieback tree.



The numbers of beetle larvae in disturbed soils were higher than in undisturbed soils (Roberts et al., 1982), exacerbating stress to the remaining trees. Stem borers, fungal pathogens, and sap-sucking insects have not been measured in terms of their impact on the Australian dry forest, but could emerge as additional stresses on the trees.

Forest diebacks are worldwide and have various origins and etiologies (Mueller-Dombois, 1986, 1987). The Australian dieback is particularly severe and directly related to land-use changes and insect outbreaks. Australians have an opportunity to make a positive impact on their environment by aiding in the regeneration of native trees. The Greening of Australia project (Nadolny, 1991) aims to plant 1 x lo9 trees during this decade. The ameliorative aspects of dieback research and land-use management in Australia may set a worldwide precedent, illustrating how scientists and landowners can work together to implement sound land management practices.

CONCLUSIONS

lnsect herbivore populations can respond rapidly and dramatically to environmental changes (especially those resulting from natural and anthropogenic disturbances) that stress or increase the abundance of particular plant species in both temperate and tropical forests. Landscapes dominated by susceptible hosts or patches contribute to large-scale population growth of herbivores. Population outbreaks typically are triggered by disturbance or post-disturbance successional processes, but propagate the effects of disturbance into surrounding forest patches, sometimes affecting ecosystem structure and processes over extensive areas. In terms of intensity, severity, scale, frequency, and regularity, herbivore outbreaks are comparable to abiotic disturbances such as fires and storms.

Few studies have addressed relationships between herbivory and disturbance, largely because of limi- tations on measurements of herbivory. Nevertheless, available data suggest that the type of herbivory (e.g., by folivores versus sap-suckers) depends on successional stage in both temperate and tropical forests. Factors triggering elevated levels and herbivore effects on ecosystem processes appear similar between tropical and temperate forests. Furthermore, the scale and severity of herbivory appear to be greater in forests with a history of human intervention than in more

natural forests. Herbivory may be considered as a keystone ecological process in an ecosystem, which can serve as an indicator of disturbance (Lowman, 1999). Low or moderate levels of foliage consumption with infrequent (but occasional) outbreaks typify a healthy forest ecosystem, whereas increased frequency of outbreaks indicates accelerated degradation. Tree condition and density are two major factors affecting herbivory. Both are influenced by disturbance and forest management practices. Forest diebacks have become a serious problem in all regions of the world and reflect a combination of changes in forest conditions, including those that promote insect outbreaks. Forest fragmentation also affects the dispcrsal of herbivore populations across the landscape and creates patches of plants in which herbivores and their predators may not be in balance.

Herbivore effects on plant reproduction, growth and survival, and on the turnover of plants and plant parts, influence a variety of ecosystem processes, including primary productivity, community dynamics, biogeochemical cycling, and canopy-atmosphere-soil interactions. Herbivore effects on these processes are complex, ranging from the alleviation of plant stresses and facilitation of ecosystem recovery under some conditions to the demise of entire forests.

A major challenge is the isolation and experimental testing of the impact of different levels of herbivory over large spatial scales - for instance at the ecosystem level. Given the apparent increase in herbivore outbreaks on vegetation, and the enormous biological and economic implications that such epidemics can have, such research is critical. Several major issues should be addressed in future studies. These include developing standardized methods for measuring long- term herbivory, directly assessing the relationships between disturbance and herbivory, and quantifying the impact of human intervention in ecosystems on herbivory. In particular, future studies must use sufficient replication to compare long-term patterns of herbivory under conditions of experimental disturbance (especially of anthropogenic activities) and recovery.

REFERENCES

Aide, T.M., 1993. Patterns of leaf development and herbivory in a tropical understory community. Ecology, 74: 455466.

Alfaro, R.I. and Shepard, R.F., 1991. Tree-ring growth of interior Douglas-fir after one year's defoliation by DougIas-fir tussock moth. For: Sci., 37: 959-964.

FOREST HERBIVORY INSECTS 267

Alstad, D.N., Edmunds Jr., G.F. and Weinstein, L.H., 1982. Effects of air pollutants on insect populations. Annu. Rev. Entomol., 27: 369-384.

