influence of woody invader control methods and seed availability on native and invasive species
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ORIGINAL PAPER
Influence of woody invader control methods and seedavailability on native and invasive species establishmentin a Hawaiian forest
Rhonda K. Loh Æ Curtis C. Daehler
Received: 5 May 2007 / Accepted: 20 November 2007 / Published online: 26 March 2008
� Springer Science+Business Media B.V. 2008
Abstract When invasive woody plants become
dominant, they present an extreme challenge for
restoration of native plant communities. Invasive
Morella faya (fire tree) forms extensive, nearly
monospecific stands in wet and mesic forests on the
Island of Hawai’i. We used logging, girdling, and
selective girdling over time (incremental girdling) to
kill stands of M. faya at different rates, with the
objective of identifying a method that best promotes
native forest re-establishment. We hypothesized that
rapid canopy opening by logging would lead to
establishment of fast-growing, non-native invaders,
but that slower death of M. faya by girdling or
incremental girdling would increase the establish-
ment by native plants adapted to partial shade
conditions. After applying the M. faya treatments,
seed banks, seed rain, and plant recruitment were
monitored over 3 years. Different plant communities
developed in response to the treatments. Increased
light and nitrogen availability in the logged treatment
were associated with invasion by non-native species.
Native species, including the dominant native forest
tree, (Metrosideros polymorpha) and tree fern (Cibo-
tium glaucum), established most frequently in the
girdle and incremental girdle treatments, but short-
lived non-native species were more abundant than
native species. A diverse native forest is unlikely to
develop following any of the treatments due to seed
limitation for many native species, but girdling and
incremental girdling promoted natural establishment
of major components of native Hawaiian forest.
Girdling may be an effective general strategy for
reestablishing native vegetation in areas dominated
by woody plant invaders.
Keywords Girdling � Morella faya �Myrica faya � Resource availability �Restoration � Seed bank � Seed rain
Introduction
Restoring plant communities that have been highly
altered by invasion presents a daunting challenge to
conservation managers. Some successful alien plant
invaders can completely displace native vegetation
and alter ecosystem functions (Vitousek et al. 1997).
Ultimately, rates and trajectories of natural succes-
sion may be altered either through continued
persistence of the invaders (Brown et al. 2006; Titus
and Tsuyuzaki 2003) or by their eventual replace-
ment by communities that differ substantially from
pre-invasion communities (Walker and Smith 1996).
R. K. Loh (&)
Division of Resources Management, Hawai’i Volcanoes
National Park (HVNP), National Park Service,
P.O. Box 52, Hawaii National Park, HI 96718, USA
e-mail: rhonda_loh@nps.gov
C. C. Daehler
Department of Botany, University of Hawai’i at Manoa,
3190 Maile Way, Honolulu, HI 96822, USA
123
Biol Invasions (2008) 10:805–819
DOI 10.1007/s10530-008-9237-y
Neither of these outcomes is desirable in areas
established as reserves for native ecosystems.
Invasions by woody species have attracted special
attention (e.g., Reichard and Hamilon 1997; Krivanek
and Pysek 2006) because of their potential to alter the
character of vegetation across extensive areas (Lons-
dale 1993; Serbesoff-King 2003) or in highly
threatened ecosystems (Zalba and Villamil 2002;
Tassin et al. 2006). The nitrogen fixing tree Morella
faya (Aiton) Wilbur (formerly Myrica faya) is one
example of a woody invader capable of dominating
native plant communities and disrupting ecosystems in
Hawai’i. Prolific seed production and widespread
dispersal by birds contribute to its rapid spread and
establishment (Smathers and Gardner 1979; LaRosa
et al. 1985; Vitousek and Walker 1989). On young
volcanic substrates, invasive M. faya stands increase
nitrogen inputs up to four-fold above natural back-
ground levels (Vitousek et al. 1987). Once established,
leaf litter may limit recruitment by other species
(Walker and Vitousek 1991). In Hawai’i Volcanoes
National Park (HAVO), M. faya occupies[12,000 ha,
replacing native forest dominated by Metrosideros
polymorpha Gaud (Whiteaker and Gardner 1985).
Recovery of native forest after cutting down
M. faya stands is expected to be low because light
levels and the legacy of elevated nitrogen in the soils
are predicted to favor establishment of fast-growing
alien species (Vitousek and Walker 1989; Adler et al.
1998; Pattison et al. 1998; Durand and Goldstein
2001). In addition, tropical forest recovery is often
limited by low seed dispersal from nearby forest
remnants and limited seed persistence in the soil seed
bank (Young et al. 1987; Holl 1999; Holl et al.
2000). With the exception of scattered M. polymor-
pha, there are few native species remaining where
M. faya has invaded in HAVO. Generally, the native
soil seed bank is not persistent in Hawaiian mesic
forests (Drake 1998). As a result, non-native species
are better poised to take advantage of large forest
gaps created by the removal of an invasive tree.
Similar situations are likely to occur in other parts of
the world where woody invaders have developed
extensive stands (e.g., Holmes et al. 2000; Weber
2003). In such cases, knowledge from research on
forest gap dynamics might be applied to improve
prognoses for restoration.
Following death of a patch of forest trees, the
environment of the forest gap and subsequent seed
rain and plant establishment may depend on how the
forest gap is created (Hopkins et al. 1990; Martinez-
Ramos and Soto-Castro 1993). Immediately follow-
ing gap creation, light availability and diurnal ranges
in air and soil temperatures increase. The degree to
which these changes occur and the resulting plant
species that establish depend on the size and severity
of the disturbance (Bazzaz and Pickett 1980; Brokaw
1985; Cabin 2002a). Small or partial gaps (e.g.,
individuals tree fall, selective logging) are character-
ized by relatively small increases in light availability
that facilitate the growth of shade-tolerant species
that pre-exist in the understory or colonize soon after
disturbance (Burton and Mueller-Dombois 1984;
Denslow 1987; Guariguata 2000). In contrast, large
gaps (e.g., from clear-cutting) facilitate the establish-
ment of light-demanding species that have pre-
existing seeds in the soil or that arrive soon after
disturbance (Brokaw 1985). Consequently, there may
be little quantitative correlation between seed avail-
ability and the establishing vegetation, depending on
the nature of the disturbance (Putz and Appanah
1987).
The objective of this study was to determine
how different rates of killing a dominant woody
invader affect subsequent plant establishment by
native and invasive plants. Studies elsewhere have
shown that the method of invasive stand removal
can influence patterns of species establishment
(Wakibara and Mnaya 2002). In our experiment,
we selected three methods (logging, girdling, and
incremental girdling) to kill stands of M. faya at
different rates and thereby produce different gap
environments. We hypothesized that logging would
lead to establishment of fast-growing, weedy, non-
native plants due to sudden increases in resource
availability, which often differentially promote the
success of invasive plants in Hawaiian forests
(Durand and Goldstein 2001; Ostertag and Verville
2002). In contrast, we hypothesized that slower
death of M. faya by girdling or incremental girdling
would provide increased opportunities for estab-
lishment by native forest species, which are often
slower growing (Pattison et al.1998), but may be
able to tolerate lower resource conditions (Daehler
2003). We also hypothesized that plant communi-
ties establishing in treated areas farther from the
M. faya forest edge would be characterized by
lower species richness.
