INVESTIGATING SEED DISPERSAL AND SEED PREDATION IN A HAWAIIAN

DRY FOREST COMMUNITY

IMPLICATIONS FOR CONSERVATION AND MANAGEMENT

A THESIS SUBMITTED TO THE GRADUATE DIVISION OF THE UNIVERSITY OF HAWAI‘I IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE

DEGREE OF

MASTER OF SCIENCE

IN

BOTANICAL SCIENCES

(BOTANY – ECOLOGY, EVOLUTION, AND CONSERVATION BIOLOGY)

DECEMBER 2004

By

Charles G. Chimera

Thesis Committee:

Donald Drake, Chairperson Gerald Carr David Duffy

Lloyd L. Loope

We certify that we have read this thesis and that, in our opinion, it is satisfactory in scope

and quality as a thesis for the degree of Master of Science in Botanical Sciences (Botany

– Ecology, Evolution and Conservation Biology).

THESIS COMMITTEE

Chairperson

ii

ACKNOWLEDGEMENTS

I would like to thank the EECB Program, the Hawai‘i Audubon Society, the Dai Ho

Chun Fellowship, and the Charles H. Lamoureux Fellowship in Plant Conservation for

providing funding for my research. I would like to thank Sumner Erdman and Tony

Durso of Ulupalakua Ranch for facilitating access to my study site through ranch lands. I

would like to thank the staff of the Natural Area Reserve System, particularly Betsy

Gagne, Bill Evanson and Bryon Stevens, for providing access to the Kanaio Natural Area

Reserve as well as for logistical support. Lloyd Loope and Art Medeiros were extremely

generous with their knowledge and guidance, factors that greatly influenced my decision

to conduct research in Hawaiian dry forests. Lloyd Loope also provided vehicle support

during the initial stages of my research. I thank my thesis committee chair, Don Drake,

and my committee members, Gerry Carr, Dave Duffy and Lloyd Loope for providing me

with guidance and assistance during all stages of my thesis research. Finally, I would

especially like to thank my wife, Melissa, for providing me with unlimited patience and

support during my entire graduate experience.

iii

TABLE OF CONTENTS

ACKNOWLEDGEMENTS.............................................................................................. III

LIST OF TABLES............................................................................................................ VI

LIST OF FIGURES ....................................................................................................... VIII

CHAPTER 1. INTRODUCTION ....................................................................................... 1 Loss of Native Seed Dispersers .................................................................................. 4 Rodents and Seed Predation........................................................................................ 5 Influence of Vegetation on Seed Dispersal and Predation.......................................... 6 Hypotheses.................................................................................................................. 8

CHAPTER 2. PATTERNS OF SEED DISPERSAL: VEGETATION STRUCTURE, FRUGIVORY AND SIZE EFFECTS .............................................................................. 10

Introduction................................................................................................................... 10 Loss of Native Seed Dispersers ................................................................................ 12 Influence of Vegetation on Seed Dispersal............................................................... 13

Methods......................................................................................................................... 15 Study Site .................................................................................................................. 15 Measurement of Seed Rain Under Trees Versus in Exposed Areas......................... 16 Seed Dispersal Analysis............................................................................................ 19 Bird Observations ..................................................................................................... 20 Seed Size Measurements........................................................................................... 21 Characterization of the vegetation ............................................................................ 22

Results........................................................................................................................... 22 Seed Rain .................................................................................................................. 22 Bird Observations ..................................................................................................... 33 Seed Size and Bird Dispersal.................................................................................... 35

Discussion..................................................................................................................... 36

CHAPTER 3: POST-DISPERSAL SEED PREDATION: LOCATION AND EFFECT ON SPECIES .................................................................................................................... 46

Introduction................................................................................................................... 46 Methods......................................................................................................................... 49

Study Site .................................................................................................................. 49 Measuring Seed Predation Levels............................................................................. 50 Treatments................................................................................................................. 51 Analysis..................................................................................................................... 53

Results........................................................................................................................... 54 Discussion..................................................................................................................... 62

iv

CHAPTER 4: CONCLUSION ......................................................................................... 68

APPENDIX A. TREE DIMENSIONS (BASAL DIAMETER, HEIGHT, CROWN DIAMETER) OF STUDY SPECIES AND DEAD TREES............................................. 73

APPENDIX B. PHENOLOGY OF STUDY TREES AND COMMON FLESHY-FRUITED SHRUBS OF THE KANAIO NATURAL AREA RESERVE....................... 75

Methods......................................................................................................................... 75

APPENDIX C. RELATIVE FREQUENCY AND ABUNDANCE OF NON-NATIVE BIRDS IN KNAR STUDY SITE ..................................................................................... 82

Methods......................................................................................................................... 82

APPENDIX D. CHARACTERIZATION OF THE VEGETATION OF KNAR STUDY SITE .................................................................................................................................. 87

Methods......................................................................................................................... 87

LITERATURE CITED ..................................................................................................... 91

v

LIST OF TABLES

Table 2.1. Mean seed density by dispersal vector and location at KNAR from April 2003 to March 2004. .......................................................................................................... 23

Table 2.2. Mean density under trees and in exposed sites, and percentage of total seed

rain for all species collected in 180 seed traps in KNAR from April 2003 to March 2004........................................................................................................................... 24

Table 2.3. Mean species richness per trap of fleshy-fruited seed rain under trees and in

exposed sites for six dry forest trees. ........................................................................ 25 Table 2.4. Mean density of seeds dispersed by birds and fleshy-fruited seed rain under

trees and in associated exposed sites for six dry forest trees in KNAR.................... 27 Table 2.5. Mean density of seeds dispersed by birds for 11 species under six dry forest

trees in KNAR........................................................................................................... 28 Table 2.6. Relative frequency of fleshy-fruited species in the seed rain under six dry

forest trees and all exposed seed traps at KNAR...................................................... 29 Table 2.7. Frequency of seedlings of fleshy-fruited species under six dry forest trees at

KNAR. ...................................................................................................................... 33 Table 2.8. Mean visits per hour and mean fruits consumed per foraging visit by seven

non-native bird species to five dry forest trees at KNAR.. ....................................... 35 Table 2.9. Mean seed length and width of 14 seeds collected in the seed rain at KNAR..

................................................................................................................................... 36 Table 3.1. Means of seed length and width and mean removal rate under trees and in

exposed sites for species used in the seed predation trials........................................ 55 Table 3.2. General linear models for number of seeds and/or fruits remaining versus

block, treatment, location, and the interaction of treatment and location after 15 days for four dry forest trees in KNAR............................................................................. 56

Table 3.3. Two-way ANOVA results for the effects of treatment and location on removal

rates for Pleomele auwahiensis................................................................................. 60 Table A.1. Tree dimensions of study species used in sampling of fleshy-fruited seeds and

seed removal trials. ................................................................................................... 73 Table B.1. Monthly percentage of fleshy-fruited trees and shrubs with mature fruit....... 76

vi

Table B.2. Mean monthly percentage of stems with immature fruits by tree species. ..... 76 Table B.3. Mean monthly percentage of stems with mature fruits by tree species .......... 77 Table B.4. Mean monthly number of immature fruits and mature fruits on branches by

shrub species. ............................................................................................................ 77 Table C.1. Relative frequency and abundance of bird species recorded at 10 sampling

stations over a 12-month period in KNAR. .............................................................. 83 Table D.1. Abundance of ground vegetation at KNAR study site.................................... 88 Table D.1 (Continued) Abundance of ground vegetation at KNAR study site. ............... 89 Table D.2. Abundance of canopy vegetation at KNAR study site.. ................................. 89 Table D.3. Abundance of native and non-native vegetation, dead trees, bare ground, and

open sky at KNAR study site.................................................................................... 90

vii

LIST OF FIGURES

Figure 1.1. Remnant dry forests of the Hawaiian Islands................................................... 3 Figure 2.1. Study area within Kanaio Natural Area Reserve............................................ 16 Figure 2.2. Mean density of seeds dispersed by birds under six dry forest tree species... 30 Figure 2.3. Mean density of conspecific seeds under five dry forest trees at KNAR....... 31 Figure 2.4. Mean density of Bocconia seeds versus conspecific seeds under four dry

forest trees at KNAR................................................................................................. 32 Figure 3.1. Removal rates of seeds and fruits from the KNAR........................................ 57 Figure 3.2. Removal rates for Bocconia frutescens and Diospyros sandwicensis seeds and