Barbosa, P. and Wagner, M.R., 1989. Introduction to Forest and Shade Tree Insects. Academic Press, San Diego, California, 639 pp.

Benedict, F., 1976. Herbivory Rates and Leaf Properties in Four Forests in Puerto Rico and Florida. Ph.D. Dissertation, University of Florida, Gainesville, Florida.

Boardman, R., 1981. Eucalypt diebacks in South Australia. In: K.M. Old, G.A. Kyle and C.P. Ohmart (Editors), Eucalypt Dieback in Forests and Woodlands, CSIRO Conference Proceedings. Canberra, Australia, pp. 55-58.

Bray, J.R., 1964. Primary consumption in three forest canopies. Ecology, 45: 165-167.

Carroll, C.R. and Hoffman, C.A., 1980. Chemical feeding deterrent mobilized in response to insect herbivory and counter adaptation by Epilachna trededimnotata. Science, 209: 414416.

Carter, J.J., Edwards, D.W. and Humphreys, F.R., 1981. Eucalypt diebacks in New South Wales. In: K.M. Old, G.A. Kyle and C.P. Ohmart (Editors), Euculypt Dkbuck in Forests and Woodlands, CSIRO Con$rence Proceedings. Canberra, Australia, pp. 27-3 1.

Coley, P.D., 1983. Herbivory and defensive characteristics of tree species in a lowland tropical forest. Ecol. Monogr, 53: 209-233.

Coley, PD. and Aide, T.M., 1991. Comparison of herbivory and plant defenses in temperate and tropical broad-leaved forests. In: P.W. Price, T.M. Lewinsohn, G.W. Fernandez and W.W. Benson (Editors), Plunt-Animal Interactions: Evolutionary Ecology in Tropical und Temperate Regions. John Wiley, New York, pp. 25- 49.

Coley, PD., Bryant, J.P. and Chapin 111, F.S., 1985. Resource availability and plant antiherbivore defense. Science, 230: 895-899.

Davidson, D.W., 1993. The effects of herbivory and granivory on terrestrial plant succession. Oikos, 68: 23-35.

Fares, Y., Sharpe, P.J.H. and Magnuson, C.E., 1980. Pheromone dispersion in forests. 1 Theoe Biol., 84: 335-359.

Feeny, P.P., 1976. Plant apparency and chemical defense. In: J.W. Wallace and R.L. Mansell (Editors), Biochemical Interaction Between Plants and Insects. Plenum Press, New York, pp. 1 4 0 .

Ford, H.A., 1989. Ecology @"Birds; An Australiun Perspective. Surrey Beatty, New South Wales, p. 287.

Fox, L.R. and Morrow, PA., 1986. On comparing herbivore damage in Australian and north temperate systems. Aust. 1 Ecol., I I: 387-394.

Furniss, R.L. and Carolin, VM., 1977. Western Forest Insects. USDA Forest Service Misc. Publ. 1339. USDA Forest Service, Washington, D.C., 654 pp.

Futuyma, D.J. and Wasserman, S.S., 1980. Resource concentration and herbivory in oak forests. Science, 2 10: 920-922.

Goheen, D.J. and Hansen, E.M., 1993. Effects of pathogens and bark beetles on forests. In: T.D. Schowalter and G.M. Filip (Editors), Beetle-Pathogen Interactions in Conifer Forests. Academic Press, London, pp. 175-196.

Golley, F.B., 1977. Insects as regulators of forest nutrient cycling. Trop. Ecol., 18: 116-123.

Hadley, K.S. and Veblen, T.T., 1993. Stand response to western spruce budworm and Douglas-fir bark beetle outbreaks, Colorado Front Range. Can. 1 Foe Res., 23: 479491.

Harborne, J.B., 1982. Introduction to Ecological Biochemistry. 2nd Ed. Academic Press, London, 278 pp.

Heatwole, H.H. and Lowman, M.D., 1986. Dieback, Death of an Australian Landscape. Reed Books, Sydney, Australia, 150 pp.

Henvitz, S.R., Slye, R.E. and Turton, S.M., 1994. Lateral shading and the differential interception of angular direct-beam radiation among neighboring tropical rainforest canopy trees. Selbyana, 15: AIO.