806 R. K. Loh, C. C. Daehler
123
Materials and methods
Study area
The study area was a 270-ha forest located on the
southeast flank of Kilauea Volcano in Hawai’i
Volcanoes National Park (HAVO) on the island of
Hawai’i (19�200 N, 155�150 E, 1,100–1,200 m eleva-
tion; 1,800 to 3,000 mm annual rainfall) (Doty and
Mueller-Dombois 1966). The soil is part of the
Puhimau Series and consists of shallow, well-drained
silt loams that formed from volcanic ash and pumice.
The vegetation was formerly native M. polymorpha
forest. Dicranopteris linearis (Burm.) Underw., a
native mat fern, had been a common understory
component along with scattered native trees, tree
ferns [Cibotium glaucum (Smith) H. and A.], shrubs,
and herbaceous species (Mueller-Dombois and Fos-
berg 1974). M. faya was first noted in the area in the
early 1960s (Kawasaki 1966). M. faya is now the
dominant species with M. polymorpha remaining as
an occasional emergent tree. The M. faya stands are
nearly devoid of understory vegetation (\1% abso-
lute cover). Sparse native sedges [Carex wahuensis
C.A. Mey., Uncinia uncinata (L. fil.) Kukenth.] and
occasional tree ferns (C. glaucum) persist. Leaf litter
(2–5-cm litter depth), primarily from M. faya,
blankets most of the forest floor. A fence erected in
1984 excludes feral pigs from the area.
Morella faya stand treatments
In summer 1999, five replicate 30 9 30-m plots of
M. faya forest were randomly assigned to one of three
M. faya stand treatments or an untreated control in
each of two sites. The three stand treatments were
selected to kill trees at different rates: (1) log
treatment: clear-cutting followed by stump applica-
tion of 50% Garlon 3A (Dow AgroSciences LLC,
Indianapolis, IN) with the cut material left in place,
(2) girdle treatment: bark girdling of all M. faya
followed by herbicide application to the wound, and
(3) incremental girdle treatment: bark girdling and
herbicide application to 33% of the M. faya trees,
followed by another 33% of trees at 12 months, and
girdling of the final 33% of trees after 20 months.
Dead trees, downed wood, and debris were left in the
plots. Any treated trees that appeared to be respro-
uting were treated a second time within 1 month. The
two sites were separated by *1 km. The interior site
was located *200 m from the M. faya forest edge,
whereas the edge site was located 10–50 m from the
M. faya forest edge in the same M. faya forest.
Measures of microenvironment
Soils were analyzed for pH, organic C, P, K, Ca, B,
Mg, and total N content before treatment and at the
end of the 3-year study. Three 10-cm deep soil cores
were collected from randomly selected sites in each
plot. Soils were passed through a 2-mm soil sieve to
remove small rocks and roots, and the three subsam-
ples from each plot were combined. Soils were sent to
the University of Hawaii at Manoa Agricultural
Diagnostic Service Center (Honolulu, HI) for extrac-
tion and analysis. Total N content was determined by
Kjeldahl digestion and colorimetric analysis (Shuman
et al. 1973; Nelson and Sommers 1972). Organic
carbon was determined by chromic acid digestion and
spectrophotometric analysis (Heanes 1984). Concen-
trations of P, K, Ca, Mg, and B were determined by
ashing and inductively coupled plasma spectroscopic
analysis (Isaac and Johnson 1983).
Nitrogen availability in the field, measured as
extractable NH�4 and NO�3 was determined at
1-month and at 3–4-month intervals following initial
treatments using the resin bag method developed by
Crews et al. (1995), and as detailed in Loh and
Daehler (2007). Nitrogen mineralization potential in
the soil was determined prior to treatments and 1.5
and 3 years following treatment using the protocols
developed by Binkley and Vitousek (1990). Three
10-cm-deep soil cores were collected from randomly
selected sites located in each plot. Soils were passed
through a 2-mm soil sieve to remove small rocks and
roots. Net mineralization (final [NO3-N + NH4-
N] - initial [NO3-N + NH4-N]) and nitrification
rates (Final [NO3-N] - Initial [NO3-N]) were deter-
mined for a 10-day incubation period. One set of 10-g
subsamples was incubated in the dark for 10 days at
room temperature prior to extraction with 2N KCL. A
second set of 10-g subsamples was extracted with 2N
KCl immediately after being sieved. After 20–24 h,
the supernatant was passed through filter paper
(Whatman #1, pre-rinsed with 90 ml of 2N KCl)
and sent to Stanford University for analysis
(P.M. Vitousek, Department of Biological Science,
Stanford University, Stanford, CA). Nitrate and
Invasive species establishment in a Hawaiian forest 807
123
ammonium concentrations were determined using an
Alpkem autoanalyzer (Alpkem Corporation, Wilson-
ville, OR).
Soil moisture was determined prior to the exper-
iment and at quarterly intervals by collecting 10-cm-
deep soil cores from five randomly selected locations
within each plot. Soils were passed through a 2-mm
soil sieve to remove small rocks and roots and then
weighed before and after drying at 100�C for 72 h to
determine percent soil moisture. Litter and downed
wood were collected annually from three 25 9
25-cm subplots randomly located in each sample
plot. Material was dried in an oven for 72 h, sorted
and weighed.
Relative humidity and air temperature 5 cm above
the soil surface were measured using HOBO H8 Pro
RH/Temperature loggers (Onset Computer Corp.,
Pocasset, MA), as detailed in Loh and Daehler
(2007). Monthly rainfall was measured from one rain
gauge placed at each site. Light availability (PAR)
was measured annually, following initial treatments,
at 10 cm and 80 cm above the forest floor using a
silicon photodiode LI-191SA line quantum sensor
(LICOR Biosciences, Lincoln, NE). The sensor was
held at each height in five randomly selected
locations in each plot. Measurements were taken at
1-s intervals for 1 min within 2 h of solar noon. Light
readings were expressed as a ratio relative to readings
taken simultaneously by a LI-190SB quantum sensor
placed in a nearby open area. Sensors were sent to
LICOR for calibration prior to sampling to ensure
readings were similar when exposed to the same
lighting.
Plant measurements
Prior to the start of the experiment, and in each of
3 years after application of the treatments, understory
vegetation (\2 m height) was surveyed in the
20 9 20-m area centered within each 30 9 30-m
M. faya treatment plot, and all species present were
recorded. Cover and density were quantified in
60-1 9 1-m subplots established along three parallel
20-m transects that spanned each 20 9 20-m sam-
pling plot. Percent cover of each vascular plant
species (\2 m height) was estimated using modified
Daubenmire cover classes, which were subsequently
converted to percentages using the midpoints of each
cover class (Mueller-Dombois and Ellenberg 1974).
Mosses, liverworts, and lichens were combined to
estimate total cover of non-vascular plants. Covers of
substrate components (downed wood, litter, soil, and
rock) were estimated for each subplot. Recruitment of
woody species and ferns was monitored and classified
into 10-cm height categories.
Seed availability
Seed availability, defined as the pool of propagules
available for plant establishment, was determined by
sampling the seed rain (input of seeds to an area) and
soil seed bank (seeds present in the soil) in the
20 9 20-m plots. Seed rain was collected in eight
randomly placed seed traps, which sampled a total
area of 0.12 m2. Traps were made from plastic pots
(14-cm diameter at the top and 10.5-cm deep) from
which the bottoms were removed and replaced with
polyester cloth secured by rubber bands (Drake
1998). Traps were placed directly on the ground
and wire mesh (2.4 9 3.4-cm aperture) was placed
over the tops of traps to prevent seed predation by
rats. Traps were collected at 2–3-month intervals.
