Download early primary succession on mount st. helens: impact of insect

Ecology, 83(1), 2002, pp. 191–202
q 2002 by the Ecological Society of America
EARLY PRIMARY SUCCESSION ON MOUNT ST. HELENS:
IMPACT OF INSECT HERBIVORES ON COLONIZING LUPINES
JOHN G. BISHOP1
Department of Botany, University of Washington, Seattle, Washington 98195-5325 USA
Abstract. Lupinus lepidus var. lobbii, the earliest plant colonist of primary successional
habitats at Mount St. Helens, can dramatically influence successional rates and ecosystem
development through N fixation and other facilitative effects. However, 15 yr after the
eruption, lupine effects remained localized because high rates of population growth in newly
founded patches (l 5 11.2, 1981–1985) were short lived ( l 5 1.51, 1991–1995), despite
widespread habitat availability. To investigate this paradox, I examined 60 colonizing lupine
patches, ranging from low-density patches at the edge of the expanding lupine population,
to high-density patches in the population core, including 41 patches created in 1992. Survival, reproduction, and herbivory (for a subset of plants and years) were measured on
;12 000 plants for up to 5 yr (1991–1995).
Stem-boring, leaf-mining, and seed-eating lepidopterans and anthomyiid flies, feeding
in edge patches and in low-density margins of core patches, strongly affected edge patch
demography, but not that of central core areas. For example, in 1994–1995, 77% of edge
plants were afflicted by tortricid stem borers vs. 24% in the core, and from 1993–1995
gelechiid leafminers infested 68% of plants in the youngest edge patches, vs. only 8% at
the core. Associated adult mortality was 88%, compared to ;30% when absent. Seed
predators consumed 36% (range 4–92%, 1991–1995) of seeds in both core and edge patches,
with seed loss negatively correlated with seedling number the following year. In contrast
to edge patches, resource-dependent seedling mortality appears to govern short-term dynamics in high-density core areas. In conclusion, edge region herbivore effects can account
for the observed 7.5-fold difference in l between patches founded 10 yr apart. Locally,
increased lupine mortality creates access to high-quality sites, facilitating succession, but
at larger spatial scales diminished population growth is likely to retard facilitation and
succession.
Key words: colonization; demography; facilitation; herbivory; invasion; legume; Lepidoptera;
Lupinus; Mount St. Helens (Washington, USA); primary succession; seed predation.
INTRODUCTION
For many colonizing plants, establishment in primary successional habitats depends upon amelioration
of physical conditions, either abiotically or through the
facilitative effect of other species. For example, studies
of soil development conclude that nitrogen is in short
supply early in primary succession, a deficiency that
appears especially striking on soils deriving from volcanic ash and lava flows, where N may be undetectable
initially (Vitousek et al. 1987, del Moral and Bliss
1993). In such systems, the presence of nitrogen-fixing
species is capable of radically altering the pace and
pattern of ecosystem and community development
(Clements 1916, Connell and Slatyer 1977, Marrs et
al. 1983, Vitousek et al. 1987, Vitousek and Walker
1989). Under the facilitative model of succession, the
rates of population increase and spatial spread of nitrogen-fixing species critically affect the rate of sucManuscript received 1 November 1999; revised 27 May 2000;
accepted 30 June 2000; final version received 15 January 2001.
1 Present address: School of Biological Sciences, Washington State University, 14204 NE Salmon Creek Avenue,
Vancouver, Washington 98686 USA.
E-mail: [email protected]
cession, and it is therefore of interest to understand the
factors that control the population dynamics of these
species.
Consumers of nitrogen-fixing species, e.g., herbivores, may positively impact rates of succession by
liberating resources, or they may decrease rates by demographically preventing facilitation. Most studies examining consumer effects on primary succession involve marine systems, where inhibition and tolerance
models of succession may be more important than facilitative models (Lubchenco and Gaines 1981, Farrell
1991, Sousa and Connell 1992, Hixon and Brostoff
1996). These studies provide examples of consumermediated deceleration, acceleration, and deflection of
successional rates and trajectories (Hixon and Brostoff
1996). The only study demonstrating herbivore effects
on terrestrial primary succession is that of Bach (1994),
showing that on lake shore sand dunes, chrysomelid
beetles specializing on willow decreased host abundance, thereby changing community trajectories. Currently, herbivores are not considered an influential factor in terrestrial primary succession (e.g., Walker and
Chapin 1987, del Moral and Wood 1988, Wood and
Anderson 1990, Miles and Walton 1993). This is par-
191
192
JOHN G. BISHOP
Ecology, Vol. 83, No. 1
PLATE 1. Lupinus lepidus in low-density edge regions attacked by (a) stem-boring Hysticophora spp. (foreground) and
(b, c) leaf-mining Chionodes spp., visible as light-colored webbing and woven gravel retreats.
ticularly true for invertebrate herbivores, whose ability
to more generally influence plant population and community dynamics is a topic of debate (Brown et al.
1987, 1988, Crawley 1989, Gange et al. 1989, Ehrlén
1995, Louda and Potvin 1995, Strong et al. 1995, Louda
and Rodman 1996, Root 1996, Maron 1997).
In this paper, I describe patterns of herbivory and
the demographic consequences of that herbivory in a
population of the legume Lupinus lepidus var. lobbii
that is colonizing the primary successional landscape
created by Mount St. Helens’ 1980 eruption. In 1981,
several individual lupines colonized the otherwise barren north slope of Mount St. Helens from remnant populations elsewhere on the volcano. For the next several
years the lupine population spread rapidly outward
from this initial invasion focus and lupines were the
most successful colonist of upland primary successional habitats (del Moral et al. 1995). By 1990, the
lupine population comprised a central core region of
extremely high lupine density surrounded by numerous, low-density edge patches up to 2 km from the core
(Fig. 1; del Moral et al. 1995, Fagan and Bishop 2000).
Intervening areas between the edge and core patches,
as well as the expanding margins of core patches, contain low-density, patchily distributed lupine. By 1992,
population growth rates (l) in recently founded, lowdensity edge region patches had dropped to an average
of l 5 1.51 (mean of 13 edge patches studied from
1991 to 1995), far below those observed during the
early stages of colonization (l 5 11.2, for the original
colonizing patch, 1981–1985; Fagan and Bishop 2000).
L. lepidus on Mount St. Helens has been the focus
of numerous studies demonstrating its ability to ameliorate physical conditions. For example, L. lepidus can
fix more than 7 kg·ha21·yr21 of N, compared to background rates of only 2 kg·ha21·yr21 of N, and dramatically increases organic matter accumulation (Kerle
1985, Halvorson et al. 1991, Halvorson et al. 1992).