Huettl, R.F. and Mueller-Dombois, D. (Editors), 1993. Forest Declines in the Atlantic and Pacific Regions. Springer-Verlag, Berlin.

Huntly, N., 1991. Herbivores and the dynamics of communities and ecosystems. Annu. Rev. Ecol. Syst., 22: 477-503.

Kareiva, P., 1983. Influence of vegetation texture on herbivore populations: resource concentration and herbivore movement. In: R.F. Denno and M.S. McClure (Editors), Variable Plants and Herbivores in Natural and Managed Systems. Academic Press, New York, pp. 259-289.

Kiilsgaard, C.W., Greene, S.E. and Stafford, S.G., 1987. Nutrient concentration in litterfall from some western conifers with special reference to calcium. Plant Soil, 102: 223-227.

Kimmins, J.P., 1972. Relative contributions of leaching, litterfall, and defoliation by Neodiprion sertifer (Hymenoptera) to the removal of cesium-134 from red pine. Oikos. 23: 226-234.

Kruess, A. and Tscharntke, T., 1994. Habitat fragmentation, species loss, and biological control. Science, 264: 1581-1584.

Landsberg, J., 1990. Dieback of rural eucalypts: does insect herbivory relate to dietary quality of tree foliage? Aust. 1 Ecol., 15: 73-87.

Leigh, E.G. and Smythe, N., 1978. Leaf production, leaf consumption and the regulation of folivory on Barro Colorado Island. In: E.G. Leigh, A.S. Rand and D.M. Windsor (Editors), The Ecology ofArborea1 Folivores. Smithsonian Institution Press, Washington, D.C., pp. 33-50.

Leigh, E.G. and Windsor, D.M., 1982. Forest production and regulation of primary consumers on Barro Colorado Island. In: E.G. Leigh, A.S. Rand and D.M. Windsor (Editors), The Ecology of'Arborea1 Folivores. Smithsonian Institution Press, Washington, D.C., pp. 109-123.

Lovejoy, T.E. and Bierregaard Jr., R., 1990. Central Amazonian forests and the Minimum Critical Size of Ecosystems Project. In: A.H. Gentry (Editor), Four Neotropical Rainforests. Yale University Press, New Haven, Connecticut, pp. 60-71.

Lovett, G.M. and Ruesink, A.E., 1995. Carbon and nitrogen mineralization from decomposing gypsy moth frass. Oecologiu, 104: 133-138.

Lovett, G.M. and Tobiessen, P., 1993. Carbon and nitrogen assimilation in red oaks (Quercus rubra L.) subject to defoliation and nitrogen stress. Tree Physiol., 12: 259-269.

Lowman, M.D., 1982. The effects of different rates and methods of leaf area removal on coachwood (Ceratopetalum apetalum). Aust. 1 Bot., 30: 477483.

Lowman, M.D., 1984. An assessment of techniques for measuring herbivory: is rainforest defoliation more intense than we thought? Biotropica, 16: 264-268.

Lowman, M.D., 1992. Leaf growth dynamics and herbivory in five species of Australian rain forest canopy trees. 1 Ecol., 80: 433447.

Lowman, M.D., 1995. Herbivory as a canopy process in rain forest trees. In: M.D. Lowman and N.M. Nadkarni (Editors), Forest Canopies. Academic Press, San Diego, California, pp. 43 1 4 5 5 .


Lowman, M.D., 1999. Life in the Treetops. Yale University Press, New Haven, Connecticut, U.S.A., 206 pp.

Lowman, M.D. and Box, J.H., 1983. Variation in leaf toughness and phenolic content among 5 species of Australian rain forest trees. Aust. J Ecol., 8: 17-25.

Lowman, M.D. and Heatwole, H.H., 1992. Spatial and temporal variability in defoliation of Australian eucalypts. Ecology, 73: 129-142.

Lowman, M.D. and Heatwole, H.H., 1993. Rural dieback in Australia and subsequent landscape amelioration. In: R.F. Huettl and D. Mueller-Dombois (Editors), Forest Decline in the Atlantic and Pacific Region. Springer-Verlag, New York, pp. 307-3 19.