Propagules were identified by comparison with
reference specimens collected from mature plants.
Unidentified seeds/spores were placed in petri dishes
containing water agar (15 g l-1), and subsequently
the germinants were transferred to soil trays where
they were grown and identified. For some species
with large numbers of seeds (Kyllinga brevifolia
Rottb., M. polymorpha, M. faya, R. rosifolius, Setaria
gracilis Kunth.), a sub-sample was counted from each
seed trap by spreading seeds onto a gridded petri dish
(14-cm diameter) and counting the seeds from at least
20% of the grid area.
To measure the soil seed bank, eight soil cores
(12.1-cm diameter and 5-cm deep, including overly-
ing litter) covering a total of 0.09 m2 were collected
at quarterly intervals from each treatment plot.
Within 24 h, litter was removed, then sub-samples
from each plot were combined and spread onto a
1-cm layer of 1:1:1 peat moss:perlite:steam-sterilized
soil (sterilization at 100 C for 3 h) in a 23 9 23-cm
tray. Trays were placed in an outdoor nursery located
in Volcano, HI (elevation approximately 1,050 m).
Four control trays of sterile soil were randomly
distributed among the samples to monitor for external
seed inputs. Seedlings that emerged over 6 months
were identified, counted, and removed from the trays.
808 R. K. Loh, C. C. Daehler
123
Individuals too small to identify were transplanted
and grown to a larger size for identification. Two
sedges, K. brevifolia (alien) and Cyperus polystachos
Rottb. (native), were treated as a single group
because seeds and seedlings of these two species
could not be consistently distinguished. Fern species
were grouped together for the same reason.
Data analysis
Except where noted, all statistical analyses were
conducted using SYSTAT (version 10, SPSS Inc.,
Chicago, IL). Differences in environmental variables,
vegetation response and composition of the soil seed
bank and seed rain among stand treatments were
analyzed by ANOVA with treatment and site as fixed
factors. A repeated measures ANOVA was used to
analyze changes over time. Data were transformed
(log-transformed for count data, arcsine transformed
for percent cover data) to fit assumptions of homo-
scedasticity. Tukey multiple comparisons were then
used to identify significant treatment responses. For
cover and density, species were grouped and ana-
lyzed by life form and origin (alien herbaceous =
AH, alien woody = AW, alien ferns = AF, native
herbaceous = NH, native woody = NW, native
ferns = NF). Densities of M. faya, M. polymorpha,
and C. glaucum were analyzed separately by species.
Understory vegetation in control plots was too sparse
to include in the analysis.
Monotonic multidimensional scaling was used to
visualize changes in microenvironment and resources
in response to the treatments. Multiple regression
analysis was used to determine relationships between
the abundance (percent absolute cover or density) of
major plant groupings (alien herbaceous, alien
woody, native herbaceous, native woody and native
fern species) and resource supply (% PAR, % soil
moisture, ammonium, and nitrate). In all cases,
stepwise backward regression was used to find the
best-fit model. Species identified in the soil seed bank
and seed rain were grouped and analyzed by major
plant groupings [alien herbaceous = AH, alien
woody = AW, native herbaceous = NH, native
woody = NW, ferns = FERN, and K. brevifoli-
a + C. polystachos (KYLCYP)], and repeated
measures ANOVA was used to identify effects of
treatment and site (fixed factors) on seed rain and
soil seed bank over 3 years.
Results
Microclimate and resources at edge versus
interior sites
Light availability after logging was generally greater
at the interior site where there were fewer
M. polymorpha trees. There were no important
differences in monthly rainfall between the two sites,
but soil moisture at the interior site (36–46%) was
consistently higher than at the edge site (28–36%).
Instantaneous NHþ4 and NO�3 concentrations at the
interior (average 228 and 905 lg ml-1 bag-1,
respectively) were higher than at the edge site (151
and 622 lg ml-1 bag-1, respectively), and in gen-
eral, nutrient concentrations were higher at the
interior site (Table 1).
Treatment effects on microclimate and resource
availability
Treatment responses were similar for plots at the
edge and interior sites (Fig. 1). Relative to control
plots, logged plots had higher PAR, higher temper-
atures (both mean and maximum), and lower
humidity (Fig. 1). Average daily temperature in the
log plots (19.9�C) was 3.8�C higher than in control
plots. Average relative humidity in the log plots was
83% vs. 95% in control plots, but soil moisture
averaged higher in the log plots (38%) relative to
controls (32%) (P = 0.032). This difference in soil
moisture occurred only in the 1st year, probably due
to lack of transpiration by M. faya in the log plots.
Average relative light availability in log plots,
Table 1 Soil characteristics in the experimental plots
Interior site Edge site
pH 5.2 ± 0.3 5.3 ± 0.2
Organic carbon 6.2 ± 1.8 4.4 ± 1.2
Total % N* 0.4 ± 0.1 0.3 ± 0.1
P (lg/g)* 54.2 ± 9.9 35.2 ± 11
K (lg/g) 28.2 ± 6.6 21.3 ± 10
Ca (lg/g)* 1464 ± 233 942 ± 289
Mg (lg/g)* 363 ± 53 211 ± 58
B (lg/g)* 0.7 ± 0.2 0.4 ± 0.1
Values were averaged across treatments and over time due to
lack of statistical differences. Asterisks indicates a significant
difference between sites (ANOVA, P \ 0.05)
Invasive species establishment in a Hawaiian forest 809
123
measured as percentage of PAR reaching the plots,
was 31–63% at 10 cm above forest floor and 45–73%
at 80 cm above forest floor, versus 1–3% in the
control plots at both heights. Full sun was never
achieved in the log plots because of shading from
scattered M. polymorpha trees.
The incremental girdle treatment in the 1st year
remained similar to controls (Fig. 1). At the same
time, the girdle plots experienced increases in light
and temperature, but they were much smaller than in
the log plots (Fig. 1). By the 2nd year, the incre-
mental girdle plots experienced increased light and
temperature, and these variables also continued to
increase in the girdle plots (Fig. 1). By the 3rd year,
all treatments had increased soil nitrification,
increased N-mineralization, and decreased litter.
These 3rd year changes were also observed in the
control, but to a lesser extent (Fig. 1).
Trends among treatments in instantaneous soil
nitrogen, as measured by the resin bag assays, were
similar between the interior (Fig. 2) and edge site
(data not shown). Soil available NO�3 increased after
9 months in the log plots (Fig. 2), but later (between
18 and 30 months), decreased relative to the other
treatments (Fig. 2), although the decrease was statis-
tically significant only in comparison with the girdle
treatment. Soil available NO�3 increased after
18 months in girdle, and 22 months in incremental
girdle plots relative to control and log plots, but were
statistically significant only between the girdle and
log plots. By month 35, both the log and girdled
treatment had significantly less available NO�3 than
the control and incremental girdle treatment (Fig. 2).
No statistical trends were apparent over time for
instantaneous NHþ4 nitrogen (data not shown). Other
soil nutrients and pH did not vary significantly from
pre-treatment measurements (Table 1).
During the 1st year following treatment, felled
trees created large amounts of downed wood in log
plots (averaging 1042 g-2), compared to control plots
(267 g-2), while the girdle and incremental girdle
plots had intermediate amounts of downed wood (396
and 355 g-2, respectively). By the 3rd year, the large
volume of downed wood in the log plots dropped to
425 g-2, whereas amounts in the other treatments and
the control remained relatively stable.