fruits under trees and in exposed sites. ..................................................................... 58 Figure 3.3. Removal rates for Pleomele auwahiensis seeds and fruits under trees and in

exposed sites ............................................................................................................. 59 Figure 3.4. Removal rates for Pleomele auwahiensis seeds and fruits on the open ground

under trees and in exposed sites................................................................................ 60 Figure 3.5. Removal rates for Reynoldsia sandwicensis and Santalum ellipticum seeds

under trees and in exposed sites................................................................................ 61 Figure A.1. Tree dimensions for study species and used in sampling of fleshy-fruited

seeds.......................................................................................................................... 74 Figure B.1 Monthly percentage of fleshy-fruited trees with mature fruit......................... 78 Figure B.2. Monthly percentage of fleshy-fruited shrubs with mature fruit..................... 79 Figure B.3. Mean monthly percentage of stems with immature and mature fruits by tree

species ....................................................................................................................... 80 Figure B.4. Mean monthly number of immature and mature fruits on branches by shrub

species ....................................................................................................................... 81 Figure C.1. Relative frequency of bird species recorded at 10 sampling stations over a 12-

month period in KNAR............................................................................................. 84

viii

Figure C.2. Relative abundance of bird species recorded at 10 sampling stations over a 12-month period in KNAR. ...................................................................................... 85

Figure C.3. Mean number of three important non-native frugivores at 10 sampling

stations over a 12-month period in KNAR. .............................................................. 86

ix

CHAPTER 1. INTRODUCTION

BACKGROUND

Tropical dry forests are among the most diverse, yet imperiled, natural

communities worldwide (Janzen 1988; Lerdau et al. 1991), and those of the Hawaiian

Islands are no exception (Loope 1998). Tropical dry forests, which are found from

Mexico to South America, Africa, Australia, Southeast Asia, and many of the world’s

tropical islands, are characterized by a mean annual rainfall of 250 to 2000 mm, mean

annual temperatures higher than 17ºC, and one to two dry periods per year (Murphy and

Lugo 1986). Unlike other tropical dry forests, which generally contain fewer tree species

than neighboring rain forest does (Murphy and Lugo 1986; Janzen 1988), Hawaiian dry

forests are richer in tree diversity than comparable areas of wet forest (Rock 1913;

Carlquist 1980; Sohmer and Gustafson 1987). Like their global counterparts, which have

experienced a rapid and significant loss of area throughout the world (Murphy and Lugo

1986; Janzen 1988; Bullock et al. 1995), these unique Hawaiian communities have now

been reduced to approximately 10% of their former extent (Bruegmann 1996; Mehrhoff

1998). Extensive impacts on and alterations of these Hawaiian ecosystems began with the

agricultural and hunting practices of the early Polynesians, their use of fire for land

clearing, and the introduction of non-native animals such as the Polynesian rat, (Rattus

exulans) (Kirch 1982; Sadler 1999; Burney et al. 2001; Athens et al. 2002). This

deterioration and loss accelerated after the arrival of Europeans. As has occurred in dry

forests throughout the world (Murphy and Lugo 1986; Janzen 1988), the introduction of

ungulates such as cattle and goats, further land clearing for agriculture and development, 1

accidental and intentional anthropogenic fires, and the introduction of aggressive weeds,

particularly fire-carrying grasses such as fountain grass (Pennisetum setaceuma), all

contributed to the rapid decline of the Hawaiian dry forests (Stone 1989; Cuddihy and

Stone 1990; Loope 1998).

Despite these losses, however, biologically diverse areas of dry forest still remain

on several of the Hawaiian Islands. Notable remnant forests can be found at Pu`u Wa`a

Wa`a on the island of Hawai‘i, on the southern slopes of Maui at Auwahi, Kanaio and

Pu`u-o-kali, on Lanai at Kanepu`u, on Molokai at Kauhako Crater, and at Mokuleia on

Oahu (Figure 1.1; Medeiros et al. 1984; Medeiros et al. 1986; Cuddihy 1989; Medeiros et

al.1996; Loope 1998; Mueller-Dombois and Fosberg 1998). Nevertheless, all of the

factors that contributed to the initial loss still threaten the integrity of the remaining dry

forests, and other threats continue to impact the health and functioning of these

ecosystems (Stone et al. 1992; Pratt and Gon 1998; Blackmore and Vitousek 2000;

Staples and Cowie 2001). One particularly troubling phenomenon that seriously affects

the persistence as well as the potential future restoration of these communities is the

almost complete lack of natural reproduction of native tree species (Medeiros et al. 1986;

Medeiros et al. 1993; Loope et al. 1995; Loope 1998; Cabin et al. 2000). Janzen (1988)

refers to adult trees that fail to reproduce as “the living dead” and indicates that this

problem is symptomatic of tropical dry forests worldwide.

a Plant taxonomy follows Wagner et al. (1999)

2

Figure 1.1. Remnant dry forests of the Hawaiian Islands

In addition to browsing and trampling of seedlings by feral ungulates, and

competition with aggressive, non-native weeds, several other factors may contribute to

the lack of seedling recruitment in tropical dry forests. These include the loss of native

pollinators with subsequent reduction or loss of outcrossing, the impacts of non-native

invertebrates and pathogens on seed and seedling survival, modifications in microclimate

and microhabitats suitable for seed germination and seedling survival, loss of native birds

that scarified and dispersed seeds, and seed predation by introduced rodents (Janzen

1988; Medeiros et al. 1993; Aide et al. 2000; Cabin et al. 2000; Holl et al. 2000;

Zimmerman et al. 2000). The loss of native dispersal agents and the likely modification

of the existing seed shadows, or seed distribution patterns around the source, as well as 3

the effects of rodent predation on seeds, are two factors that could be influencing

community composition and structure of dry forests throughout the Hawaiian Islands.

Loss of Native Seed Dispersers

Direct and indirect impacts of both Polynesians and Europeans have led to the

widespread extinction of native birds throughout Polynesia, including the Hawaiian

Islands (Kirch 1982; James and Olson 1991; Olson and James 1982, 1991; Steadman

1995; Burney et al. 2001; Ziegler 2002). What roles these birds may have played in the

scarification and dispersal of native tree seeds is unknown, but others have suggested that

a loss of native dispersers and replacement by more generalist, non-native dispersal

agents could affect both the shape of the seed shadow and the types of seeds being

dispersed (Temple 1977; Cole et al. 1995; Willson and Traveset 2001; Loiselle and Blake

2002; Meehan et al. 2002). As plants that depend on animals for seed transport are

susceptible to dispersal failure when their seed vectors become rare or extinct, disruptions

to the mutualism can have serious consequences for the maintenance of plant populations

(Willson and Traveset 2001). This may be particularly true of Hawaiian dry forest areas,

in which almost two-thirds of native tree and shrub species have fleshy fruits presumably

adapted for some type of bird dispersal (Medeiros et al. 1993). These areas once

supported a diverse and conspicuous native avifauna prior to the arrival of the

Polynesians (James and Olson 1991; Olson and James 1991). Other than the native owl

or pueo (Asio flammeus sandwicensis), the Pacific Golden plover or kolea (Pluvialis

fulva) and the infrequently seen Hawaiian goose or Nene (Branta sandvicensis), recent

4

surveys within Kanaio recorded only introduced game birds and small, generalist

frugivores such as the Japanese white-eye (Zosterops japonicus) and the common myna

(Acridotheres tristis) (Medeiros et al. 1993; Chimera pers. obs.). This replacement of

native with non-native birds is common to the lowland dry forests of the Hawaiian

Islands (Pratt et al. 1987). Although little is known about how the loss or gain of a

dispersal agent alters a plant’s seed shadow, potential consequences could include

changes in the amounts and sizes of seeds being dispersed (Willson and Traveset 2001).

The size and structure of the fruits and seeds that birds can consume or transport are

constrained by the size of the bird (Wheelwright 1985; Stiles 2001), such that small, alien

birds may be preferentially consuming only smaller fruits and seeds. Anecdotal

observations of seedling distributions suggest that birds are dispersing some small-seeded

native and non-native taxa, of which several of the latter are serious, habitat-modifying

weeds. It is unknown to what degree these introduced birds act as surrogates of the

extinct avifauna in the dispersal of seeds, how frequently the seeds of these and other

species, if any, are being dispersed, and whether or not there is a size limit to seeds that

may be dispersed.