Lowman, M.D. and Wittman, PK., 1996. Forest canopies: methods, hypotheses and future directions. Annu. Reu. Ecol. Syst., 27: 55-8 1.

Lowman, M.D., Burgess, A.D. and Higgins, W.D., 1987. The biomass of New England peppermint (Eucalyptrrs nouu-unglicu) in relation to insect damage associated with rural dieback. Aust .I Ecol.. 12: 361-371.

Lowman, M.D., Moffett, M. and Rinker, H.B., 1993. A technique for taxonomic and ecological sampling in rain forest canopies. Selbyana, 14: 75-79.

Marquis, R.J., 1984. Leaf herbivores decrease fitness of a tropical plant. Science, 226: 537-539.

Mason, R.R. and Luck, R.F., 1978. Population growth and regulation. In: M.H. Brooks, R.W. Stark and R.W. Campbell (Editors), The Douglas-jir Tussock Moth: a Synthesis. USDA Forest Service Tech. Bull. 1585, USDA Forest Service, Washington, D.C., pp. 4 1 4 7 .

Matthews, R.W. and Matthews, J.R., 1978. Insect Behauior. Wiley, New York, 507 pp.

Mattson, W.J. and Haack, R.A., 1987. The role of drought in outbreaks of plant-eating insects. BioScience, 37: 110-1 18.

McClure, MS., 1991. Density-dependent feedback and population cycles in Adelges tsugae (Homoptera: Adelgidae) on Tsuga canadensis. Enuiron. Entomol., 20: 258-264.

Miller, J.C. and Hansen, PE., 1989. Laboratory Feeding Tests on the Deuelopment of' Gypsy Moth Larvae with Reference to Plant Tma and Allelochemicals. Agricultural Experiment Station Bulletin 674, Oregon State University, Corvallis, Oregon, 63 pp.

Miller, K.K. and Wagner, M.R., 1984. Factors influencing pupal distribution of the pandora moth (Lepidoptera: Saturniidae) and their relationship to prescribed burning. Enuiron. Entomol., 13: 4 3 w 3 1.

Mueller-Dombois, D., 1986. Perspectives for an etiology of stand- level dieback. Annu. Reu. Ecol. Syst., 17: 221-243.

Mueller-Domhois, D., 1987. Natural diehack in forests. BioScience, 37: 575-583.

Mueller-Domhois, D., 1990191. Forest decline and soil nutritional problems in Pacific areas. Water Air Soil Pollut., 54: 195-207.

Nadolny, C., 1991. The dilemma of rural tree clearing. Search, 22: 4 3 4 6 .

Nielsen, B.O., 1978. Above ground food resources and herbivory in a beech forest. Oikos, 31: 273-279.

Norton, A,, 1886. On the decadence of Australian forests. Proc. R. Soc. Queensl., 3: 15-22.

Oboyski, P., 1995. Macmarthropod Communities on Vine Maple, Red Alder and Sit& Alder Along Riparian Zones in the Central Western Cascade Range, Oregon. M.S. Thesis, Oregon State University, Cowallis, OR, 49 pp.

Odum, H.T. and Ruiz Reyes, J., 1970. Holes in leaves and the grazing control mechanism. In: H.T. Odum and R.F. Pigeon (Editors), A Tropical Rain Forest. U.S. Atomic Energy Commission, Oak Ridge, Tennessee, pp. 1-69-1-80.

Ohmart, C.P., Stewart, L.G. and Thomas, J.R., 1983. Phytophagous insect communities in the canopies of three Eucalyptus forest types in south-eastern Australia. Aust. J Ecol., 8: 395403.

Old, K.M., Kile, G.A. and Ohmart, C.P. (Editors), 1981. Eucalypt dieback in forests and woodlands. Proceedings o f a Conference o f the Forest Research Diuision. CSIRO, Canberra, Australia, 285

PP. Paine, T.D. and Baker, F.A., 1993. Abiotic and biotic predisposition.

In: T.D. Schowalter and G.M. Filip (Editors), Beetle-Pathogen Interactions in Conifer Forests. Academic Press, London, pp. 61- 79.

Parker, G.G., 1983. Throughfall and stemflow in the forest nutrient cycle. Adu. Ecol. Res.. 13: 58-133.