Species richness
Log plots initially had the highest total species
richness, averaging 13 and 15 species for the interior
and edge sites, respectively. Girdle plots had inter-
mediate species richness (4 and 10 species for interior
and edge sites, respectively), while incremental girdle
and controls tended to have the lowest species
richness (Fig. 3). The edge site often had greater
species richness of alien species, but not native
species (Fig. 3). The first native species to establish
were sedges (C. polystachos, C. wahuensis and
U. uncinata). After 3 years, native species’ richness
averaged greater in the girdle (9–10 spp.) and
incremental girdle (6–9 spp.) plots than in the log
MDS Axis 1 -10 -5 0 10
MD
S A
xis
2
-6
-4
-2
0
2
4
6
C1
C2
C3
C1C2
C3
IG1
IG2
IG3
IG1IG2
IG3
G1
G2
G3
G1
G2
G3
L1
L2
L3
L1
L2
L3
PAR
Temp
N-mineralizationNitrification
Humidity
LitterC = controlL = loggedG= girdleIG = incremental girdle
5
Fig. 1 Comparison of environmental parameters after differ-
ent Morella faya control treatments over time. Numbers
indicate years after treatment. Black lettering indicates forest
interior plots while grey lettering indicates forest edge plots.
Stress = 0.078, Proportion of variance (RSQ) = 0.97
Fig. 2 Extractable soil NO�3 in log (LOG), girdle (GIR),
incremental girdle (INCGIR), and control (CTRL) treatments
at the forest interior site over time. Letters indicate significant
differences between treatments, as specified on the figure
810 R. K. Loh, C. C. Daehler
123
plots (5–7 spp., Fig. 3). This was because of the
greater prevalence of Isachne distichophylla Munro
ex Hbd., Pipturus albidus (Hook. and Arnott) A.
Gray, M. polymorpha, and native fern species [e.g.,
C. glaucum, Dryopteris wallachiana (Spreng.) Hyl.]
in girdle and incremental girdle plots. By the 3rd
year, there were no differences in alien species
richness among treatments (Fig. 3). M. faya seedlings
were present in all treatments, but individuals
[10 cm height were seldom observed.
Species abundance
In the 1st year, alien plants established most rapidly
in the log plots, with 20% and 13% absolute cover in
the forest edge and interior sites, respectively
(Fig. 4). The initial alien cover was dominated by
sedges, grasses (S. gracilis, Paspalum conjugatum
Berg., Holcus lanatus L.) and other herbaceous
species (A. arvensis, Oxalis corniculata L., Anemone
hupehensis). Alien woody species (Buddleia asiatica,
Rubus argutus, Rubus ellipticus, R. rosifolius)
became increasingly evident in the 2nd and 3rd
years. By the 3rd year, dense shrub thickets and alien
grasses dominated the log plots. Native species cover
in the log plots was mainly sedges.
In contrast to the log plots, alien plant establish-
ments were much slower in the incremental girdle
and girdle plots. After the 1st year, absolute percent
plant cover was \1% and was composed mainly of
native sedges that were already present prior to
treatment. In the 2nd and 3rd years, cover increased
more rapidly in the girdle plots than in the incre-
mental girdle plots (Fig. 4). Unlike the log plots,
there was relatively little alien grass present in girdle
and incremental girdle treatments, with the exception
of Ehrharta stipoides, a shade-tolerant grass that
invades mesic forest in Hawai’i. The most abundant
alien species were K brevifolia, B. asiatica, and R.
rosifolius. Low numbers of M. faya establishment
(individuals [10 cm height) occurred in incremental
girdle (2–12 individuals 60 m-2) and girdle (4–6
individuals 60 m-2) plots. The overall cover of
native species was relatively low, averaging \5%
cover; however, there was more recruitment of
native trees (Ilex anomala Hook. and Arnott,
Fig. 3 Mean species
richness (±1 SE) at 1, 2 and
3 years following treatment
following Morella fayastand removal in log (LOG),
girdle (GIR), incremental
girdle (INCGIR) and
control (CTRL) treatments.
Means that share the same
letter do not differ
significantly between
treatments (Tukey multiple
comparison test performed
on significant treatments
responses across both sites)
Invasive species establishment in a Hawaiian forest 811
123
M. polymorpha, P. albidus) (1–5 individu-
als 60 m-2) and ferns [C. glaucum, D. wallachiana,
Sadleria spp., Ctenitis rubiginosa (Brack.) Co-
pel.] [10 cm in height in the girdle (3–7
individuals 60 m-2) and incremental girdle (1–
6 individuals 60 m-2) plots than in the log plots
(0–1 individuals 60 m-2). Native and alien plant
understory cover in the untreated control plots
remained low (\1%) throughout the study.
Relationship between resource availability
and plant establishment
Multiple regressions showed that abundance (abso-
lute percent cover, density) of the several plant
categories (alien herbaceous, alien woody, native
herbaceous, native woody, native ferns) at 3 years
was most strongly influenced by PAR (Table 2).
Among alien herbaceous species, there was a strong
relationship between PAR (positive), NHþ4 soil con-
centrations (positive), and soil moisture (negative).
Together these factors accounted for 55% of the
observed variation in alien herbaceous cover among
plots and sites. When sites were analyzed separately,
soil moisture was not a significant factor. At the
forest interior site, light was the single most impor-
tant resource, accounting for 37.6% of the variation
in alien herbaceous cover. At the forest edge, light
and NHþ4 concentrations were important, accounting
for 76.4% of the variation in alien herbaceous cover.
Light availability was also positively associated with
alien woody (R2 = 30.6, P \ 0.001) and native
herbaceous (R2 = 30.5, P \ 0.001) cover. Positive
associations between light and native fern (R2 = 22.8,
P = 0.035) and native woody (R2 = 19.6, P = 0.018)
densities were weaker (Table 2).
Seed rain
In the 1st year, only nine alien species and two native
species were identified in the seed rain. Alien species
richness in the seed rain increased during the 2nd
year as fast-growing alien species from the seed bank
matured and contributed to the seed rain (Fig. 5).
Seed rain for alien species tended to be greater at the
forest edge site (Fig. 5). Native M. polymorpha and
alien M. faya, the two dominant tree species in the
study area, made up over 90% of the seed rain.
M. polymorpha seed rain (800–29,000 seeds
m-2 year-1) did not differ among treatments. During
Fig. 4 Absolute percent
cover (±1 SE) of native
(left) and alien (right)
vegetation \2 m height in
the understory at 1, 2, and
3 years following treatment
in log (LOG), girdle (GIR),
incremental girdle
(INCGIR), and control
(CTRL) treatments. Note
that the y-axis scale differs
for native and alien species.