Live lupines competitively suppress other ruderal species, whereas dead lupines facilitate these same species
(Morris and Wood 1989, del Moral and Bliss 1993,
Titus and del Moral 1998a). However, L. lepidus’s potential impact on community development has been
tempered by its surprisingly slow rate of spread in recent years, associated with decreased l (Halvorson et
al. 1992, del Moral and Bliss 1993).
A number of hypotheses could explain recent decreases in rates of lupine population increase and spatial expansion. For example, dispersal limitation,
caused by lupines’ relatively large seeds (4–8 mg),
could curtail spatial spread (Wood and del Moral 1987).
However, long-distance dispersal does occur. Founder
individuals were .4 km from the nearest surviving
patches, in 1991 edge patches existed .1 km from the
January 2002
HERBIVORY, DEMOGRAPHY, AND COLONIZATION
193
FIG. 1. Map of study patches on the Pumice Plains and debris avalanche. Locations were obtained by standard survey
methods using a TopCon Total Station. Patches with dates denote location and founding dates of core patches. Contour
intervals are 14.7 m. Elevations are in feet (3600–4400 feet 5 1100–1340 m). The crater is 1.5 km southeast. Large solid
circles 5 core patches; small solid circles 5 young edge, experimentally established; solid squares 5 young edge, natural;
solid triangle 5 old edge; diamonds 5 source of transplants. Southern 1984 core patch is drawn to scale. The fourth core
patch is ;700 m north-northeast of the 1981 patch. Inset shows location of Mount St. Helens National Volcanic Monument,
denoted by open rectangle. Tick marks in inset represent 468159 N latitude and 1228109 W longitude.
nearest seed source (personal observation), and seeds
are occasionally deposited in traps many meters from
lupine populations (del Moral and Wood 1988, Wood
and del Moral 1988, Wood and del Moral 2000). Thus,
while dispersal limitation may play a role, it is not an
adequate explanation for changes in spread rates. Another plausible explanation is that lupines colonized
most suitable habitat shortly after the eruption, and fare
poorly in currently available habitats. This explanation
is also inadequate. L. lepidus is characteristic of highaltitude pumice communities, and resource-poor gla-
cial moraines (Franklin and Dyrness 1973, Kruckeberg
1987). Following the eruption, L. lepidus colonized
substrates 200–600 m below pre-eruption elevations,
.1000 m below its elevation elsewhere in the region.
At these lower elevations, decreased drought stress and
longer growing seasons are known to increase lupine
growth rates, size, and fecundity relative to plants at
pre-eruption elevations (Braatne and Bliss 1999).
Another plausible hypothesis is that decreased l is
the demographic consequence of insect herbivory.
High levels of seed predation and other insect herbivore
JOHN G. BISHOP
194
damage have been documented in L. lepidus at Mount
St. Helens (Halvorson et al. 1992, Bishop and Schemske 1998, Fagan and Bishop 2000; see Plate 1). In
addition, short-term removal experiments demonstrate
that these herbivores can reduce population growth and
rates of spread (Fagan and Bishop 2000). Here I present
the results of a large-scale (60 patches), long-term (5
yr) demographic survey of core and edge patches that
allows extrapolation from the experimental results of
Fagan and Bishop. Specifically, I describe patterns of
herbivory that differ between core and edge patches
and that are closely related to host density. I also use
extensive demographic data to ask whether the damage
by different herbivore guilds likely affects lupine demography (survivorship, fecundity, and population
growth), and whether these effects vary among core
and edge portions of the population.
METHODS
Species and study site
The study site was formed on 18 May 1980 by a
massive debris avalanche, followed by pyroclastic
flows (incandescent gas and pumice), lahars, and tephra
deposition. The lupine patches in this study lie in a 60km2 primary successional landscape, atop 4–80 m of
newly emplaced rock, and 300–1500 m from the nearest secondary successional habitats. Thus, there was
no pre-eruption seed bank, herbivores, or other biotic
legacy surviving in these sites. The study patches are
scattered across 6 km2, with an elevation range of
1150–1350 m (3700 ft–4400 ft in Fig. 1). Environmental conditions on the Pumice Plains are stressful;
the pumice and rock substrates initially had little waterholding capacity, few nutrients, and nitrogen was undetectable (Nuhn 1987, Halvorson et al. 1991, 1992,
del Moral and Bliss 1993). Although yearly precipitation is high (2.4 m), there is pronounced summer
drought (rainfall in July and August is frequently ,10
mm/mo), and surface temperatures reach 508C (Reynolds and Bliss 1986). Lupinus lepidus possesses several
features, e.g., the ability of seedling roots to quickly
reach moister subsurface soil, that allow it to avoid
drought stress (Braatne and Bliss 1999).
Mature plants of L. lepidus var. lobbii are ,15 cm
high, with prostrate herbaceous stems spreading up to
a 45-cm radius from a fibrous caudex. L. lepidus is
self-compatible, and pollen addition experiments in
large patches show that fruit set is not pollinator limited
(J. G. Bishop, unpublished data). Seeds are explosively
dehisced, but have no obvious adaptations for secondary dispersal, moving primarily by wind and during
snowmelt. The flowers, flower stalks, and fruits are
frequently attacked by lupine aphids (Macrosiphum albifrons), pentatomid bugs (Chlorochroa ligata), larvae
of an anthomyiid fly (Crinurina sp.), and by two lepidopteran species specializing on lupines, the microlepidopteran Schinea sueta and the lycaenid Plebejus
Ecology, Vol. 83, No. 1
icarioides montis (Bishop and Schemske 1998). Larvae
of several lepidopteran species also mine the leaves
(Gelechiidae: Chionodes spp.) and bore into the caudex
(Tortricidae: Hystricophora sp.); these are also lupine
specialists and rarely attack adjacent L. latifolius. Damage by leafminers is easily recognized as a woven mass
of whitened or yellowed leaves, and damage by caudex
borers as a gray pallor accompanied by wilting (see
Plate 1); both types of damage are readily quantified.
Damage to fruits and seeds by chewing insects is more
subtle and must be assessed through dissection.
Field methods
Twelve patches (3 core, 9 edge) were selected for
demographic study in 1991, and 7 additional patches
(1 core, 6 edge) were selected in 1992. Core patches,
founded 1981–1984, include the original colonizing
patch and several high-density satellites. Edge patches
are widely spaced secondary or tertiary foci, generally
located uphill of the core, and can be divided into two
age groups: those founded 1987–1988 (‘‘old edge’’)
and those founded 1990–1991 (‘‘young edge’’). These
19 patches represent a substantial proportion of the core
and older edge patches that existed by 1991. Extensive
surveys in 1994 identified only three additional large
patches in the region illustrated in Fig. 1, and about
40 additional patches among edge patches, mostly comparable to young edge patches.