Rodents and Seed Predation

The flora and fauna of Pacific island ecosystems, including the Hawaiian Islands,

evolved in the absence of native rodents, and presumably lack defenses against them.

Introduced rodents such as the Polynesian rat (Rattus exulans), the black or roof rat (R.

rattus), the Norway rat (R. norvegicus) and the house mouse (Mus musculus) have been

5

inadvertently introduced to many island chains throughout Polynesia, with devastating

impacts on the insular biota (Tomich 1986). In the Hawaiian Islands, rats have negatively

impacted the native avifauna (Atkinson 1977; Scott et al. 1986), the endemic snail and

arthropod fauna (Hadfield et al. 1993; Cole et al. 2000) and the flora, including many

rare and endangered species (Wirtz 1972; Baker and Allen 1978; Russel 1980; Scowcroft

and Sakai 1984; Stone et al. 1984; Stone 1985; Sugihara 1997; Cole et al. 2000).

Although introduced rodents are notorious plant and seed predators in other insular

ecosystems (Fall et al. 1971; Clark 1982; Campbell et al. 1984; Allen et al. 1994; Moles

and Drake 1999; Williams et al. 2000; Campbell and Atkinson 2002; Delgado Garcia

2002; McConkey et al. 2003), and have been anecdotally implicated in the destruction of

seeds in dry forest ecosystems, including the Kanaio Natural Area Reserve (Medeiros et

al. 1986; Medeiros et al. 1993; Cabin et al. 2000), no studies have been published that

quantitatively document the impacts of these introduced rodents on the seeds of particular

dry forest taxa. As many of these species produce fruits with one to few larger seeds

(Wagner et al. 1999), and as rodents in general have been shown to be voracious

predators of intermediate and larger seeds (Stiles 2001), the effects that introduced

rodents have on seed mortality and subsequent levels of seedling recruitment could be

profound (Crawley 2001).

Influence of Vegetation on Seed Dispersal and Predation

A modified seed dispersal regimen due to losses of native dispersal agents would

be expected over time to influence structure and composition of dry forest communities.

6

A scattered distribution of dry forest trees would, in turn, play an influential role in the

deposition of seeds throughout an area. Remnant Hawaiian dry forests like that in

Kanaio, Maui, typically consist of isolated trees and scattered groves of trees surrounded

by native shrublands, non-native grasslands or barren lava flows (Medeiros et al. 1993;

Mueller-Dombois and Fosberg 1998). The highest densities in a seed shadow normally

occur close to the seed source (Willson and Traveset 2001). In the case of fruiting trees,

greatest seed densities would therefore be expected under tree crowns. In addition,

directed dispersal often results in non-random seed shadows due to the preferential

movements of birds to these feeding and perch sites (Howe and Primack 1975; Jordano

and Schupp 2000; Wenny 2001; Clark et al. 2004). By providing suitable microsites for

the growth of seedlings, isolated trees or groves of trees in otherwise open areas often act

as recruitment foci for bird-disseminated species (Yarranton and Morrison 1974;

Debussche et al. 1982; McDonnell and Stiles 1983; Izhaki et al. 1991; Guevara and

Laborde 1993; Debussche and Isenmann 1994; Ferguson and Drake 1999). Isolated trees

may therefore play an important role in the succession and regeneration of formerly

disturbed or degraded areas (Uhl et al. 1982; Guevara et al. 1986; Eriksson and Ehrlen

1992; Loiselle and Blake 1993; Martinez-Ramos and Soto-Castro 1993; Uhl 1998;

Wijdeven and Kuzee 2000; Cordell et al. 2002).

Despite the potential benefits, however, this non-random deposition under trees,

combined with the higher seed rain predicted by the seed shadow under parent trees,

could also influence seed predation rates (Hulme 1998). Several authors have

documented higher predation rates under trees and vegetation, as opposed to more open

areas (Howe et al. 1985; Callaway 1992; Aide and Cavelier 1994; Osunkoya 1994;

7

Chapman and Chapman 1996; Hau 1997; Hulme 1997; Wenny 2000; Holl 2002). Others

have found no difference (DeSteven and Putz 1984; Terborgh et al. 1993), or higher rates

of predation in exposed areas or grasslands (Hay and Fuller 1981; Uhl 1998; Wijdeven

and Kuzee 2000). One study reported lower predation rates under isolated pasture trees as

opposed to either open pastures or intact forests (Holl and Lulow 1997).

To better predict how the interaction between the scattered distribution of dry

forest trees, seed dispersal and seed predation may shape the dry forest community

composition of the Kanaio Natural Area Reserve, I addressed the following hypotheses as

part of my Master’s Thesis:

Hypotheses

Seed Dispersalb

1) There is no difference in the seed rain of fleshy-fruited, bird-dispersed seeds

under trees versus in exposed areas.

2) Non-native birds such as the Japanese white-eye (Zosterops japonicus), the

common mynah (Acridotheres tristis), the zebra dove (Geopelia striata), the

northern cardinal (Cardinalis cardinalis) and the northern mockingbird (Mimus

polyglottos) are not dispersing the seeds of fleshy-fruited native and non-native

trees.

3) There is no relationship between seed size and quantity of seeds dispersed for

fleshy-fruited species.

Seed Predation

8

b The term “seed” is used broadly in this proposal to refer to the persistent part of the diaspore. Thus, for Reynoldsia sandwicensis, the units studied will be the multiple pyrenes within the drupe.

4) Rodents are not removing and destroying seeds of dry forest species.

5) There is no difference in removal rates of seeds under trees versus exposed sites.

6) There is no relationship between seed size and removal rates for dry forest tree

species.

9

CHAPTER 2. PATTERNS OF SEED DISPERSAL: VEGETATION

STRUCTURE, FRUGIVORY AND SIZE EFFECTS

INTRODUCTION

Tropical dry forests, including those of the Hawaiian Islands, are among the

world’s most diverse, yet imperiled, natural communities (Janzen 1988; Lerdau et al.

1991; Loope 1998). Unlike other tropical dry forests, with generally fewer species than

neighboring rain forest (Murphy and Lugo 1986; Janzen 1988), Hawaiian dry forests

have greater tree diversity than comparable areas of wet forest (Rock 1913; Carlquist

1980; Sohmer and Gustafson 1987). Like their global counterparts, which have

experienced a rapid and significant loss in area (Murphy and Lugo 1986; Janzen 1988;

Bullock et al. 1995), these unique Hawaiian communities have now been reduced to

approximately 10% of their former cover (Bruegmann 1996; Mehrhoff 1998). Extensive

alterations of these ecosystems began with the agricultural and hunting practices of the

early Polynesians, and with the introduction of non-native animals such as the Polynesian

rat, (Rattus exulans) (Kirch 1982; Sadler 1999; Burney et al. 2001; Athens et al. 2002).

This deterioration and loss accelerated after the arrival of Europeans. Hawaiian dry

forests rapidly declined following the introduction of ungulates such as cattle (Bos

taurus) and goats (Capra hircus), land clearing for agriculture and development,

accidental and intentional anthropogenic fires, and the introduction of aggressive weeds,

factors commonly attributed to the loss of dry forests throughout the world (Murphy and

Lugo 1986; Janzen 1988; Stone 1989; Cuddihy and Stone 1990; Loope 1998).

10

Despite these losses, however, biologically diverse, if fragmented, areas of dry

forest still remain on several of the Hawaiian Islands (Medeiros et al. 1984; Medeiros et

al. 1986; Loope 1998; Mueller-Dombois and Fosberg 1998). Nevertheless, all of the

factors that contributed to the initial loss still threaten the integrity of the remaining dry

forests, and other threats continue to impact the health and functioning of these

ecosystems (Stone et al. 1992; Pratt and Gon 1998; Blackmore and Vitousek 2000;

Staples and Cowie 2001). One particularly troubling phenomenon that seriously affects

the persistence as well as the potential future restoration of these communities is the

almost complete lack of natural reproduction of native tree species (Medeiros et al. 1986;

Medeiros et al. 1993; Loope et al. 1995; Loope 1998; Cabin et al. 2000). Janzen (1988)

refers to adult trees that fail to reproduce as “the living dead” and indicates that this

problem is symptomatic of tropical dry forests worldwide.