Means that share the same
letter do not differ
significantly between
treatments (Tukey multiple
comparison test). The
control treatment was not
included in the statistical
comparison, but is shown
for illustrative purposes
812 R. K. Loh, C. C. Daehler
123
1st two years, M. faya was more abundant in control,
incremental girdle, and girdle plots (7,600–
23,000 seeds m-2 year-1) than in log plots (3,300–
8,400 seeds m-2 year-1). By the 3rd year, M. faya
seed rain greatly decreased in girdle and incremental
girdle plots (550–930 m-2 year-1). In the log plots,
Table 2 Multiple regression analysis for vegetation (cover or density) by native and alien life forms in the 3rd year
Independent variable Forest interior Forest edge Forest Interior & edge
Co-efficient T P Co-efficient T P Co-efficient T P
Alien herbs, sedges, grasses
%PAR 35.7 3.5 0.002 40.0 5.66 0.001 37.01 5.77 0.001
% soil moisture -1.06 5.05 0.001
Soil NHþ4 0.49 5.04 0.001 0.30 12.35 0.001
Alien woody species
%PAR 44.0 4.43 0.001 21.31 1.80 0.09 31.5 4.05 0.001
Native herbs, sedges, grasses
%PAR 3.11 3.52 0.001 3.39 3.24 0.005 3.13 18.1 0.001
Native woody species density
%PAR 2.613 1.762 0.096 1.976 2.486 0.018
Soil NO�3 0.006 2.211 0.041 0.005 3.107 0.004
Soil NHþ4 0.008 1.847 0.081
Native fern species density
%PAR 4.361 2.259 0.037 2.782 2.196 0.035
Only the step-wise best fit model is show for each category of plants. Missing values indicate variables that were dropped from the
model (P [ 0.15). Variables potentially included in each model were %PAR (photosynthetically active radiation), % soil moisture,
soil NO�3 ; and soil NHþ4
Fig. 5 Mean number of
species found in seed rain 1,
2, 3 years following
removal of Morella faya in
log (LOG), girdle (GIR),
incremental girdle
(INCGIR), and control
(CTRL) treatments in the
forest interior and forest
edge sites. Results of Tukey
multiple comparison tests
are displayed on forest
interior graphs (left). Means
that share the same letter do
not differ significantly
between treatment
Invasive species establishment in a Hawaiian forest 813
123
the seed rain of other alien species (2,400–
33,000 seeds m-2 year-1 in aggregate) mainly con-
sisted of A. arvensis, H. lanatus, Melinis minutiflora
P. Beauv., Oxalis corniculata L., R. rosifolius,
Sacciolepis indica (L.) Chase, Schizachyrium con-
densatum (Kunth) Nees, Setaria gracilis, and
Youngia japonica L. In girdle, incremental girdle
and control plots, seed rain of alien species, other
than M. faya, remained relatively low throughout the
experiment (8–120 m-2 year-1). Seed rain of native
species other than M. polymorpha remained relatively
low across all plots and sites (\400 m-2 year-1).
Seed bank
Seeds of 23 alien and 8 native species were detected
in the seed bank the 1st year (Fig. 6). Unlike in the
seed rain, M. faya (0–190 seeds m-2) and M. poly-
morpha (19–420 seeds m-2) did not dominate the
seed bank. Instead, K. brevifolia and C. polystachos,
R. rosifolius, S. gracilis, and other alien species not
initially found in the seed rain (Anemone hupehensis,
B. asiatica, M. minutiflora, S. indica, Sonchus spp.,
and Youngia japonica) were the major contributors to
the soil seed bank. Likely propagule sources were
from individuals that occurred outside of treatment
sites in small forest gap openings and beyond the
forest edge. Alien seed abundance and diversity were
greater at the forest edge site (6–9 species, 1,400–
2,000 seeds m-2) than at the forest interior
(6–7 species, 490–630 seeds m-2), but no differences
were detected among treatments. Native species that
were not initially found in the seed rain were detected
in very small quantities in the soil seed bank (\1
seed m-2), including three sedges (M. angustifolia,
C. polystachos, C. wahuensis), a small herb (Luzula
hawaiiensis Buchenau), and two shrubs (Dodonaea
viscosa Jacq., Leptecophylla tameiameiae (Cham. &
Shlechtend.) F. v. Muell.]. Spores of ferns
(C. glaucum, and alien Pityrogramma spp.) were
widespread in the soil in all treatments and sites
(50–200 germinants m-2). Native species, excluding
C. polystachos and ferns, made up \1% of the
propagules in the soil seed bank.
Discussion
Effect of stand treatments on plant establishment
As predicted, the large gaps created by logging led to
rapid establishment of non-native species. The log
Fig. 6 Mean number of
species found in the seed
bank 1, 2, 3 years following
removal of Morella faya in
log (LOG), girdle (GIR),
incremental girdle (INC
GIR), and control (CTRL)
treatments in the forest
interior and forest edge
sites. There were no
statistical differences
among treatments
814 R. K. Loh, C. C. Daehler
123
plots were characterized by high-light conditions, an
initial increase in available nitrogen, and wide
temperature fluctuations, all of which may have
directly promoted the germination and establishment
of fast-growing herbaceous species present in the soil
seed bank and seed rain (Pons 1992; Probert 1992).
Leaf litter, while initially high in all the treatment
plots, declined most rapidly in log and girdle plots.
Consequently, germination of small seeds from the
seed bank (mainly non-native) was probably also less
constrained by the physical barrier of the litter in the
latter treatments (Molofsky and Augspurger 1992).
Other studies have also documented rapid establish-
ment and dominance by non-native weedy species
associated with clear-cutting of non-native woody
species in plant communities ranging from temperate
forest (Webb et al. 2001) to the South African fynbos
(Holme et al. 2000). In contrast to the log plots,
partial shade environments, combined with leaf litter,
may have limited expression of the weedy plant seed
bank in incremental girdle and girdle plots (Scow-
croft 1992). The result was slower rates of
establishment by light-demanding non-native herbs
and selective establishment of native species as well
as some aliens that could tolerate partial shade (e.g.,
B. asiatica, R. argutus, and R. rosifolius, E. stipoides,
K. brevifolia, and A. hupehensis).
The partial shade environments created by girdle
and incremental girdle treatments supported the most
diverse native species assemblage that included
M. polymorpha and C. glaucum, the two major
canopy species in nearby native rain forest (Mueller-
Dombois and Fosberg 1974; Lipp 1994). Light levels
at the forest floor in the girdle and incremental girdle
plots were near optimal for M. polymorpha germina-
tion (4–15% relative irradiance, cf. Burton 1982;
Burton and Mueller-Dombois 1984). Likewise,
C. glaucum establishment is limited to moist, shady
areas (Becker 1976). In girdle and incremental girdle
plots, fern establishment was largely confined to
moss beds and nurse stumps that emerged from the
litter. These microenvironments became more evi-
dent after the 2nd and 3rd year in the girdle and
incremental girdle treatments when the thick litter
layer began decomposing. Because of their slow
growth, only small individuals of M. polymorpha and
C. glaucum were recorded in the 1st 3 years after
girdling. Given sufficient time, these native species
are expected to increase in size and dominance as
long as there is no inhibitory dominance by alien
species, such as E. stipoides, in the plots (McDaniel
2007). M. polymorpha and C. glaucum are long-lived,
and they may outlast or displace the shorter-lived
alien species, such as B. asiatica, R. rosifolius and
R. argutus, which characteristically decline or die
back following fruit production (Engard 1945;
Wright and Mueller-Dombois 1988).
Many native species in other forest ecosystems
may also respond positively to slow increases in light
associated with girdling an invader, relative to the
rapid light increases associated with clear-cutting.
For example, in an African tropical forest, girdling of
Senna spectabilis, an invasive tree, led to greater
establishment and cover by native forest species
compared to clear-cutting the invader (Wakibara and
Mnaya 2002). Likewise, many native fynbos trees in
riparian forest establish better beneath closed or
partial canopies, relative to in open areas (Gala-
towitsch and Richardson 2005).