To increase the sample size of newly-founded patches and to examine the probability of patch establishment, I established 41 additional edge patches prior to
the start of the 1992 growing season. Each experimental
patch consisted of eight 1-yr-old individuals arranged
in a circle 1 m in diameter. Founders were from one
or two nearby patches (Fig. 1), representing likely
sources of natural colonists. Sites were chosen without
regard to microsite characteristics, except that they
were not placed within gullies and were located .30
m from existing patches so as to avoid external demographic influence. These patches are similar in density, percent cover, and size distribution to natural edge
patches, so they are grouped with ‘‘young edge’’ patches.
To provide an indication of successional status, percent cover of lupines and percent cover of all other
plants were estimated in 1993 for each patch by estimating the number of 10 3 10 cm squares occupied
by each species. Differences among patch types (core,
old edge, young edge) were examined using Tukey tests
for multiple comparisons (Zar 1984).
All edge patches were censused by establishing a
permanent grid around the patch. In core patches, because of their much higher density, five or nine 25 3
25 cm quadrats were placed randomly along transects
running from the periphery to the center of the patch.
An additional four or seven 1 3 1 m quadrats were
located at patch margins in order to sample the lowdensity periphery of the core. The number of quadrats
January 2002
HERBIVORY, DEMOGRAPHY, AND COLONIZATION
varied because additional quadrats were placed in larger patches. When increased patch size made sampling
entire patches and quadrats infeasible, sections of grids
or quadrats were subsampled and some patches were
dropped from the study (2 in 1994, and 13 in 1995).
All patches were censused in late August (15–28 August), but plants were usually located and tagged in
June and July. All plants were marked, and their location determined to the nearest centimeter, to allow
relocation. Plant diameter, which correlates strongly
with biomass (Pearson’s r 5 0.95, df 5 49, P ,
0.0001), and herbivore damage were measured at each
census.
Herbivory.—Foliage damage by leaf-mining
Chionodes caterpillars, which is conspicuous, was
rare until 1993. It was quantified first as present or
absent (1993), then, in 1994 –1995, on a 0 –5 scale
of leaf destruction: 0 5 0 –5%, 1 5 6–25%, 2 5 26–
50%, 3 5 51–75%, 4 5 76–95%, 5 5 95–100% destruction). Caudex-boring Hystricophora sp. was not
apparent from 1990 to 1992, but in 1993–1994 it
caused nearly 100% mortality in several patches not
under study. Mortality caused by Hystricophora between August 1994 and July 1995 was quantified by
surveying live plants and plants that had died since
August 1994 for Hystricophora presence. Because
dead lupines take . 1 yr to decay, and because Hystricophora forms extensive galleries in the root and
caudex, stem borer presence was easily identified in
summer, 1995. Its presence in live plants is detected
by wilting, gray pallor, and the accumulation of frass
surrounding holes in the caudex.
Reproduction.—I censused each plant’s infructescences and the number of fruits on a sample of seven
infructescences per plant, if available, at roughly 2-wk
intervals. In addition, a sample of infructescences was
collected to estimate seeds per fruit and damage by
seed predators. Sample sizes fluctuated because (a) in
some years (i.e., 1993) weather conditions resulted in
little or no reproduction, and (b) to minimize demographic effects, fruits were not collected from patches
with low fruit set. Overall for the years 1991–1996,
seed production and damage was measured for 2957
fruits from 347 plants from core patches, and 2594
fruits from 369 plants from edge patches.
Fruits were checked for (a) the presence of damage,
indicated by an entry or exit hole produced by a seed
predator; (b) the total number of mature or nearly mature seeds, not counting aborted seeds; and (c) the number of these seeds damaged by chewing. Not counting
aborted seed results in underestimation of insect damage, because many are caused by inflorescence-attacking hemipterans and homopterans. Seed production before predation was estimated for each plant by multiplying total fruits (known for every plant) by average
seeds per fruit (from that patch or a nearby one of the
same patch type). Using average seeds per fruit to calculate seeds per plant is reasonable, as most of the
195
interplant variation within patches in seed production
is due to fruit number rather than seeds per fruit (Bishop
and Schemske 1998). The number of seeds matured
(i.e., after predispersal predation) was estimated by discounting each plant’s seed production by estimates of
proportion of seeds damaged by chewing insects.
Analyses.—For each of the three types of herbivore
damage, leaf mining, caudex boring, and seed predation, there are two classes of analyses. I first asked
whether each type of herbivory differs between patch
type (core and edge), and whether there is an additional
effect of plant density. Second, I analyzed the effects
of patch type, herbivory, and density, on mortality, seed
and seedling production, and population growth. Regression analyses included year as a covariate, when
appropriate. In edge patches, density was calculated by
defining area as the rectangle bounding the outermost
plants. Variables were transformed to improve model
fit, but not to de-correlate residuals. Instead, significance levels for multiple and stepwise regressions were
determined by randomization tests performed as follows. For each regression analysis, dependent variable
data were randomly permuted among cases 4999 times,
while holding the independent variable data constant,
to create 4999 new data sets. The regressions were
repeated for each permuted data set, and the P value
was calculated as [rank of the actual regression coefficients among 4999 random ones 4 5000] (Manly
1991). Table 3 lists dependent and independent variables. Proportion of seeds damaged in each patch was
analyzed with ANOVA using sequential sums of
squares, with year and patch type as factors. Significance levels were determined using randomization tests
as for the regression analysis, with P values calculated
as the [rank of the actual F statistic 4 5000] (Manly
1991). All analyses were done using S-Plus 3.3 (Statsci
1994).
Logistic regressions were used to determine (1) the
effect of patch type (core vs. edge), density, and plant
diameter on Hystricophora infestion (1994–1995 only,
for four edge and four core patches) and (2) whether
leaf damage or Hystricophora infestation affected the
probability of surviving until the next census, with
patch type, year, and plant diameter as covariates. For
years in which degree of leaf damage was measured,
logistic regression was used to determine whether
plants with the most damage (.74%) had an increased
probability of death. Survivorship distributions (probability of surviving until a given age) were estimated
for 1991–1995 for all patch type and cohort combinations using Kaplan-Meier estimates for 5246 plants
(Kalbfleisch and Prentice 1980). Effects of patch type
and cohort on survivorship distributions were analyzed
with an accelerated life model, assuming a Weibull
distribution. Details of this analysis are available in
Appendix A. Appendix B contains photographs showing the impact of insect herbivores.