In addition to browsing and trampling of seedlings by feral ungulates, and

competition with non-native weeds, several other factors may contribute to the lack of

seedling recruitment in tropical dry forests. These include the loss of native pollinators,

the impacts of non-native invertebrates and pathogens on seed and seedling survival,

modifications in microclimate and microhabitats suitable for seed germination and

seedling survival, loss of native birds that scarified and dispersed seeds, and seed

predation by introduced rodents (Janzen 1988; Medeiros et al. 1993; Aide et al. 2000;

Cabin et al. 2000; Holl et al. 2000; Zimmerman et al. 2000). The loss of native dispersal

agents and the likely modification of the existing seed shadows are factors that could be

influencing community composition and structure of Hawaiian dry forests.

11

Loss of Native Seed Dispersers

Direct and indirect impacts of both Polynesians and Europeans have led to the

widespread extinction of native birds throughout Polynesia, including the Hawaiian

Islands (Kirch 1982; James and Olson 1991; Olson and James 1982, 1991; Steadman

1995; Burney et al. 2001; Ziegler 2002). What roles these birds may have played in the

scarification, dispersal and predation of native tree seeds is unknown, but others have

suggested that a loss of native dispersers and replacement by more generalist, non-native

dispersal agents could affect both the shape of the seed shadow and the types of seeds

being dispersed (Temple 1977; Cox et al. 1991; Cole et al. 1995; Loiselle and Blake

2002; Meehan et al. 2002; McConkey and Drake 2002). Plants that depend on animals

for seed transport are susceptible to dispersal failure when their vectors become rare or

extinct, and disruptions to the mutualism can have serious consequences for the

maintenance of plant populations (Willson and Traveset 2001). This may be particularly

true of Kanaio and adjacent dry forest areas on Maui, in which over 59% of native trees

and shrubs have fleshy fruits presumably adapted for some type of bird dispersal

(Medeiros et al. 1993). These areas once supported a diverse and conspicuous native

avifauna prior to the arrival of the Polynesians (James and Olson 1991; Olson and James

1991) but are now almost entirely dominated by introduced game birds and non-native

passerines such as the Japanese white-eye (Zosterops japonicus) (Medeiros et al. 1993;

Chimera pers. obs.). This replacement of native with non-native birds is common in the

lowland dry forests of the Hawaiian Islands (Pratt et al. 1987). Little is known about how

the loss or addition of a dispersal agent alters the seed shadow of a plant, but potential

consequences could include changes in the amounts and sizes of seeds being dispersed

12

(Willson and Traveset 2001). The size and structure of the fruits and seeds that birds can

consume or transport are constrained by the size of the bird (Wheelwright 1985; Stiles

2001), such that small, non-native birds may be preferentially consuming only smaller

fruits and seeds. Yet it is unknown to what degree these introduced birds act as surrogates

of the extinct avifauna in the dispersal of seeds, how frequently the seeds of these and

other species, if any, are being dispersed, and whether or not there is a size limit to seeds

that may be dispersed.

Influence of Vegetation on Seed Dispersal

A modified seed dispersal regimen due to losses of native dispersal agents would

be expected to influence structure and composition of dry forest communities. The spatial

distribution of dry forest trees would, in turn, play an influential role in the deposition of

seeds throughout an area. Remnant Hawaiian dry forests like that in Kanaio, Maui,

typically consist of isolated trees and scattered groves of trees surrounded by native

shrublands, non-native grasslands or barren lava flows (Medeiros et al. 1993; Mueller-

Dombois and Fosberg 1998). The highest densities in a seed shadow normally occur

close to the seed source (Willson and Traveset 2001). In the case of fruiting trees,

greatest seed densities would therefore be expected under tree crowns. In addition,

directed dispersal often results in non-random seed shadows due to the preferential

movements of birds to these feeding and perch sites (Howe and Primack 1975; Jordano

and Schupp 2000; Wenny 2001; Clark et al. 2004). By providing suitable microsites for

the growth of seedlings, isolated trees or groves of trees in otherwise open areas often act

13

as recruitment foci for bird-disseminated species (Yarranton and Morrison 1974;

Debussche et al. 1982; McDonnell and Stiles 1983; Izhaki et al. 1991; Guevara and

Laborde 1993; Debussche and Isenmann 1994; Ferguson and Drake 1999). Isolated trees

may therefore play an important role in the succession and regeneration of formerly

disturbed or degraded areas (Uhl et al. 1982; Guevara et al. 1986; Eriksson and Ehrlen

1992; Loiselle and Blake 1993; Martinez-Ramos and Soto-Castro 1993; Uhl 1998;

Wijdeven and Kuzee 2000; Cordell et al. 2002).

This study attempts to predict how the interaction between the distribution of dry

forest trees and seed dispersal patterns by non-native frugivores may shape the dry forest

community composition of KNAR by addressing the following questions:

1) Is there a difference in the seed rain of fleshy-fruited, bird-dispersed seeds

under trees versus in exposed areas?

2) Are non-native birds acting as surrogates of the native avifauna and dispersing

the seeds of fleshy-fruited native trees, or are they primarily dispersing seeds

of non-native plants?

3) Is there a relationship between seed size and amount of seeds being dispersed

by the non-native birds?

14

METHODS

Study Site

This study was conducted in the Kanaio Natural Area Reserve (KNAR), on the

leeward side of East Maui, Hawai‘i, at an elevation between 750-850 meters (20º36’N,

156º20’W; Figure 2.1). The climate in the reserve is hot and dry, with annual

temperatures between 20-30° C and with mean annual rainfall of approximately 750 mm,

mostly falling between the months of October to April (Giambelluca et al. 1986). The

substrate is estimated to be less than 10,000 years old and consists mostly of `a`a lava

with some accumulation of overlying soils patchily distributed throughout the area

(Crandell 1983). Medeiros et al. (1986; 1993) have provided a thorough description of

the community composition of the dry forest of KNAR and the leeward slopes of

Haleakala, Maui. Supplemental characterization of the vegetation was provided using a

modified point-intercept method (Bonham 1989; Appendix D). The 354-hectare reserve

was established in 1990 to protect exemplary representatives of Hawaiian dry forest

species and plant communities, including Dodonaea (`A`ali`i) lowland shrublands,

Diospyros (Lama) forest and Erythrina (Wiliwili) forests (Gagne and Cuddihy 1990). The

reserve also contains individuals and groves of at least twenty-two relatively short-

statured native tree species, with typical tree heights ranging between three to ten meters.

Anecdotal observations suggest that only eight of these species may currently be

reproducing (Medeiros et al. 1993). Certain non-native trees, however, including the

widespread Bocconia frutescens (Papaveraceae), are able to reproduce without the

apparent difficulties experienced by native species (Chimera pers. obs.).

15

Figure 2.1. Study area within Kanaio Natural Area Reserve (20º36’N, 156º20’W).

Measurement of Seed Rain Under Trees Versus in Exposed Areas

To quantify the difference in seed rain under trees and in exposed areas, one seed

trap was placed under each study tree and one in an exposed site at least five meters from

the edge of each study tree’s crown. This distance was sufficient to prevent fruits or seeds

from falling off the parent trees directly into seed traps in exposed sites, as all study trees

were relatively short-statured (Appendix A). Traps under trees were placed in a random

16

compass direction approximately half the distance from the trunk to the crown edge. Seed

traps in the exposed areas were placed in a random compass direction away from the

parent tree, such that the position of the seed trap was not within five meters of any other

tree. These exposed sites consisted of barren rock, low-statured grasses, woody shrubs

and annual or perennial herbaceous species generally less than 50 cm tall. Seed traps in

exposed sites were not placed directly under woody shrubs, although they were placed

next to shrubs when encountered. Fifteen trees of each of four different native dry forest

species, of one non-native species and fifteen dead, standing trees were selected from the

northeastern portion of KNAR, the area of greatest tree diversity and abundance

(Medeiros et al. 1993). For purposes of comparison, the 15 dead, standing trees were

arbitrarily considered a sixth study “species”. Trees were selected from those that were

not directly adjacent to a conspecific individual. Species of native trees were also selected

on the basis of availability of enough individuals meeting the spacing criterion and

having fruits with adaptations for bird dispersal (Jordano 2001; Stiles 2001). Native

species were selected to represent a range of fruit and seed sizes. Trees were selected

along 12 parallel, east-west oriented transects, ranging in length from 100 to 675 meters

and spaced 50 meters apart. Transects stopped within 5 m of non-forest vegetation on

adjacent ranchlands, thus accounting for differences in transect length. All appropriate

study trees were first recorded along the length of each transect. Study trees were

randomly selected from this subset of individuals such that no trees were closer than 25

meters to each other. The entire study site encompassed an area of approximately 0.25

km2. Native fleshy-fruited taxa used for this study include Reynoldsia sandwicensis,

(Araliaceae), Diospyros sandwicensis (Ebenaceae), Santalum ellipticum (Santalaceae),

17

and Pleomele auwahiensis (Agavaceae). For the dioecious D. sandwicensis, only female,

fruit-producing trees were used. One non-native tree, Bocconia frutescens

(Papaveraceae), was also used. This tree is common in the reserve, and its spread is likely

being facilitated by non-native birds attracted to the red, pulpy aril attached to the shiny

black seeds (Wagner et al. 1999). The fifteen dead standing trees were selected so that

they were roughly similar in branch architecture and height. Basal diameter, height, and

crown diameter were recorded for each study tree (Appendix A). Seed vouchers were

collected from all fruiting trees in the vicinity during the duration of the study to aid in

the identification of seeds in the traps.