In the log plots, one might expect a transition from
shade-intolerant, fast-growing herbs to taller, more
shade-tolerant woody species in the future (Brown
and Southwood 1987). Such a transition from non-
native sedges and grasses to non-native shrubs was
becoming apparent in log plots by the 3rd year of the
study, but few native species became established in
the log plots. Experiments in nearby sub-montane
forest at HAVO showed that native shrub recruitment
is limited by invasive perennial grasses (D’Antonio
et al. 1998), and this inhibitory effect on native
shrubs persists for at least 20 years, even in the
absence of fire (Hughes and Vitousek 1993). In other
cleared areas that have been left fallow, alien-
dominated thickets commonly form that are similar
to those in the log plots, but native M. polymorpha
forest has not regenerated, even after more than
10 years (C. Daehler, personal observation).
Clear-cutting, which resulted in alien shrub
(30–53% absolute cover) and grass (25–49% absolute
cover) dominance and very little native forest tree
and fern establishment (\1% absolute cover) does not
appear to be a feasible strategy for increasing the
abundance of native forest trees, unless the area is to
be managed as a garden, with constant weeding and
dense outplantings of native species. Clear-cutting of
invasive trees also led to poor regeneration by native
species in the South African fynbos (Galatowitsch
and Richardson 2005). In contrast, bull-dozing was
Invasive species establishment in a Hawaiian forest 815
123
an effective strategy for clearing an invasive grass
and facilitating restoration of a Hawaiian dry forest
(Cabin et al. 2002b). In this case, ‘‘clear-cutting’’
with a bulldozer effectively removed the invader
along with the upper substrate and much of the alien
seed bank. Outplanted native species could then be
established with the assistance of artificial watering,
at a time when relatively few weed seeds were being
dispersed from surrounding non-native vegetation.
The latter example of successful restoration initiated
by deliberate ‘‘clear cutting’’ is rather unusual; more
often, clear-cutting does not seem to be a practical
method of restoring native forests (e.g., Viisteensaari
et al. 2000; Webb et al. 2001; Loh and Daehler
2007), and it is often more costly than girdling.
Seed limitations on plant establishment
As predicted, the M. faya forest edge site received a
higher diversity and abundance of seeds relative to
the forest interior site, but this effect was due
primarily to an increase in the non-native component
of the seed rain at the forest edge site. Native species
associated with pioneer communities and forest gaps,
other than C. polystachos, seemed to be limited by
seed availability. For example, Pipturus albidus and
Isachne distichophylla are endemic species that
readily colonize following native forest disturbance,
but were rarely detected in the soil seed bank or seed
rain. Individuals grow in the adjacent uninvaded
forest and only occasionally found in small light gaps
near treatment sites. The few P. albidus and
I. disticthophylla plants that established appeared
near the end of the 3-year study. Species with fleshy
fruits (e.g., P. albidus, L. tameiameiae, Vaccinium
spp.) are apt to be dispersed by birds (Medeiros
2004). Native frugivorous birds were rare in the area,
but the introduced Japanese white-eye (Zosterops
japonica) was common, and this bird is known to
disperse seeds in Hawaiian forests (LaRosa et al.
1985). Introduced ground-foraging birds (e.g., Kalij
pheasant, Locura leucomelana) and rats (Rattus
rattus) could also serve as seed dispersers. However,
the overwhelming abundance of palatable fruit in the
vicinity of the treatment plots was supplied by
M. faya and R. rosifolius (in gaps).
Besides the two dominant native forest species (M.
polymorpha and C. glaucum), native species were
rare in the seed rain largely due to the absence of
nearby propagule sources. Other native species were
successfully established in the area though seed
additions and outplanting, demonstrating dispersal
limitation for some components of the native forest
(Loh and Daehler 2007). Selective losses of native
species in the seedbank and/or seed rain are common
in areas that have been highly invaded by non-native
species (e.g., Holmes and Cowling 1997; Holmes
2002; Wearne and Morgan 2006; Mason et al. 2007).
These losses have the potential to change the
character of regenerating vegetation. For example,
Holmes (2002) found that the seed bank in sand plain
fynbos that had been densely invaded by Acacia
saligna was missing many woody natives and would
likely develop into a herbland, rather than the original
native shubland if A. saligna were removed. In our
study, the two dominant native forest species
(M. polymorpha and C. glaucum) remained abundant
in the seed rain and appear capable of establishing
after M. faya has been girdled. Establishment by
these two species can provide the structural frame-
work of a native Hawaiian forest, but much of the
endemic diversity would likely remain missing
without seed supplementation and/or outplanting.
Reinvasion by Morella faya?
Although M. faya was dominant prior to treatment,
it was unable to dominate the early stages of forest
recovery. In log plots, re-establishment of M. faya
was hampered by a limited supply of propagules. In
dry storage, 30% of seeds retain viability for up to
78 weeks (Walker 1990), but in field, viable seeds
may have been quickly depleted through germina-
tion, decay, or predation by rodents and birds
(Medeiros 2004). Consequently, re-establishment
was largely dependent on the seed rain from
individuals located outside of plots, and establish-
ment was extremely low (1 individual 30 m-2 in
log plots). Inability of M. faya to colonize open
sites has also been observed in other studies
(Vitousek and Walker 1987; Adler et al. 1998).
Nitrogen fixers commonly lose their competitive
edge after they have sufficiently altered their
environment to facilitate the growth of more
nitrophilous non-native or native species (Walker
1993; Adler et al. 1998).
In the partial shade environments of girdle and
incremental girdle plots, abundant litter and
816 R. K. Loh, C. C. Daehler
123
insufficient light may have contributed to poor
seedling establishment by M. faya (Vitousek and
Walker 1987; Walker 1990; Lipp 1994). However by
the 3rd year, as more light reached the forest floor and
leaf litter decomposed, a growing number of taller
M. faya seedlings (up to 1 individuals 5 m-2) were
evident in girdle and incremental girdle plots,
suggesting the potential for M. faya to regain
dominance in those plots in the absence of some
follow-up control.
Conclusions
An increase in resource supply, combined with the
presence of non-native propagules, can make com-
munities more vulnerable to invasion by alien species
following disturbance (Johnstone 1986; Hobbs and
Huennecke 1992; Davis et al. 2000). Relative to
instantaneous removal of M. faya by logging, slower
rates of killing M. faya stands (by girdling) decreased
rates of invasion by alien plants and allowed the
dominant native forest species to establish, but a
limited native propagule supply probably limited the
potential for establishment of a diverse native forest.
A likely outcome is that a partial restoration of native
forest will occur in the girdle and incremental girdle
plots. Seed supplementation and/or outplantings have
proven successful in establishing a wider range of
native species, especially in girdle and incremental
girdle treatments (Loh and Daehler 2007). A strategy
of re-establishing patches of native forest, then
girdling intervening M. faya, could increase effi-
ciency of the restoration process by providing natural
sources of native seeds. It is not clear whether re-
invasion by M. faya will be a problem, but there are
other alien species (Psidium cattleianum, Hedychium
gardnerianum and Rubus ellipticus) found \1 km
from the M. faya forest that have the potential to
develop dense stands. Seeds of these species were not
detected in the seed rain or seed bank, but they are all
bird-dispersed, and their early removal upon arrival at
restoration sites should be a high priority. Although
we can manipulate resource supply rates to favor the
establishment of key native forest species and
minimize invasive species establishment, there are
still likely to be a few invaders that require direct
control if we are to succeed in restoring a native-
dominated forest.