JOHN G. BISHOP
196
Ecology, Vol. 83, No. 1
TABLE 1. Site and sample characteristics by patch type. Statistics are reported as median (range).
Site
Patches
surveyed
Year
founded
Area sampled
(m2)
Edge, young
Edge, old
Core
50
6
4
1990–1991
1987–1988
1981–1985
90 (48–600)
225 (144–300)
4 (1–10)
Initial
population
size
Density
(plants/m2)
3 (1–24)
131 (35–160)
.40,000
0.13 (0.001–1.70)a
1.01 (0.03–5.06)b
73 (0.25–1622)c
Notes: Number of plants sampled does not include dead plants. Values of density and percent cover with different superscript
letters are significantly different, P , 0.01.
RESULTS
density and herbivory results from herbivory concentrated at the margins of the patch.
Similarly, though quantified only in 1994–1995,
damage by caudex-boring Hystricophora increased
through the study period, and was much higher in edge
patches (Table 4), and Hystricophora presence was positively density dependent in edge patches but negatively density dependent in core patches. Logistic regression indicated that core plants at 1.16-m spacing
were 153 more likely to be infested by Hystricophora
than plants at 0.16 m, whereas the odds of infestation
for widely spaced plants at the edge is 0 relative to
plants at 0.16-m spacing (odds 5 probability of infestation 4 probability of no infestation; Table 4).
Seed production and predation.—Seed production
per plant (post predation) decreased dramatically
through the study period (compare 1991–1992 with
1993–1995 in Table 2). The presence of Chionodes
leafminers, which increased over time, was associated
with 32 fewer seeds per plant (Table 3, line 4), a 97%
decrease compared to 33 6 71 seeds per plant (mean
6 1 SD of all populations, after seed predation). Seed
loss to predispersal seed predators, primarily lycaenid
caterpillars (Plebejus icarioides) and anthomyiid fly
larvae (Crinurina spp.), averaged 36% across all populations and years. ANOVA and post hoc comparisons
indicated that mean seed loss was greater in core patches and decreased over time (effect of patch type: F 5
49, df 5 1, P , 0.0001; effect of year: F 5 11, df 5
4, P 5 0.001; patch type 3 year interaction: F 5 6, df
Percent cover of lupines was .11003 higher in the
core than at the edge; percent cover of other species
was 2653 higher at the core than at the edge, and
median patch densities were 5613 greater in the core
than in young edge patches (Table 1). These results
provide independent support for lumping patches into
classes representing different stages of colonization.
Note that the densities in edge patches appear too low
to invoke either intraspecific or interspecific resource
competition as an explanation for the 7.5-fold decrease
in population growth in these regions (Table 2). All 41
transplant patches survived the first growing season,
with a median of 3 survivors per patch, and 38 of 41
survived for 6 yr. Natural and transplant young edge
patches were similar in percent cover (both 0.4%) and
lupine density (mean 6 SD; 0.03 6 0.1 plants/m2 vs.
0.08 6 0.05 plants/m2 in 1992).
Foliage and caudex damage.—Foliage damage by
leaf-mining Chionodes caterpillars increased dramatically through the study period. While leaf damage was
not apparent in 1991–1992, by 1995 a median of 68%
of plants were damaged in young edge patches, and the
median damage score averaged across patches was 1.85
(2 5 26–50%, Table 2). Edge patches suffered much
greater damage than core patches (Table 2, Fig. 2), and
increasing density of L. lepidus resulted in decreased
damage at both the core and the edge, even with year
and patch type entered first as covariates (Table 3, lines
1 and 2). At the core, this negative relationship between
TABLE 2. Mean of population and quadrat means (6 1
SD ).
Edge
Measure
Years
Younger
Older
Core
1993–1995
0.68 6 0.4
a
0.48 6 0.3b
0.08 6 0.1c
Proportion plants damaged
Average area damage
1994–1995
1.85 6 1.1a
1.43 6 1.5a
0.18 6 0.3b
Seed number
1991–1992
1993–1995
231 6 425
51 6 120a
174 6 358
7 6 33b
b
161 6 394b
35 6 60ab
l ( 5 e r)
1981–1985
1991–1995
NA
1.51 6 0.75
NA
1.46 6 0.93
11.3 6 2.7
0.91 6 0.19
a
Notes: Plant damage refers to leaf mining by Chionodes spp. Average area damaged is the mean leafminer damage class
on a scale of 0 (0–5%) to 5 (95–100%). Values in the same row with different superscript letters are significantly different
(P , 0.05) after applying variable-wide sequential Bonferroni correction. Population growth rate (l) is the median of the
geometric mean Nt11/Nt for each patch taken across years. Core data from 1981–1985 are courtesy of C. M. Crisafulli. NA
5 not applicable.
January 2002
HERBIVORY, DEMOGRAPHY, AND COLONIZATION
197
TABLE 1. Extended.
Percent
cover lupine
0.03 (0.01–0.1)a
0.06 (0.02–0.15)a
66.3 (4–92)b
Percent
cover other
plants
0.12 (0.00–1.9)a
0.12 (0.06–0.5)a
31.8 (0.00–108.8)c
Number of live plants sampled
1991
1992
1993
1994
1995
41
399
291
213
603
757
1300
1715
1894
1433
1828
1496
1148
98
298
5 4, P 5 0.018; P-values estimated by randomization
test; total df 5 215). The strong patch type 3 year
interaction corresponds to the observation that over 7
yr of observations, percent fruit damage in core patches
undergoes a damped oscillation between 4% and 82%,
whereas damage in edge patches declines after 1991
(Fig. 3). Regression analysis reveals that the effect of
seed predation carries over into the next year’s seedling
abundance. Seedlings per square centimeter of leaf area
were significantly negatively affected by the proportion
of fruits damaged, but not affected by the total number
of seeds (before predation) in the previous year (Table
3, line 5).
Survivorship and annual mortality.—The odds of dying increased through the study period and were twice
as great in core patches as in edge patches, corresponding to a decrease in survivorship from 0.65 to 0.32
(logistic regression: effect of each additional year on
odds 5 4.413, P , 0.0001; effect of edge vs. core 5
1.933, P , 0.0001; N 5 8867; effect indicates increase
in odds of death). Logistic regression did not detect
significant effects of Chionodes presence on mortality
from 1993 to 1995, an effect that may have been
masked by Hystricophora, but did indicate that a leaf
damage score .4 (.75% leaf loss, N 5 109) raised
the overall probability of death in 1994–1995 from 0.80
to 0.90 (P 5 0.03, df 5 2261).