Seed traps were constructed of pairs of circular plastic pots with the bottoms

removed. One pot was placed inside the other with a cotton cloth in between, a

modification of the design described by Drake (1998). Traps were 23 cm deep with a

25.4 cm diameter, such that the total area sampled under each of the six tree species (15

traps/species) and in the adjacent open area (15 traps/species) was 0.76 m2 each. Trap

tops were covered with wire screens with 4 x 3 x 2.5 cm hexagonal apertures, large

enough to allow fruits and seeds to fall through and small enough to deter rodents or

game birds from consuming seeds.

Traps were emptied and the liners replaced every month, for a twelve-month

period (April 2003 to March 2004). Numbers of seeds in traps were counted by species

and the number of intact fruits from the parent tree was also recorded. Because fruits of

Diospyros sandwicensis, Pleomele auwahiensis and Reynoldsia sandwicensis potentially

contained multiple seeds, 50 of each haphazardly-selected fruit were cut open and the

seeds inside counted. Means were 1.54 ± 0.11 seeds per D. sandwicensis fruit, 1.3 ± 0.09

18

seeds per P. auwahiensis fruit and 9.08 ± 0.23 seeds per R. sandwicensis fruit. Any fruits

from these species found in traps were therefore multiplied by the mean. In addition,

seeds were investigated for any invertebrate or rodent damage. The presence of immature

and ripe fruits on sample trees and common taxa in the area was recorded on a monthly

basis from March 2003 to February 2004 to document part of the potential pool of bird-

dispersed seeds that could be disseminated in the area (Appendix B).

Seed Dispersal Analysis

All seeds collected in traps were counted and categorized by one of four general

dispersal vectors: bird-dispersed; gravity-dispersed; uncertain; wind-dispersed. Fleshy-

fruited seeds collected in traps under trees were quantified as “bird-dispersed” if they

were from species other than that of the overhanging tree. All fleshy-fruited seeds

collected in traps in exposed areas or under dead, standing trees were counted as “bird-

dispersed”. Seeds from intact fruits that had fallen directly from the parent tree into seed

traps were categorized as “gravity-dispersed”. Seeds from one weedy annual, Bidens

pilosa (Asteraceae), were present in many traps and were categorized as “gravity-

dispersed” because they often fell from plants overtopping seed traps. Seeds lacking pulp

but captured beneath a conspecific tree were counted as “uncertain”. This category

accounts for the possibilities that birds may have dropped seeds from the parent tree or

that invertebrates consumed pulp from fruits after they fell from parent trees. Smaller,

wind-borne seeds and those with adaptations for dispersal by wind currents (e.g. tufts of

hairs, winged-fruits etc.) were classified as “wind-dispersed”.

19

As several seeds without pulp were collected in traps under conspecifc trees, a

large number of seeds were categorized as having an “uncertain” dispersal vector. These

data were combined with “bird-dispersed” data, and fleshy-fruited “gravity-dispersed”

data into a single category: all fleshy-fruited seeds, to indicate that all seeds in this

category possess adaptations for bird dispersal. Fleshy-fruited seed rain was then

compared between trees and exposed areas for all species. Total seed rain numbers were

converted into seed densities (seeds m-2) and square root transformed to improve

normality. Bird-dispersed and fleshy-fruited seed densities were compared between trees

and associated exposed sites with paired t-tests. Species richness of fleshy-fruited seeds

was also compared between trees and exposed sites with paired t-tests. (Zar 1999).

Fleshy-fruited seed rain of native and non-native species was compared under trees of

each species using paired t-tests. Fleshy-fruited seed rain of particular seed species (see

Results) was compared between different tree species using a one-way analysis of

variance, with Tukey tests determining which densities were significantly different (Zar

1999).

As an indirect estimate of seed dispersal, the presence/absence of all species of

fleshy-fruited seedlings less than 10 cm tall was noted under the crowns of all study trees.

For presentation and discussion of results, dead trees are treated as a “species”.

Bird Observations

To determine what birds, if any, were consuming and potentially dispersing the

seeds of the study trees, 30-40 hours of timed watches per study tree species were

20

conducted to coincide with fruiting. A different haphazardly-selected fruiting tree was

observed each day for a total of seven to ten days. Only trees estimated to have mature

fruit on at least 25 percent of tree branches were used. Observations took place in the

morning, from sunrise until as late as 1130 a.m., and only occurred on days when

climatological conditions were favorable (i.e. absence of heavy rains or extremely strong

winds). Observations were made with binoculars and a spotting scope at a distance and in

a position that was unlikely to affect bird activity. Bird species, foraging visits per hour

and fruit consumed per visit were recorded for each study tree. Observations were

conducted on Pleomele auwahiensis for eight days in May and June 2003, on Bocconia

frutescens for ten days in May and June 2003, on Diospyros sandwicensis for six days in

December 2003 and January 2004, on Reynoldsia sandwicensis for nine days in January

2004, and on Santalum ellipticum for seven days in February 2004. Anecdotal

observations of frugivory on non-study species were also noted. Because observations

were conducted on different numbers of days and under different weather conditions, bird

visitation data were not compared statistically between species.

To assess the relative frequency and abundance of bird species at the study site,

additional sampling was conducted from April 2003 to March 2004 at 10 stations along a

bird transect roughly bisecting the seed sampling transects. For more information on

methods used, and for results, consult Appendix C.

Seed Size Measurements

21

To determine the range of sizes of dispersed seeds, length and width were

recorded for those seeds collected in seed traps and classified as “bird-dispersed” (see

Seed Dispersal Analysis section). Spearman’s rank correlation was used to test for

relationships between seed length, width, and total numbers of bird-dispersed seeds

collected in all seed traps. Although no seeds of Pleomele auwahiensis, Santalum

ellipticum, or Reynoldsia sandwicensis were recorded as “bird-dispersed”, correlations

were run both with and without size dimensions of these species as they are common to

the area and were specific subjects of the seed dispersal analysis.

Characterization of the vegetation

Plant cover in the study site was sampled at 1000 points using the point-intercept

method (Bonham 1989). For a more detailed description of methods used, and results,

consult Appendix D.

RESULTS

Seed Rain

A total of 12,769 seeds (Table 2.1) from at least 32 different species (Table 2.2)

were found in the 180 traps under trees and in exposed sites. Over 90% of seeds were

non-native. Of these, two wind-dispersed grasses, Melinis repens and M. minutiflora,

dominated the seed rain, comprising 55.6% and 10.6% of the total density respectively

(Table 2.2).

22

Table 2.1. Mean density (seeds m-2 ± 1 SEM) by dispersal vector and location (tree vs. exposed sites) at KNAR from April 2003 to March 2004. Densities under trees versus exposed sites were compared with paired t-tests on square root transformed data. Seed densities were not compared between different dispersal vectors.

Dispersal Vector Tree Exposed P-value % TotalWind 595.6 (96.5) 1362.4 (283.6) 0.167 70Gravity 321.0 (80.1) 42.5 (17.7) <0.001 13.0Birds 237.9 (30.2) 7.7 (2.97) <0.001 8.8Uncertain 230.5 (68.0) 1.3 (0.75) <0.001 8.3

23

Table 2.2. Mean density (seeds m-2) under trees and in exposed sites (Exp.), and percentage of total seed rain for all species collected in 180 seed traps in KNAR from April 2003 to March 2004. N = native; A = alien/non-native.