Acknowledgements We thank Tim Tunison, Peter Urias,
Alison Ainsworth, and the staff of the Division of Resources
Management at Hawai’i Volcanoes National Park for the
tremendous support in the field and laboratory. Thanks to Peter
Vitousek, Doug Turner, and Heraldo Farrington of Stanford
University and the staff at the Agricultural Diagnostic Service
Center, College of Tropical Agriculture and Human Resources,
University of Hawai’i at Manoa for their generosity in
performing laboratory analysis, and providing equipment,
technical advice, and laboratory space.
References
Adler PB, D’Antonio CM, Tunison JT (1998) Understory
succession following a dieback of Myrica faya in Hawaii
Volcanoes National Park. Pac Sci 52:69–78
Bazzaz FA, Pickett STA (1980) The physiological ecology of
tropical succession: a comparative review. Annu Rev Ecol
Syst 11:287–310
Becker RE (1976) The phytosociological position of tree ferns
(Cibotium spp.) in the montane rainforest on the island of
Hawaii. Dissertation, University of Hawaii at Manoa
Binkley D, Vitousek PM (1990) Soil nutrient availability. In:
Pearcy RW, Ehleringer JR, Mooney HA, Rundel PW (eds)
Plant physiological ecology. Chapman and Hall, London,
pp 75–96
Brokaw NVL (1985) Gap-phase regeneration in a tropical
forest. Ecology 66:682–687
Brown VK, Southwood TRE (1987) Secondary succession:
patterns and strategies. In: Gray AJ, Crawley MJ, Edwards
PJ (eds) Colonization, succession, and stability. Blackwell
Scientific Publications, Oxford, pp 315–337
Brown KA, Scatena FN, Gurevitch J (2006) Effects of an inva-
sive tree on community structure and diversity in a tropical
forest in Puerto Rico. For Ecol Manage 226:145–152
Burton PJ (1982) The effect of temperature and light on Met-rosideros polymorpha seed germination. Pac Sci 36:
229–240
Burton PJ, Mueller-Dombois D (1984) Response of Metro-sideros polymorpha seedlings to experimental canopy
opening. Ecology 65:779–791
Cabin RJ, Weller SG, Lorence DH, Cordell S, Hadway LJ
(2002a) Effects of microsite, water, weeding, and direct
seeding on the regeneration of native and alien species
within a Hawaiian dry forest preserve. Biol Conserv
104:181–190
Cabin RJ, Weller SG, Lorence DH, Cordell S, Hadway LJ,
Montgomery R, Goo D, Urakami A (2002b) Effects of
light, alien grass, and native species additions on
Hawaiian dry forest restoration. Ecol Appl 12:1595–1610
Crews TE, Kitayama K, Fownes JH, Riley RH, Herbert DA,
Mueller-Dombois D, Vitousek PM (1995) Changes in soil
phosphorus fractions and ecosystem dynamics across a
long chronosequence in Hawaii. Ecology 76:1407–1424
Daehler CC (2003) Performance comparisons of co-ocurring
native and alien invasive plants: implications for conser-
vation and restoration. Annu Rev Ecol Syst 34:183–211
D’Antonio CM, Hughes RF, Mack M, Hitchcock D, Vitousek
PM (1998) The response of native species to removal of
Invasive species establishment in a Hawaiian forest 817
123
invasive exotic grasses in a seasonally dry Hawaiian
woodland. J Veg Sci 9:699–712
Davis MA, Grime JP, Thompson K (2000) Fluctuating
resources in plant communities: a general theory of in-
vasibility. J Ecol 88:528–534
Denslow J (1987) Tropical rainforest gaps and tree species
diversity. Annu Rev Ecol Syst 18:431–451
Doty MS, Mueller-Dombois D (1966) Atlas for bioecological
studies in Hawaii Volcanoes National Park. Miscellaneous
Publication 89, College of Tropical Agriculture, Univer-
sity of Hawaii, Honolulu
Drake DR (1998) Relationship among the seed rain, seed bank
and vegetation of a Hawaiian forest. J Veg Sci 9:103–112
Durand LZ, Goldstein G (2001) Photosynthesis, photoinhibi-
tion, and nitrogen use efficiency in native and invasive
tree ferns in Hawaii. Oecologia 126:345–354
Engard CJ (1945) Habit and growth of Rubus rosaefolius Smith
in Hawaii. Am J Bot 32:536–548
Galatowitsch S, Richardson DM (2005) Riparian scrub
recovery after clearing of invasive alien trees in headwater
streams of the Western Cape, South Africa. Biol Conserv
122:509–521
Guariguata MR (2000) Seed and seedling ecology of tree
species in neotropical secondary forests: management
implications. Ecol Appl 10:145–154
Heanes DL (1984) Determination of total organic-C in soils by
an improved chromic acid digestion and spectrophoto-
metric procedure. Commun Soil Sci Plant Anal 15:
1191–1213
Hobbs RJ, Huenneke LF (1992) Disturbance, diversity, and
invasion: Implications for conservation. Conserv Biol
6:324–337
Holl KD (1999) Factors limiting tropical rain forest regenera-
tion in abandoned pastures: seed rain, seed gremination,
microclimate and soil. Biotropica 31:229–242
Holl KD, Loik ME, Lin EHV, Samuels IA (2000) Tropical
montane forest restoration in Costa Rica: overcoming
barriers to dispersal and establishment. Rest Ecol 8:
339–349
Holmes PM (2002) Depth distribution and composition of
seed-banks in alien-invaded and uninvaded fynbos vege-
tation. Austral Ecol 27:110–120
Holmes PM, Cowling RM (1997) The effects on invasion by
Acacia saligna on the guild structure and regeneration
capabilities of South African fynbos shrublands. J Appl
Ecol 34:317–332
Holmes PM, Richardson DM, Van Wilgen BW, Gelderblom C
(2000) Recovery of South African fynbos vegetation fol-
lowing alien woody plant clearing and fire: implications
for restoration. Austral Ecol 25:631–639
Hopkins MS, Tracey JG, Graham AW (1990) The size and
composition of soil seed-banks in remnant patches of
three structural rainforest types in North Queensland. Aust
J Ecol 15:43–50
Hughes F, Vitousek PM (1993) Barriers to shrub reestablish-
ment following fire in the seasonal submontane zone of
Hawai’i. Oecologia 93:557–563
Isaac RA, Johnson WC (1983) High speed analysis of agri-
culture samples using inductively coupled plasma-atomic
emission spectroscopy. Spectrochim Acta 38B:277–282
Johnstone IM (1986) Plant invasion windows: a time-based
classification of invasion potential. Biol Rev 61:369–394
Kawasaki A (1966) Survey of Myrica faya in volcano area
Sept. 7–9, 1966. Typescript and map. State of Hawaii,
Department of Agriculture, Honolulu
Krivanek M, Pysek P (2006) Predicting invasions by woody
species in a temperate zone: a test of three risk assessment
schemes in the Czech Republic (Central Europe). Divers
Distrib 12:319–327
LaRosa AM Smith CW, Gardner DE (1985) Role of alien and
native birds in the dissemination of firetree (Myrica fayaAit.—Myricaceae) and associated plants in Hawaii. Pac
Sci 39:372–378
Lipp CC (1994) Ecophysiological and community-level con-
straints to the invasion of Myrica faya, an alien tree in
Hawaii Volcanoes National Park. Dissertation, University
of Hawaii at Manoa
Loh RK, Daehler CC (2007) Influence of invasive tree kill rates
on native and invasive plant establishment in a Hawaiian
Forest. Rest Ecol 15:199–211
Lonsdale WM (1993) Rates of spread of an invading species -
Mimosa pigra in northern Australia. J Ecol 81:513–521
Martinez-Ramos M, Soto-Castro A (1993) Seed rain and
advanced regeneration in a tropical rain forest. Vegetatio
107/108:299–318
Mason TJ, French K, Russell KG (2007) Moderate impacts of
plant invasion and management regimes in coastal hind
dune seed banks. Biol Conserv 134:428–439
McDaniel S (2007) Effects of light availability on the germi-
nation, growth and survival of alien grasses and native
trees and shrubs. MS Thesis, University of Hawaii, Hilo
Medeiros AC (2004) Phenology, reproductive potential, seed
dispersal and predation, and seedling establishment of
three invasive plant species in a Hawaiian rain forest.