Graphical inspection of age-specific survivorship
curves suggests that patch type, cohort, and year have
strong effects on survivorship (Fig. 4). These conclusions are supported by the accelerated life model analysis for patch type and cohort, which shows decreasing
survivorship in younger patches and later cohorts (P
, 0.0001 for each level of each factor, df 5 5239, r2
5 0.41; see Appendix A for details). Of particular note
is that seedling mortality (as indicated by the transition
from the first to the second growing season) was 40%
higher in core patches than in edge patches (excluding
core 1991 and edge 1994). In contrast, postseedling
mortality was comparable between patch types. The
pattern at the core is attributable to high lupine establishment and rapid growth of other species in 1991,
followed by low establishment thereafter. In 1994 survivorship was low in all patch types for all cohorts
(Fig. 4) demonstrating that high mortality late in the
study is not attributable to a deterministic 4 or 5 yr
lifespan. In edge patches this appears to be due to Hystricophora-caused mortality, whereas in core patches
Hystricophora was less common (Table 4), and survivorship is comparable to previous years.
Population growth (l 5 Nt11/Nt) was higher on average in edge patches than in core patches, and highest
in the young edge patches (Table 2), but even in edge
patches l averaged seven times less than in the original
patch 10 yr earlier. l fluctuated greatly between patches
and years, with r (5 loge l) ranging between 21.2 and
.5 in edge patches in 1992–1993 alone (data not
shown, see Fig. 5 for range over all years). In addition,
population growth appears to be strongly affected by
density (Fig. 5), even when patch type and year are
entered first as covariates (Table 3, line 3). In other
words, negative density dependence is as strong across
the range of densities found in young edge patches
(median density 5 0.13 plants/m2) as it is in core patches (median density 5 73 plants/m2).
DISCUSSION
FIG. 2. Magnitude of leaf damage caused by Chionodes
spp. (patch averages), vs. density. The change of damageclass scale on the bottom graph reflects a difference in scoring
method between 1993 (present/absent) and later years [scale
of 0 (no damage) to 5 (100% damage)]. Note logarithmic
plant density scale. Lines are local regression (LOESS) estimates.
Stem-boring, leaf-mining, and seed-eating lepidopterans and anthomyiid flies, feeding in edge patches
and in the low-density margins of core patches, markedly increased mortality and decreased propagule production. This conclusion, in combination with the experimental results of Fagan and Bishop (2000), con-
JOHN G. BISHOP
198
Ecology, Vol. 83, No. 1
TABLE 3. Results of regression analyses of leafminer damage and demographic rates.
Independent variables†
Year
Dependent variable
Coeff.
A) Leafminer damage
Plants damaged by Chion- 20.03
odes, 1993–1995 (%)
Average Chionodes dam0.08
age, 1994–1995
r in patch or quadrat
20.33
[5loge(l)]
B) Lupine seed production
Seed production per reproductive plant§\
C) Seedling density
Seedlingst11/cm2 of lupine\
Patch type
P
Coeff.
P
Density (Nt /m2)‡
Coeff.
P
Patch type
3 Density
Coeff.
P
r2
N
0.16
0.34
0.0002
20.09 ,0.0001
20.06
0.008
0.56
236
0.61
0.93
0.0001
20.33 ,0.0001
20.30
0.004
0.51
143
1.97 ,0.0001
20.21 ,0.0001
20.22
0.008
0.33
249
,0.0001
1.4
0.0006
0.01
0.64
224.6
214.5
20.08
0.0002
0.005
0.006
232.2
20.12
0.0002
0.02
3.2
0.0002 0.27 2023
20.001 0.87
0.16
109
Notes: Results are from stepwise regression with variables entered in order from left to right; patch type is a dichotomous
variable with core 5 0, edge 5 1, except as noted. Listed under each independent variable are regression coefficients (Coeff.)
and P values.
† All significance levels determined by randomization tests are based on regressions of 4999 permuted data sets.
‡ Variable was loge-transformed to improve model fit.
§ Top coefficient is for old edge relative to young edge, bottom values are for core relative to young edge.
\ Multiple regression.
tradicts the view that insect herbivores do not influence
colonization by native species or of primary successional habitats, suggesting instead that specialist insect
herbivores restrict population growth and spatial spread
of lupines colonizing Mount St. Helens. For example,
in 1994, 77% of edge plants were infested with the
Hystricophora stem borer, and infested plants had 97%
mortality, compared to 59% for uninfested plants and
25–40% for edge plants in the previous 3 yr (Fig. 4).
TABLE 4. (A) Hystricophora infestation and associated percentage mortality in edge vs. core patches; (B) results of
logistic regression analysis for the effect of patch type and
nearest-neighbor distance (NN) on Hystricophora presence.
Mortality in 1995 not caused by Hystricophora was
partly attributed to the Chionodes leafminer, which infests the majority of edge plants and removes .25%
of leaf area on average (Fig. 2, Table 2). Herbivores
also curtailed edge plant seed production. In 4 of 6 yr,
pre-dispersal seed predators removed 37–69% of maturing seeds (Fig. 3) and directly affected seedling
numbers the next season. Moreover, leaf damage by
Chionodes prevented many plants from producing any
seeds. Because established seedlings often have high
survivorship in edge patches (Fig. 4), and L. lepidus
A) Hystricophora infestation
Category
Infested
Uninfested
Edge (N 5 703)
Core (N 5 345)
Plants
(%)
Mortality
(%)
Plants
(%)
Mortality
(%)
77
23
97
59
24
76
95
93
B) Regression analysis (R2 5 0.78, N 5 1048)
Variable
Intercept
Edge relative to core
NN at core (m2)
NN at edge (m2)
Coefficient
Odds
P
0.29
0.47
2.71
210.04
1.60
15.03
0.00
0.22
0.000
0.000
0.000
Notes: (A) The ‘‘Plants’’ columns report the percentages
of infested and uninfested plants. (B) Coefficients indicate
the increment in log odds. ‘‘Odds’’ refers to the odds of
infestation relative to core or edge plants with 0.16-m spacing.
For example, edge plants are 1.63 more likely to be infested
than are core plants.
FIG. 3. Time series of percentage fruit damaged with
pointwise 95% binomial confidence intervals. Estimates are
means of plant means. Values for 1990 are anecdotal, but
based on a large number of fruit dissections (Bishop and
Schemske 1998); values for 1996 are from Fagan and Bishop
(2000).
January 2002
HERBIVORY, DEMOGRAPHY, AND COLONIZATION
199
Edge vs. core
Herbivory in core patches was lower compared to
edge patches and had relatively little demographic impact. Hystricophora and Chionodes caterpillars were
scarce in core patches. When present, they were concentrated in the lower density margins of the core
patches, where their presence might easily be overlooked, yet where models predict the greatest effects
of decreased propagule production on spatial spread.