Seed Category Family Status Tree Exp. % Total Melinis repens Poaceae A 373.7 1181.7 55.55 Bocconia frutescens Papaveraceae A 473.0 6.1 17.11 Melinis minutiflora Poaceae A 162.3 133.1 10.55 Bidens pilosa Asteraceae A 78.3 41.0 4.26 Reynoldsia sandwicensis Araliaceae N 96.0 0.0 3.43 Dodonaea viscosa Sapindaceae N 33.6 26.8 2.14 Diospyros sandwicensis Ebenaceae N 51.8 0.0 1.85 Lantana camara Verbenaceae A 47.6 3.5 1.82 Santalum ellipticum Santalaceae N 16.2 0.0 0.58 Chenopodium oahuense Chenopodiaceae N 11.4 0.0 0.41 Galinsoga parviflora Asteraceae A 1.3 8.8 0.36 Pleomele auwahiensis Agavaceae N 9.2 0.0 0.33 Nothocestrum latifolium Solanaceae N 7.2 0.0 0.26 Sonchus oleraceus Asteraceae A 2.2 4.0 0.22 Osteomeles anthyllidifolia Rosaceae N 5.3 0.0 0.19 Unidentified (one spp.) NA NA 2.9 2.0 0.17 Ageratina adenophora Asteraceae A 1.5 0.9 0.09 Emilia fosbergii Asteraceae A 0.4 2.0 0.09 Wikstroemia monticola Thymelaeaceae N 2.4 0.0 0.09 Asclepias physocarpa Ascelpiadaceae A 2.2 0.0 0.08 Sporobolus africanus Poaceae A 0.0 2.2 0.08 Cocculus orbiculatus Menispermaceae N 1.8 0.0 0.06 Digitaria ciliaris Poaceae A 1.8 0.0 0.06 Zinnia peruviana Asteraceae A 0.0 1.8 0.06 Petroselinum crispum Apiaceae A 0.7 0.9 0.05 Leptecophylla tameiameiae Epacridaceae N 1.1 0.0 0.04 Conyza bonariensis Asteraceae A 0.0 0.4 0.02 Schinus terebinthifolius Anacardiaceae A 0.4 0.0 0.02 Alyxia oliviformis Apocynaceae N 0.2 0.0 0.01 Hyochoeris radicata Asteraceae A 0.2 0.0 0.01 Myoporum sandwicense Myoporaceae N 0.2 0.0 0.01 Tridax procumbens Asteraceae A 0.2 0.0 0.01

24

Table 2.3. Mean species richness (± 1 SE) per trap of fleshy-fruited seed rain under trees and in exposed sites for six dry forest tree species. All pairwise comparisons between trees and exposed sites are significantly different (paired t-test; P < 0.05).

Tree Species Tree Exposed Bocconia frutescens 2.07 (0.18) 0.07 (0.07) Dead Trees 2.53 (0.24) 0.13 (0.09) Diospyros sandwicensis 3.40 (0.21) 0.27 (0.12) Pleomele auwahiensis 2.60 (0.40) 0.40 (0.16) Reynoldsia sandwicensis 3.33 (0.33) 0.07 (0.07) Santalum ellipticum 2.27 (0.36) 0.00 (0.00)

Mean species richness of fleshy-fruited seeds per trap ranged from 0-6 species

under trees and 0-2 species in exposed sites. In all cases, species richness was higher

under trees than in exposed sites (Table 2.3). For all study tree species, there were

significantly more bird-dispersed seeds collected under trees than in associated exposed

areas (Table 2.4). In the fleshy-fruited seed categories, only 11 species were conclusively

bird-dispersed. Of these, eight were native species and three were non-native (Table 2.5).

Of the 1085 bird-dispersed seeds in the seed rain, 96.8% were collected under trees, and

3.2% were collected in exposed sites. Although more species of natives were dispersed

by birds, total seed rain was dominated by two non-native species, Bocconia frutescens

and Lantana camara, comprising 74.7% and 17.7% of the bird-dispersed seed rain

respectively. These two species were also the only ones to be bird-dispersed to exposed

sites. There were also significant differences in bird-dispersed seed densities among the

six study trees (Figure 2.2). In many cases, it was difficult to assess whether conspecific

seeds collected under parent trees were dispersed by birds from other conspecific trees, or

25

merely fell into seed traps. Numbers of bird-dispersed seeds in the seed rain are therefore

a conservative estimate.

A total of 14 species with fleshy-fruited seeds were collected in the seed rain, 11

that had been dispersed by birds and three (Pleomele auwahiensis, Reynoldsia

sandwicensis and Santalum ellipticum) that had not. Of these fleshy-fruited species,

98.7% were collected under trees, and 1.3% were collected in exposed sites. Fleshy-

fruited seed rain was significantly higher under trees versus exposed areas in all cases

(Table 2.4). Bocconia frutescens comprised 66.6% of the fleshy-fruited seed rain,

whereas all native species combined comprised only 26.6%. Bocconia seeds were also

the most widespread, collected under 93.3% of trees and 5.6% of exposed sites (Table

2.6). In contrast, all native species combined were collected under only 61.1% of trees

and in 0% of exposed sites (Table 2.6). The second most widespread species in the

fleshy-fruited seed rain was Lantana camara, collected under 74.4% of all trees and in

11.1% of all exposed sites (Table 2.6). Seeds collected beneath conspecific fruiting trees

likely fell directly from the parent tree into seed traps. This conspecific seed rain was

highest under Bocconia trees, but did not significantly differ among the four native trees

(Figure 2.3). In comparisons of Bocconia seed rain under native trees versus conspecific

native seed rain, Bocconia seed densities did not differ from conspecific seed rain under

three native trees, but there were significantly more Bocconia than Santalum seeds under

Santalum ellipticum trees (Figure 2.4).

Seedlings of only eight fleshy-fruited species were present under study trees

(Table 2.7). Bocconia seedlings were the most widespread, occurring under 84% of all

trees and present under all six study species. Diospyros seedlings were present under

26

Diospyros trees, but were not recorded under other species. Reynoldsia seedlings were

present under 13% of Reynoldsia trees, but were not found under any other species. No

Pleomele or Santalum seedlings were noted under any trees. Lantana camara seedlings

were the second most widespread species, occurring under 72% of all trees. Two native

fleshy-fruited shrubs, Osteomeles anthyllidifolia and Wikstroemia monticola, were also

relatively widespread, occurring under 22% and 32% of all trees respectively (Table 2.7).

Table 2.4. Mean density (seeds m-2 ± 1 SE) of seeds dispersed by birds and fleshy-fruited seed rain under trees and in associated exposed sites for six dry forest tree species in KNAR. All pairwise comparisons between trees and exposed sites are significantly different (paired t-test; P < 0.05).

Bird- dispersed Fleshy- fruited Tree Species Tree Exposed Tree Exposed Bocconia frutescens 105.3(35.8) 6.58(6.58) 1886.7(494.3) 6.58(6.58) Dead 202.6(48.7) 1.32(1.32) 203.9(48.4) 2.63(1.79) Diospyros sandwicensis 442.1(131.6) 21.05(12.25) 752.6(177.2) 22.37(12.16)Pleomele auwahiensis 165.8(55.9) 15.79(10.58) 227.6(61.9) 23.68(11.59)Reynoldsia sandwicensis 263.1(50.0) 1.32(1.32) 840.7(313.2) 1.32(1.32) Santalum ellipticum 248.7(59.8) 0.00(0.00) 347.3(76.5) 1.32(1.32)

27

Table 2.5. Mean density (seeds m-2) of seeds dispersed by birds for 11 species under six dry forest tree species in KNAR. Boc fru: Bocconia frutescens; Dio san: Diospyros sandwicensis; Ple auw: Pleomele auwahiensis; Rey san: Reynoldsia sandwicensis; San ell: Santalum ellipticum. Non-native species are marked with an asterisk (*).