Dissertation, University of Hawaii at Manoa
Molofsky J, Augspurger CK (1992) The effect of leaf litter on
early seedling establishment in a tropical forest. Ecology
73:68–77
Mueller-Dombois D, Ellenberg H (1974) Aims and methods of
vegetation ecology. Wiley, New York
Mueller-Dombois D, Fosberg FR (1974) Vegetation commu-
nities of Hawaii Volcanoes National Park. Cooperative
Park Studies Unit Technical Report No. 4. University of
Hawaii at Manoa, Honolulu
Nelson DW, Sommers LE (1972) A simple digestion procedure
for estimation of total nitrogen in soils and sediments. J
Environ Qual 1:423–425
Ostertag R, Verville JH (2002) Fertilization with nitrogen and
phosphorus increases abundance of non-native species in
Hawaiian montane forests. Plant Ecol 162:77–90
Pattison RR, Goldstein G, Ares A (1998) Growth, biomass
allocation and photosynthesis of invasive and native
Hawaiian rainforest species. Oecologia 117:449–459
Pons TL (1992) Seed response to light. In: Fenner M (ed)
Seeds: the ecology and regeneration in plant communities.
Cab International, Wallingford, UK pp 259–284
Probert RJ (1992) The role of temperature in germination
ecophysiology. In: Fenner M (ed) Seeds: the ecology and
regeneration in plant communities. CAB International,
Wallingford, UK, pp 285–327
818 R. K. Loh, C. C. Daehler
123
Putz FE, Appanah S (1987) Buried seeds, newly dispersed
seeds, and the dynamics of a lowland forest in Malaysia.
Biotropica 19:326–333
Reichard SH, Hamilton CW (1997) Predicting invasions of
woody plants introduced into North America. Conserv
Biol 11:193–203
Scowcroft PG (1992) Role of decaying logs and other organic
seedbeds in natural regeneration of Hawaiian forest spe-
cies in abandoned montane pasture. USDA Forest Service
Gen. Tech. Rep. PSW-129
Serbesoff-King K (2003) Melaleuca in Florida: a literature
review on the taxonomy, distribution, biology, ecology,
economic importance and control measures. J Aquat Plant
Manage 41:98–112
Shuman GE, Stanley MA, Kundsen D (1973) Automated total
nitrogen analysis of soil and plant materials. Soil Sci Soc
Am Proc 37:480–481
Smathers GA, Gardner DE (1979) Stand analysis of an
invading firetree (Myrica faya Aiton) population, Hawaii.
Pac Sci 33:239–255
Tassin J, Riviere JN, Cazanove M, Bruzzese E (2006) Ranking
of invasive woody plant species for management on
Reunion Island. Weed Res 46:388–403
Titus JH, Tsuyuzaki S (2003) Influence of a non-native inva-
sive tree on primary succession at Mt. Koma, Hokkaido,
Japan. Plant Ecol 169:307–315
Viisteensaari J, Johansson S, Kaarakka V, Luukkanen O (2000)
Is the alien tree species Maesopsis eminii Engl. (Rhamn-
aceae) a threat to tropical forest conservation in the East
Usambaras, Tanzania?. Environ Conserv 27:76–81
Vitousek PM, Walker LR (1987) Colonization, succession, and
resource availability: ecosystem level interactions. In:
Gray AJ, Crawley MJ, Edwards PJ (eds) Colonization,
succession, and stability. Blackwell Scientific Publica-
tions, Oxford, UK, pp 207–224
Vitousek PM, Walker LR (1989) Biological invasion by
Myrica faya in Hawaii: plant demography, nitrogen fixa-
tion, ecosystem effects. Ecol Monogr 59:247–265
Vitousek PM, Walker LR, Whiteaker LD, Mueller-Dombois D,
Matson PA (1987) Biological invasion by Myrica fayaalters ecosystem development in Hawaii USA. Science
238:802–804
Vitousek PM, D’Antonio CM, Loope LL, Rejmanek M,
Westbrooks R (1997) Introduced species: a significant
component of human-caused global change. NZ J Ecol
21:1–16
Wakibara JV, Mnaya BJ (2002) Possible control of Sennaspectabilis (Caesalpiniaceae), an invasive tree in Mahale
Mountains National Park, Tanzania. Oryx 36: 357–363
Walker LR (1990) Germination of an invading tree species
(Myrica faya) in Hawaii. Biotropica 22:140–145
Walker LR (1993) Nitrogen fixers and species replacements in
primary succession. In: Mills J, Walk DWH (eds) Primary
succession on land. Blackwell Scientific Publications,
Oxford, pp 249–273
Walker LR, Smith SD (1996) Impacts of invasive plants on
community and ecosystem properties. In: Luken JO
Theiret JW (Eds) Assessment and management of plant
invasions. Springer Verlag, New York, pp 69–86
Walker LR, Vitousek PM (1991) An invader alters germination
and growth of a native dominant tree in Hawaii. Ecology
72:1440–1455
Wearne LJ, Morgan JW (2006) Shrub invasion into subalpine
vegetation: implications for restoration of the native
ecosystem. Plant Ecol 183:361–376
Webb SL, Pendergast TH, Dwyer ME (2001) Response of
native and exotic maple seedling banks to removal of the
exotic, invasive Norway maple (Acer platanoides). J
Torrey Bot Soc 128:141–149
Weber E (2003) Invasive plant species of the world: a guide to
environmental weeds. CABI Publishing, Bristol
Whiteaker LD, Gardner DE (1985) The distribution of Myricafaya Ait. in the state of Hawaii. Cooperative Park Studies
Unit Technical Report No. 55. University of Hawaii at
Manoa, Honolulu
Wright RA, Mueller-Dombois D (1988) Relationships among
shrub population structure, species associations, seedling
root form and early volcanic succession, Hawaii. In:
Werger MJA, Van der Aart PJM, During HJ, Verhoeven
JTA (eds) Plant form and vegetation structure. SPB
Academic Publishing, The Hague, The Netherlands,
pp 87–104
Young KR, Ewel JJ, Brown BJ (1987) Seed dynamics during
forest succession in Costa Rica. Vegetation 71:157–174
Zalba SM, Villamil CB (2002) Woody plant invasion in
relictual grasslands. Biol Invasions 4:55–72
Invasive species establishment in a Hawaiian forest 819
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