Fagan and Bishop (2000) found no effect of herbivore
removal on l in core patches, but their experiment was
not designed to distinguish between central and marginal zones of the core.
Greater herbivory in edge patches and core margins,
compared to the denser center of the core, runs counter
to expectations based on resource concentration,
wherein the densest host populations sustain the highest
herbivore density. While negative relationships between host density and herbivory are not uncommon,
the underlying mechanisms remain obscure (Kunin
1999). It seems unlikely here that greater herbivore
load at low-density is attributable to the larger basin
of attraction of each plant, because the herbivores are
FIG. 4. Survivorship curves: comparison of patch type
and cohort (year of germination); N 5 8867 plants. Age 1 is
the end of the first growing season for each cohort. A small
fraction of plants survive past 5 yr.
has a short-lived seed bank (Bishop 1996), the loss of
seeds may directly decrease recruitment to adult stages.
Removal of the Chionodes leafminer and several less
abundant herbivores from two edge patches in 1995
produced a temporary fivefold increase in population
growth rate (l), mediated primarily by increased plant
growth and seed production (Fagan and Bishop 2000).
Our combined studies show that such leafminer effects
occur throughout edge region patches and have persisted from 1993 to 1999. Fagan and Bishop did not
find any significant effect of herbivore removal on mortality, nor of predispersal seed predators in edge patches. The current study suggests that the fivefold effect
of removal actually underestimates the impact of insect
herbivores because both pre-dispersal seed predation
and Hystricophora-caused mortality are frequently
greater than they were in 1995–1996. Thus, herbivore
impacts likely account for the observed 7.5-fold decrease in l compared to initial rates of l 5 11.3 (Table
2). Models of spatial spread predict that diminished l
at the edge of an expanding population will greatly
affect spread rates, suggesting that herbivory in edge
regions may explain L. lepidus’ failure to rapidly exploit available habitat (Skellam 1951, Moody and Mack
1988, Andow et al. 1990, Fagan and Bishop 2000).
FIG. 5. Annual population growth (r 5 loge l) for patches
and quadrats as a function of density, by patch type, for 1991–
1995. Line slopes are from multiple regression; horizontal
lines denote r 5 0. For the abscissa label, the original units
of density were Nt /m2.
200
JOHN G. BISHOP
nearly absent from the central core areas, and because
this pattern has persisted for 7 yr. Alternatively, high
loads on edge and margin plants may reflect resource
quality for herbivores, which may in turn be a function
of plant stress, plant vigor, or nutrient stoichiometry
(Price 1991, Elser and Urabe 1999). Margin and edge
plants do appear more vigorous than their central counterparts.
A third hypothesis for core–edge differences is that
faunal buildup is more advanced in core areas, and
supports higher trophic levels that suppress herbivores.
In fact, spatial variation in tritrophic interactions is
likely at Mount St. Helens. The abundance of generalist
predators and parasitoids of lepidopteran larvae is
much higher in the core than at the edge, as are vertebrate predators, especially birds (Fagan and Bishop
2000; J. Edwards, unpublished data). Subtle differences in faunal succession might also explain the
damped oscillation in fruit damage on core plants over
a 7-yr period (1990–1996), and the contrasting monotonic decrease at the edge (Fig. 3).
Density-dependent mortality and population growth
Several observations imply that high mortality in
core patches (Fig. 4, 1992–1994 cohorts) is attributable
to resource competition. Intraspecific density reaches
1600 plants/m2 in some quadrats, median intraspecific
neighbor distance was .7 cm only in 1991, and cover
of all species averages 54% (Table 1). Seedling mortality was much higher than that for established plants,
except in 1991 (Fig. 4), a hallmark of competition (Silvertown and Lovett Doust 1993). Finally, population
growth in core patches appears negatively density dependent (Fig. 5), even when year is entered as a covariate (Table 3), and despite the fact that herbivory is
concentrated in low-density margins. This suggests that
insect-caused mortality and propagule loss may be unimportant to core population dynamics. This suggestion
is confirmed by the effect of insect removal, which
increased seed production in core patches without affecting r (Fagan and Bishop 2000).
Whether plant populations are often limited by propagule loss to herbivores is a continuing controversy
(Crawley 1989, 1990, 1992, Louda 1989). In lupines
at Mount St. Helens, the demographic significance of
propagule loss varies spatially. High levels of seed predation in core patches may have little effect on l because competition predominates, but propagule production in edge patches is closely linked to l and rate
of spread. However, core patches may depend on seed
production, since recruitment following disturbances
caused by erosion, lahars, and large mammals depends
on the newly formed, but short-lived, seed bank (Bishop 1996). Long-lived seed banks are thought to dampen
population-level effects of seed predation in other lupine species, but these systems too are frequently disturbed and seed banks could become depleted by seed
predation (Breedlove and Ehrlich 1972, Dollinger et
Ecology, Vol. 83, No. 1
al. 1973, Harrison and Maron 1995, Maron and Simms
1997). In lupines at Mount St. Helens, competition and
herbivory govern dynamics in different portions of the
population and their relative importance changes over
time. Negative density-dependent population growth
also occurred in edge patches, where the median density is only 0.13 plants/m2 and total cover averaged
only 0.4%, even when year is included first as a covariate (Table 3). Resource competition is an unlikely
explanation for negative density dependence at the
edge, but positive density-dependent herbivory in edge
areas by Hystricophora may be partially responsible.
Community and ecosystem implications
Farrell (1991) presents a model of consumer-driven
succession, wherein the effect of consumers depends
on the model of succession (e.g., tolerance, inhibition,
or facilitation). According to this view, consumers of
pioneer species will increase rates of succession when
pioneers inhibit later successional species, but consumers will decrease rates or alter trajectories of succession
when early species facilitate later ones. Vertebrate herbivores, for example, can severely delay succession in
N-limited grasslands by feeding on a legume that conrols the size of N pools (Ritchie et al. 1998). In contrast,
effects of herbivory on dune primary succession (Bach
1994) and marine primary succession (e.g., Hixon and
Brostoff 1996) are mediated by interactions that fall
under tolerance or inhibition models of succession. At
Mount St. Helens, episodic lupine mortality may increase local rates of succession, because invading ruderals are inhibited by live lupines but facilitated by
dead lupines (Morris and Wood 1989). This effect is
similar to that of L. arboreus in coastal grasslands,
where mortality caused by hepialid moth larvae mediates access by exotic plants to lupine-enriched N
pools (Maron and Jefferies 1999). However, at the landscape scale, delay in L. lepidus arrival caused by herbivores is likely to slow primary succession through
decreased rates of N mineralization, accumulation of
organic matter, and other facilitative effects (Halvorson
et al. 1991, Halvorson et al. 1992, del Moral 1993,
Titus and del Moral 1998b). Thus, at Mount St. Helens
insect herbivores may simultaneously prevent buildup
of N pools at large spatial scales and mediate access
to those pools at smaller scales.