Seed species Boc fru Dead Dio san Ple auw Rey san San ell Alyxia oliviformis 0.0 0.0 1.3 0.0 0.0 0.0 Bocconia frutescens* 28.9 114.5 388.1 127.6 177.6 227.6 Cocculus orbiculatus 0.0 0.0 4.0 5.3 0.0 1.3 Diospyros sandwicensis 0.0 0.0 0.0 0.0 2.6 0.0 Lantana camara* 50.0 48.7 35.5 23.7 75.0 17.1 Leptecophylla tameiameiae 0.0 0.0 5.3 0.0 1.3 0.0 Myoporum sandwicense 0.0 1.3 0.0 0.0 0.0 0.0 Nothocestrum latifolium 21.1 19.7 0.0 0.0 2.6 0.0 Osteomeles anthyllidifolia 4.0 11.8 5.3 6.6 4.0 0.0 Schinus terebinthifolius* 0.0 0.0 0.0 0.0 0.0 2.6 Wikstroemia monticola 1.3 6.6 2.6 2.6 0.0 0.0

28

Table 2.6. Relative frequency of fleshy-fruited species in the seed rain under six dry forest tree species and all exposed (Exp.) seed traps at KNAR. Boc fru = Bocconia frutescens; Dio san = Diospyros sandwicensis; Ple auw = Pleomele auwahiensis; Rey san = Reynoldsia sandwicensis; San ell = Santalum ellipticum. Non-native species are marked with an asterisk (*).

Seed Species Boc fru Dead Dio

san Ple auw

Rey san

San ell Exp. All

TreesNon-natives 100.0 100.0 100.0 93.3 100.0 86.7 15.6 96.7Natives 33.3 53.3 100.0 53.3 80.0 46.7 0.0 61.1 Alyxia oliviformis 0.0 0.0 6.7 0.0 0.0 0.0 0.0 1.1Bocconia frutescens* 100.0 93.3 100.0 86.7 93.3 86.7 5.6 93.3Cocculus orbiculatus 0.0 0.0 13.3 6.7 0.0 6.7 0.0 4.4Diospyros sandwicensis 0.0 0.0 100.0 0.0 13.3 0.0 0.0 18.9Lantana camara* 73.3 86.7 66.7 73.3 93.3 53.3 11.1 74.4Leptecophylla tameiameiae 0.0 0.0 20.0 0.0 6.7 0.0 0.0 4.4Myoporum sandwicense 0.0 6.7 0.0 0.0 0.0 0.0 0.0 1.1Nothocestrum latifolium 6.7 13.3 0.0 0.0 6.7 0.0 0.0 4.4Osteomeles anthyllidifolia 13.3 33.3 20.0 20.0 13.3 0.0 0.0 16.7Pleomele auwahiensis 0.0 0.0 0.0 46.7 0.0 0.0 0.0 7.8Reynoldsia sandwicensis 0.0 0.0 0.0 0.0 80.0 0.0 0.0 13.3Santalum ellipticum 0.0 0.0 0.0 0.0 0.0 46.7 0.0 7.8Schinus terebinthifolius* 0.0 0.0 0.0 0.0 0.0 13.3 0.0 2.2Wikstroemia monticola 13.3 20.0 6.7 6.7 0.0 0.0 0.0 7.8

29

Figure 2.2. Mean density of seeds dispersed by birds under six dry forest tree species at KNAR. Boc fru = Bocconia frutescens; Dio san = Diospyros sandwicensis; Ple auw = Pleomele auwahiensis; Rey san = Reynoldsia sandwicensis; San ell = Santalum ellipticum. Letters at the base of bars show the results of a one-way ANOVA and Tukey test. Trees sharing the same letter are not significantly different (P > 0.05).

30

Figure 2.3. Mean density of conspecific seeds under five dry forest tree species at KNAR. Boc fru = Bocconia frutescens; Dio san = Diospyros sandwicensis; Ple auw = Pleomele auwahiensis; Rey san = Reynoldsia sandwicensis; San ell = Santalum ellipticum. Letters at the base of bars show the results of a one-way ANOVA and Tukey test. Trees sharing the same letter are not significantly different (P > 0.05).

31

Figure 2.4. Mean density of Bocconia seeds versus conspecific seeds under four dry forest tree species at KNAR. Letters at the base of bars show the results of paired t-tests for Bocconia seeds and conspecific seeds under each tree species. Densities were square root transformed for analyses. Bars sharing the same letter are not significantly different (P > 0.05). Most conspecific seeds likely fell directly from the parent tree into seed traps.

32

Table 2.7. Relative frequency of seedlings of fleshy-fruited species under six dry forest tree species at KNAR. Boc fru = Bocconia frutescens; Dio san = Diospyros sandwicensis; Ple auw = Pleomele auwahiensis; Rey san = Reynoldsia sandwicensis; San ell = Santalum ellipticum. Non-native species are marked with an asterisk (*). (n = 15 for individual species; n = 90 for all trees).

Seedling Species Boc fru Dead Dio san Ple auw Rey san San ell All TreesAlyxia oliviformis 0.0 13.3 0.0 0.0 0.0 6.7 3.3 Bocconia frutescens* 93.3 93.3 93.3 53.3 80.0 93.3 84.4 Diospyros sandwicensis 0.0 0.0 20.0 0.0 0.0 0.0 3.3 Lantana camara* 80.0 86.7 86.7 73.3 46.7 60.0 72.2 Leptecophylla tameiameiae 0.0 0.0 0.0 0.0 0.0 0.0 0.0 Myoporum sandwicense 0.0 0.0 0.0 6.7 0.0 0.0 1.1 Nothocestrum latifolium 0.0 0.0 0.0 0.0 0.0 0.0 0.0 Osteomeles anthyllidifolia 13.3 26.7 53.3 20.0 0.0 20.0 22.2 Pleomele auwahiensis 0.0 0.0 0.0 0.0 0.0 0.0 0.0 Psydrax odorata 0.0 0.0 13.3 0.0 0.0 0.0 2.2 Reynoldsia sandwicensis 0.0 0.0 0.0 0.0 13.3 0.0 2.2 Santalum ellipticum 0.0 0.0 0.0 0.0 0.0 0.0 0.0 Schinus terebinthifolius* 0.0 0.0 0.0 0.0 0.0 0.0 0.0 Wikstroemia monticola 6.7 20.0 73.3 33.3 20.0 40.0 32.2

Bird Observations

A total of 162 hours were spent observing bird visits to the five species of fleshy-

fruited study trees. Birds were observed feeding on the fruits or seeds of four of the five

species, but no birds were observed feeding on any Pleomele auwahiensis fruits or seeds

(Table 2.8). Bocconia frutescens had the greatest number of bird foraging visits per hour

and the highest number of seeds consumed per visit (Table 2.8).

A total of seven species of non-native birds visited the five study tree species

during observation hours, but only three were observed feeding on fruits or seeds (Table

2.8). Japanese white-eyes (Zosterops japonicus) were the most frequent visitor to all five

tree species, but only fed on the fruits of three (Table 2.8). These birds generally 33

swallowed entire Bocconia seeds and arils, but on several occasions were observed

passing seeds from one to another. In 35 of 39 (89.7%) foraging visits to Reynoldsia

fruits, white-eyes only pecked at pulp and dislodged berries from panicles, apparently

without swallowing seeds. White-eyes were also only observed pecking at the pulp of

Diospyros fruits, although in one instance a bird carried the fruit a short distance to a

nearby tree before consuming the fruit pulp. Northern cardinals (Cardinalis cardinalis)

visited all five tree species and swallowed whole fruits and seeds of two (Bocconia

frutescens and Reynoldsia sandwicensis). On four occasions, cardinals were observed

clipping the unripe fruits of Santalum ellipticum in half and consuming the seed inside.

They were never observed consuming only pulp of ripe fruits. Northern mockingbirds

(Mimus polyglottos) visited all five tree species and swallowed whole fruits and seeds of

two (B. frutescens and R. sandwicensis).

34

Table 2.8. Mean visits per hour and mean fruits consumed (± SE) per foraging visit by seven non-native bird species to five dry forest tree species at KNAR. Boc fru: Bocconia frutescens; Rey san: Reynoldsia sandwicensis; San ell: Santalum ellipticum; Dio san: Diospyros sandwicensis; Ple auw: Pleomele auwahiensis.