Parker, C;.G., 1995. Structure and microclimate of forest canopies. In: M.D. Lowman and N.M. Nadkarni (Editors), Forest Canopies. Academic Press, San Diego, California, pp. 73-106.

Parks, C.G., 1993. The Injuence of'lnduced Host Moisture Stress on the Growth and Development of Western Spruce Budworm and Armillaria ostoyae on Grand Fir Seedlings. Ph.D. Dissertation, Oregon State University, Corvallis, Oregon, 154 pp.

Price, PW., 1980. Euolutionury Biology of Parasites. Monographs in Population Biology 15. Princeton University Press, Princeton, New Jersey, 237 pp.

Reichle, D.E., Goldstein, R.A. and Van Hook, R.I., 1973. Analysis of insect consumption in a forest canopy. Ecology, 54: 1076-1084.

Rhoades, D.F. and Cates, R.G., 1976. Toward a general theory of plant antiherhivore chemistry. In: J.W. Wallace and R.L. Mansell (Editors), Biochemical Interaction Between Plants and Insects. Plenum Press, New York, pp. 168-2 13.

Risch, S., 1980. The population dynamics of several herbivorous beetles in a tropical agroecosystem: the effect of intercropping corn, beans and squash in Costa Rica. J Appl. Ecol., 17: 593-612.

Risley, L.S. and Crossley, D.A., 1993. Contribution of herbivore- caused greenfall to litterfall nitrogen flux in several southern Appalachian forested watersheds. Am. Midl. Nut., 129: 67-74.

Roberts, R.J., Campbell, A.J., Porter, M.R. and Sawtell, N.L., 1982. The distribution and abundance of pasture scarabs in relation to Eucalyptus trees. In: K.L. Lee (Editor), Proceedings of the 3rd Australasian Con@rence on Grassland Inuertebrate Ecology. South Australian Government Printer, Adelaide, South Australia, pp. 207-214.

Roland, J., 1993. Large-scale forest fragmentation increases the duration of tent caterpillar outbreak. Oecologia, 93: 25-30.

Romoser, W.S. and Stoffolano Jr., J.G., 1994. The Science of Entomology. 3rd Ed. W.C. Brown, Publishers, Dubuque, lowa, pp. 532.

Salati, E., 1987. The forest and the hydrologic cycle. In: R.E. Dickinson (Editor), The Geophysiology of' Amazonia: Vegetation and Climate Interactions. John Wiley, New York, pp. 273-296.

Schowalter, T.D., 1985. Adaptations of insects to disturbance. In: S.T.A. Pickett and P.S. White (Editors), The Ecology oJ'Natura1 Disturbance and Patch Dynamics. Academic Press, New York, pp. 235-252.

Schowalter, T.D., 1989. Canopy arthropod community structure

FOREST HERBIVORY: INSECTS 269

and herbivory in old-growth and regenerating forests in western Oregon. Can. 1 For Res., 19: 318-322.

Schowalter, T.D., 1994. Invertebrate community structure and herbivory in a tropical rainforest canopy in Puerto Rico following Hurricane Hugo. Biotropica, 26: 3 12-3 19.

Schowalter, T.D., 1995. Canopy arthropod communities in relation to forest age and alternative harvest practices in western Oregon. For. Ecol. Manage., 78: 115-125.

Schowalter, T.D. and Crossley, D.A., 1983. Forest canopy arthropods as sodium, potassium, magnesium and calcium sinks in forests. For. Ecol. Manage.. 7: 143-148.

Schowalter, T.D. and Ganio, L.M., 1999. Invertebrate communities in a tropical rain forest canopy in Puerto Rico following Hurricane Hugo. Ecol. Entomol.. 24: 191-201.

Schowalter, T.D. and Sabin, T.E., 1991. Litter microarthropod response to canopy herbivory, season and decomposition in litterbags in a regenerating conifer ecosystem in western Oregon. Soil Biol. Ferfil., 11: 93-96.

Schowalter, T.D. and Turchin, P., 1993. Southern pine beetle infestation development: interaction between pine and hardwood basal areas. For. Sci., 39: 201-210.