ACKNOWLEDGMENTS
I thank D. Schemske and members of his lab for guidance,
discussion, and comments at all stages of this study, and D.
McCrumb, D. Kline, M. Jackson for extensive field assistance. L. Anderson, E. Anderson, P. Chilsen, B. Cook, I.
Dews, B. Fagan, A. Holzapfl, K. Koerner, B. Nakamura, I.
Parker, D. Parks, J. Ramsey, D. Schemske, J. Titus, K. Warner,
and R. Williamson, also provided invaluable field assistance.
C. Crisafulli kindly provided unpublished data. Comments
by K. Clay, B. Fagan, J. Kingsolver, J. Maron, J. Titus, and
three anonymous reviewers improved the manuscript. This
work was supported by the UW Plant Molecular Integration
and Function Committee, NSF dissertation improvement
grant DEB-9213143 and NSF training grant BIR-9256532.
January 2002
HERBIVORY, DEMOGRAPHY, AND COLONIZATION
LITERATURE CITED
Andow, D. A., P. K. Kareiva, S. A. Levin, and A. Okubo.
1990. Spread of invading organisms. Landscape Ecology
4:177–188.
Bach, C. E. 1994. Effects of a specialist herbivore (Altica
subplicata) on Salix cordata and sand dune succession.
Ecological Monographs 64:423–445.
Bishop, J. G. 1996. Demographic and genetic variation during colonization by the herb Lupinus lepidus on Mount St.
Helens. Dissertation. Department of Botany, University of
Washington, Seattle, Washington, USA.
Bishop, J. G., and D. W. Schemske. 1998. Variation in flowering phenology and its consequences for lupines colonizing Mount St. Helens. Ecology 79:534–546.
Braatne, J. H., and L. C. Bliss. 1999. Comparative physiological ecology of lupines colonizing early successional
habitats on Mount St. Helens. Ecology 80:891–907.
Breedlove, D. E., and P. R. Ehrlich. 1972. Coevolution: patterns of legume predation by a lycaenid butterfly. Oecologia
10:99–104.
Brown, V. K., A. C. Gagne, I. M. Evans, and A. L. Storr.
1987. The effect of insect herbivory on the growth and
reproduction of two annual Vicia species at different stages
in plant succession. Journal of Ecology 75:1173–1189.
Brown, V. K., M. Jepsen, and C. W. D. Gibson. 1988. Insect
herbivory: effects on early old field succession demonstrated by chemical exclusion methods. Oikos 52:293–302.
Clements, F. E. 1916. Plant succession: an analysis of the
development of vegetation. Carnegie Institution of Washington Publication, Washington, D.C., USA.
Connell, J. H., and R. O. Slatyer. 1977. Mechanisms of succession in natural communities and their role in community
stability and organization. American Naturalist 111:1119–
1144.
Crawley, M. J. 1989. The relative importance of vertebrate
and invertebrate herbivores in plant population dynamics.
Pages 45–71 in E. A. Bernays, editor. Insect–plant interactions. CRC Press, Boca Raton, Florida, USA.
Crawley, M. J. 1990. The population biology of plants. Philosphical Transactions of the Royal Society of London
B330:125–140.
Crawley, M. J. 1992. Seed predators and plant population
dynamics. Pages 157–191 in M. Fenner, editor. Seeds: the
ecology of regeneration in plant communities. C.A.B. International, London, UK.
del Moral, R. 1993. Mechanisms of primary succession on
volcanoes: a view from Mount St. Helens. Pages 79–100
in J. Miles and D. Walton, editors. Primary succession on
land. Blackwell Scientific, London, UK.
del Moral, R., and L. C. Bliss. 1993. Mechanisms of primary
succession: insights resulting from the eruption of Mount
St. Helens. Advances in Ecological Research 24:1–66.
del Moral, R., J. H. Titus, and A. M. Cook. 1995. Early
primary succession on Mount St. Helens, Washington,
USA. Journal of Vegetation Science 6:107–120.
del Moral, R., and D. M. Wood. 1988. Dynamics of herbaceous vegetation recovery on Mount St. Helens, Washington, USA, after a volcanic eruption. Vegetatio 74:11–27.
Dollinger, P. M., P. R. Ehrlich, W. L. Fitch, and D. E. Breedlove. 1973. Alkaloid and predation patterns in Colorado
lupine populations. Oecologia 13:191–204.
Ehrlén, J. 1995. Demography of the perennial herb Lathyrus
vernus. II. Herbivory and population dynamics. Journal of
Ecology 83:297–308.
Elser, J. J., and J. Urabe. 1999. The stoichiometry of consumer-driven nutrient recycling: theory, observations, and
consequences. Ecology 80:735–751.
Fagan, W. F., and J. G. Bishop. 2000. Trophic interactions
during primary succession: herbivores slow a plant rein-
201
vasion at Mount St. Helens. American Naturalist 155:238–
251.
Farrell, F. M. 1991. Models and mechanisms of succession:
an example from a rocky intertidal community. Ecological
Monographs 61:95–113.
Franklin, J. F., and C. T. Dyrness. 1973. Natural vegetation
of Oregon and Washington. Oregon State University Press,
Corvallis, Oregon, USA.
Gange, A. C., V. K. Brown, I. M. Evans, and A. L. Storr.
1989. Variation in the impact of insect herbivory on Trifolium pratense through early plant succession. Journal of
Ecology 77:537–551.
Halvorson, J. J., E. H. Franz, J. L. Smith, and R. A. Black.
1992. Nitrogenase activity, nitrogen fixation, and nitrogen
inputs by lupines at Mount St. Helens. Ecology 73:87–98.
Halvorson, J. J., J. L. Smith, and E. H. Franz. 1991. Lupine
influence on soil carbon, nitrogen and microbial activity in
developing ecosystems at Mount St. Helens. Oecologia 87:
162–170.
Harrison, S., and J. L. Maron. 1995. Impacts of defoliation
by tussock moths (Orgyia vetusta) on the growth and reproduction of bush lupine (Lupinus arboreus). Ecological
Entomology 20:223–229.
Hixon, M. A., and W. N. Brostoff. 1996. Succession and
herbivory: effects of differential fish grazing on Hawaiian
coral-reef algae. Ecological Monographs 66:67–90.
Kalbfleisch, J. D., and R. L. Prentice. 1980. The statistical
analysis of failure time data. Wiley, New York, New York,
USA.
