Influence of woody invader control methods and seed availability on native and invasive species

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: [email protected]

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.


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)


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


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)


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


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


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


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


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


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.

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Influence of woody invader control methods and seed availability on native and invasive species