Schowalter, T.D., Webb, J.W. and Crossley, D.A., 1981. Community structure and nutrient content of canopy arthropods in clearcut and uncut forest ecosystems. Ecology, 62: 1010-1019.

Schowalter, T.D., Hargrove, W.W. and Crossley, D.A., 1986. Herbivory in forested ecosystems. Annu. Reu. Entomol., 31: 277-295.

Schowalter, T.D., Sabin, T.E., Stafford, S.G. and Sexton, J.M., 1991. Phytophage effects on primary production, nutrient turnover, and litter decomposition of young Douglas-fir in western Oregon. For. Ecol. Manage., 42: 229-243.

Seastedt, T.R.. Crossley. D.A. and Hargrove, W.W., 1983. The effects of low-level consumption by canopy arthropods on the growth and nutrient dynamics of black locust and red maple trees in the southern Appalachians. Ecology, 64: 1040-1048.

Shure, D..I. and Phillips, D.L., 1991. Patch size of forest openings and arthropod populations. Oecologia, 86: 325-334.

Stanton, N., 1975. Herbivore pressure on 2 types of forests. Biotropica, 7: 8-1 1.

Swank, W.T., Waide, J.B., Crossley, D.A. and Todd. R.L., 1981. Insect defoliation enhances nitrate export from forest ecosystems. Oecologia, 5 1 : 297-299.

Torres, J.A., 1988. Tropical cyclone effects on insect colonization and abundance in Puerto Rico. Actu Cient., 2: 4 0 4 4 .

Torres, J.A., 1992. Lepidoptera outbreaks in response to successional changes after the passage of Hurricane Hugo in Puerto R~co . 1 Trop. Ecol., 8: 285-298.

Trumble, J.T., Kolodny-Hirsch, D.M. and Ting, I.P., 1993. Plant compensation for arthropod herbivory. Annu. Rev. Entomol., 38: 93-1 19.

U.S. Department of Agriculture, 1979. Forest Insect and Disease Management. Forest Service document NA-FR-4. U.S. Government Printing Office, Washington, D.C., pp. 127.

Veblen, T.T., Hadley, K.S., Nel, E.M., Kitzberger, T., Reid, M. and Villalba, R., 1994. Disturbance regime and disturbance interactions in a Rocky Mountain subalpine forest. 1 Ecol., 82: 125-135.

Visser, J.H., 1986. Host odor perception in phytophagous insects. Annu. Rev. Entomol., 31: 121-144.

Webb, W.L., 1978. Effects of defoliation and tree energetics. In: M.H. Brooks, R.W. Stark and R.W. Campbell (Editors), The Douglas- Fir Tussock Moth: a Synthesis. USDA Forest Service Tech. Bull. 1585, USDA Forest Service, Washington, D.C., pp. 77-8 1.

Wickman, B.E., 1980. Increased growth of white fir after a Douglas- fir tussock moth outbreak. 1 For., 78: 31-33.

Willig, M.R. and Camilo, G.R., 1991. The effect of Hurricane Hugo on six invertebrate species in the Luquillo Experimental Forest of Puerto Rico. Biotropica. 23: 455461 .

Wint, G.R.W., 1983. Leaf damage in tropical rain forest canopies. In: S.L. Sutton, T.C. Whitmore and A.C. Chadwick (Editors), Tropical Rain Forest: Ecology and Management. Blackwell, Oxford, England, pp. 229-240.

Witcosky, J.J., Schowalter, T.D. and Hansen, E.M., 1986. The influence of time of precommercial thinning on the colonization of Douglas-fir by three species of root-colonizing insects. Can. 1 For. Res., 16: 745-749.

Wolda, H. and Foster, R., 1978. Zunacetha annulata (Lepidoptera: Dioptidae), an outbreak insect in a Neotropical forest. GeoEcoTrop, 2: 443454.

Wong, M., Wright, S.J., Hubbell, S.P. and Foster, R.B., 1990. The spatial pattern and reproductive consequences of outbreak defoliation in Quararihea a.sterolepis, a tropical tree. 1 Ecol., 78: 579-588.

Documents

ECOSYSTEMS OF DISTURBED GROUND - CanopyMeg - Official Website