Kerle, E. A. 1985. The ecology of lupines in Crater lake
National Park. Thesis. Oregon State University, Corvallis,
Oregon, USA.
Kruckeberg, A. R. 1987. Plant life on Mount St. Helens before 1980. Pages 3–17 in D. Bilderback, editor. Mount St.
Helens 1980. University of California Press, Berkeley, California, USA.
Kunin, W. E. 1999. Patterns of herbivore incidence on experimental arrays and field populations of ragwort, Senecio
jacobea. Oikos 84:515–525.
Louda, S. M. 1989. Predation in the dynamics of seed regeneration. Pages 25–51 in M. A. Leck, V. T. Parker, and
R. L. Simpson, editors. Ecology of soil seed banks. Academic Press, New York, New York, USA.
Louda, S. M., and M. Potvin. 1995. Effect of inflorescencefeeding insects on the demography and lifetime fitness of
a native plant. Ecology 76:229–245.
Louda, S. M., and J. E. Rodman. 1996. Insect herbivory as
a major factor in the shade distribution of a native crucifer
(Cardamine cordifolia A. Gray, bittercress). Journal of
Ecology 84:228–237.
Lubchenco, J., and S. D. Gaines. 1981. A unified approach
to marine plant–herbivore interactions. I. Populations and
communities. Annual Review of Ecology and Systematics
12:405–437.
Manly, B. F. J. 1991. Randomization and Monte Carlo methods in biology. Chapman and Hall, London, UK.
Maron, J. L. 1997. Interspecific competition and insect herbivory reduce bush lupine (Lupinus arboreus) seedling survival. Oecologia 110:284–290.
Maron, J. L., and R. L. Jefferies. 1999. Bush lupine mortality,
altered resource availability and alternative vegetation
states. Ecology 80:443–454.
Maron, J. L., and E. L. Simms. 1997. Effect of seed predation
on seed bank size and seedling recruitment of bush lupine
(Lupinus arboreus). Oecologia 111:76–83.
Marrs, R. H., R. D. Roberts, R. A. Skeffington, and A. D.
Bradshaw. 1983. Nitrogen and the development of ecosystems. Pages 113–136 in J. A. Lee, S. McNeill, and I.
H. Rorison, editors. Nitrogen as an ecological factor. The
202
JOHN G. BISHOP
22nd Symposium of the British Ecological Society. Blackwell Scientific, Oxford, UK.
Miles, J., and D. W. H. Walton. 1993. Primary succession
on land. Blackwell Scientific, London, UK.
Moody, M. E., and R. N. Mack. 1988. Controlling the spread
of plant invasions: the importance of nascent foci. Journal
of Applied Ecology 25:1009–1021.
Morris, W. F., and D. M. Wood. 1989. The role of Lupinus
lepidus in succession on Mount St. Helens: facilitation or
inhibition? Ecology 70:697–703.
Nuhn, W. W. 1987. Soil genesis on the 1980 pyroclastic flows
of Mount St. Helens. Thesis. University of Washington,
Seattle, Washington, USA.
Price, P. W. 1991. The plant vigor hypothesis and herbivore
attack. Oikos 62:244–251.
Reynolds, G. D., and L. C. Bliss. 1986. Microenvironmental
investigations of tephra covered surfaces at Mount St. Helens. Pages 147–152 in S. Keller, editor. Mount St. Helens:
five years later. Eastern Washington Press, Cheney, Washington, USA.
Ritchie, M. E., D. Tilman, and J. M. H. Knops. 1998. Herbivore effects on plant and nitrogen dynamics in oak savanna. Ecology 79:165–177.
Root, R. B. 1996. Herbivore pressure on goldenrods (Solidago altissima): its variation and cumulative effects. Ecology 77:1074–1087.
Silvertown, J. W., and J. Lovett Doust. 1993. Introduction
to plant population biology. Blackwell Scientific, Oxford,
UK.
Skellam, J. G. 1951. Random dispersal in theoretical populations. Biometrika 38:196–218.
Sousa, W. P., and J. H. Connell. 1992. Grazing and succession
in marine algae. Pages 425–441 in D. M. John, S. H. Hawkens, and J. H. Price, editors. Plant–animal interactions in
the marine benthos. Clarendon, Oxford, UK.
Statsci. 1994. S-Plus 3.3. Statsci, Seattle, Washington, USA.
Ecology, Vol. 83, No. 1
Strong, D. R., J. L. Maron, P. G. Connors, A. Whipple, S.
Harrison, and R. L. Jefferies. 1995. High mortality, fluctuation in numbers, and heavy subterranean insect herbivory in bush lupine, Lupinus arboreus. Oecologia 104:85–
92.
Titus, J. H., and R. del Moral. 1998a. The role of mycorrhizal
fungi and microsites in primary succession on Mount St.
Helens. American Journal of Botany 85:370–375.
Titus, J. H., and R. del Moral. 1998b. Seedling establishment
in different microsites on Mount St. Helens, Washington,
USA. Plant Ecology 134:13–26.
Vitousek, P. M., and L. R. Walker. 1989. Biological invasion
by Myrica faya in Hawaii: plant demography, nitrogen fixation, ecosystem effects. Ecological Monographs 59:247–
266.
Vitousek, P. M., L. R. Walker, L. D. Whiteaker, D. MuellerDombois, and P. A. Matson. 1987. Biological invasion by
Myrica faya alters ecosystem development in Hawaii. Science 238:802–804.
Walker, L. R., and F. S. I. Chapin. 1987. Interactions among
processes controlling successional change. Oikos 50:131–
135.
Wood, D. M., and M. C. Anderson. 1990. The effect of predispersal seed predators on colonization of Aster ledophyllus on Mount St. Helens. American Midland Naturalist
123:193–201.
Wood, D. M., and R. del Moral. 1987. Mechanisms of early
primary succession in subalpine habitats on Mount St. Helens. Ecology 68:780–790.
Wood, D. M., and R. del Moral. 1988. Colonizing plants on
the Pumice Plains, Mount St. Helens, Washington. American Journal of Botany 75:1228–1237.
Wood D. M., and R. del Moral. 2000. Seed rain during primary succession on Mount St. Helens, Washington. Madroño 47:1–9.
Zar, J. H. 1984. Biostatistical analysis. Prentice-Hall, Englewood Cliffs, New Jersey, USA.
APPENDIX A
Failure time analysis of lupine survival is available in ESA’s Electronic Data Archive: Ecological Archives E083-004-A1.
APPENDIX B
Photographs showing the impact of insect herbivores are available in ESA’s Electronic Data Archive: Ecological Archives
E083-004-A2.