Boc fru Rey san San ell Dio san Ple auw Observation days 10 9 7 6 8 Observation hours 40 30 30 30 32 Visits per hour by bird spp. Acridotheres tristis 0.00 0.10 0.07 0.00 0.00 Cardinalis cardinalis 0.23 0.17 0.17 0.03 0.03 Carpodacus mexicanus 0.00 0.00 0.03 0.03 0.63 Lonchura punctulata 0.03 0.00 0.07 0.00 0.13 Mimus polyglottos 0.08 0.07 0.03 0.03 0.03 Streptopelia chinensis 0.03 0.00 0.03 0.00 0.00 Zosterops japonicus 5.75 3.86 1.97 1.37 0.97 Total visits per/hour 6.10 (0.83) 4.19 (0.94) 2.37 (0.37) 1.46 (0.40) 1.78 (0.40)Total foraging visits/hour 3.13 (0.46) 1.40 (0.40) 0.13 (0.06) 0.10 (0.06) 0.00 (0.00) Fruits consumed per visit Acridotheres tristis 0.00 0.00 0.00 0.00 0.00 Cardinalis cardinalis 6.60 (3.39) 1.00 (0.00) 1.50 (0.29)b 0.00 0.00 Carpodacus mexicanus 0.00 0.00 0.00 0.00 0.00 Lonchura punctulata 0.00 0.00 0.00 0.00 0.00 Mimus polyglottos 5.00 (0.00) 3.00 (1.00) 0.00 0.00 0.00 Streptopelia chinensis 0.00 0.00 0.00 0.00 0.00 Zosterops japonicus 2.03 (0.12) 1.18 (0.13)a 0.00 0.67 (0.33)c 0.00 Total consumed/foraging visit 2.26 (0.19) 1.26 (0.14) 1.50 (0.29) 0.67 (0.33) 0.00 (0.00)amostly pecked at pulp bseeds depredated conly pulp consumed

Seed Size and Bird Dispersal

Of the 11 bird-dispersed species collected in the seed rain at KNAR (Table 2.9),

there was a significant negative relationship between numbers of bird-dispersed seeds

35

and seed width (rs = -0.699, P = 0.015) but not seed length (rs = -0.374, P = 0.245). When

the three additional fleshy-fruited species present in the seed rain were included in the

analysis, there were significant negative relationships between seed length (rs = -0.586, P

= 0.026), seed width (rs = -0.686, P = 0.006) and numbers of seeds dispersed by birds.

Table 2.9. Mean (± 1 SE) seed length and width of 14 seeds collected in the seed rain at KNAR. Sample sizes are for seed measurements.

Seed species Greatest length

without pulp (mm)

Greatest width without pulp

(mm) n # bird-

dispersed

Bocconia frutescens 4.08 (0.03) 2.71 (0.02) 100 809 Lantana camara 4.27 (0.06) 2.64 (0.04) 59 190 Nothocestrum latifolium 2.79 (0.04) 1.45 (0.13) 33 33 Osteomeles anthyllidifolia 5.28 (0.12) 2.92 (0.10) 16 24 Wikstroemia monticola 5.48 (0.32) 2.73 (0.10) 9 10 Cocculus orbiculatus 5.01 (0.28) 3.21 (0.23) 7 8 Leptecophylla tameiameiae 2.88 (0.05) 2.40 (0.11) 6 5 Schinus terebinthifolius 3.90 (0.00) 3.40 (0.00) 1 2 Diospyros sandwicensis 13.20 (1.90) 6.90 (0.09) 100 2 Myoporum sandwicense 4.90 (0.00) 3.20 (0.00) 1 1 Alyxia oliviformis 6.80 (0.00) 6.20 (0.00) 1 1 Reynoldsia sandwicensis 5.34 (0.05) 2.92 (0.03) 100 0 Pleomele auwahiensis 7.62 (0.06) 6.15 (0.06) 100 0 Santalum ellipticum 7.88 (0.05) 6.96 (0.04) 100 0

DISCUSSION

The results of this study demonstrate the importance of trees as a factor

influencing the distribution of bird-dispersed seeds. The much higher densities and

greater species richness of bird-dispersed seeds under all trees contrasts with the relative

absence of these seeds in exposed sites. The lack of suitable perches appears to limit

36

deposition of bird-dispersed seeds into exposed areas, as birds tend to concentrate seed

rain of fleshy-fruited species under trees or artificial perches of similar structure

(McDonnell and Stiles 1983; Guevara and Laborde 1993; Ferguson and Drake 1999;

Jordano and Schupp 2000; Shiels and Walker 2003). The significantly higher seed rain

and species richness of bird-dispersed seeds under dead standing trees versus exposed

sites further emphasizes this fact (Table 2.4). Data collection under fleshy-fruited species

in this study could also increase seed deposition under trees, as other studies have

documented higher bird-dispersed seed rain under fruiting versus non-fruiting trees

(Slocum and Horvitz 2000; Clark et al. 2004). Although this may be true in some cases,

the bird-dispersed seed rain under dead trees did not differ from any of the fleshy-fruited

tree species (Figure 2.2). Nevertheless, comparisons of bird-dispersed seed rain during

the fruiting seasons of fleshy-fruited trees may detect differences between fruiting and

non-fruiting trees that were not apparent in the 12-month analysis.

The higher densities of fleshy-fruited seeds under trees versus exposed sites,

whether bird-dispersed or not, conform to expectations of a leptokurtic seed shadow, with

greatest densities occurring close to the seed source (Willson and Traveset 2001).

Disperser limitation may exaggerate this pattern, as three of the species used in this study,

(Pleomele auwahiensis, Reynoldsia sandwicensis, and Santalum ellipticum), were not

recorded in the bird-dispersed seed rain and another, (Diospyros sandwicensis), was

relatively scarce (Table 2.5). In contrast, densities of wind-dispersed seeds did not differ

significantly between trees and exposed areas (Table 2.1). This could be partly due to the

relative abundance of wind-adapted grasses at the study site (Table D.1, Appendix D),

but the lack of dependence on animal vectors for dissemination must also play an

37

influential role. Of the 11 bird-dispersed species recorded in this study, two fleshy-fruited

non-natives, Bocconia frutescens and Lantana camara, were the most abundant and

widespread in the bird-dispersed seed rain (Table 2.5-2.6). Several factors could account

for this dominance. Lantana camara is the most common shrub in the study site (Table

D.1, Appendix D), and B. frutescens is the most common non-native tree (Table D.2,

Appendix D). Nevertheless, all native fleshy-fruited trees combined are more abundant

than Bocconia alone (Tables D.2-D.3, Appendix D), so additional factors must account

for the latter’s higher seed rain. Lantana and Bocconia may produce larger fruit crops of

relatively small-sized seeds that, by sheer number, have a greater probability of being

deposited in seed traps, or that are removed at higher rates by frugivores attracted to the

abundant food source (Howe and Estabrook 1977; Izhaki 2002). Although fruit crop sizes

were not estimated for the different trees in this study, comparisons of conspecific seed

rain under trees indicate that Bocconia seeds fall at significantly higher densities under

parent trees than do any of the native species (Figure 2.3). In addition, similar or greater

densities of Bocconia seeds versus conspecific seeds (i.e. seeds of same species as

overhanging tree) were collected under the four native fleshy-fruited trees (Figure 2.4).

These results suggest that fruit crops of Bocconia are much larger than those of the four

native study trees, despite the fact that Bocconia trees are of relatively equal or smaller

stature than the natives (Figure A.1, Appendix A).

Fruiting phenology of trees may also account for the disparities in the bird-

dispersed seed rain (Carlo et al. 2003). Frugivores have been shown to track changes in

availability of important food supplies (Loiselle and Blake 1991; Price 2004), and may

focus on particular species that produce fruit when most others are sterile (Howe and

38

Primack 1975). The large fruit crop of Bocconia trees, in particular, which peaked during

the summer months of June through August, may have been one of the few foods

available to frugivores at a time of year when other fruit resources were relatively scarce

(Table B.1, Figures B.1 and B.3, Appendix B). This could be a result of annual variation,

however, so longer-term comparisons of bird-dispersed seed rain, fruiting phenology and

frugivore abundance would be useful in determining to what degree fruiting seasonality

influences amounts and rates of fruit removal and dispersal of particular fleshy-fruited

species.

Because many of the extinctions of the avifauna in the Hawaiian Islands and

throughout Polynesia occurred during the early periods of human colonization (Olson and

James 1982, 1991; Steadman 1995; Burney et al. 2001; Ziegler 2002), it is impossible to

determine conclusively all of the community level effects these extinctions have had on

plant demographics. Nevertheless, the abundance of fleshy-fruited taxa in the Hawaiian

Islands and similar insular ecosystems such as New Zealand suggests that some of these

extinct birds would have been attracted to and likely played an important role in the

dispersal and potential scarification of fleshy-fruited species (Clout and Hay 1989;

Wagner et al. 1999). Because isolated islands tend to have more depauperate assemblages

of dispersers than continental areas do, islands may be even more dependent on the

limited dispersal pool for the preservation of floral biodiversity (Bleher and Böhning-

Gaese 2001; Cox et al. 1991; McConkey and Drake 2002). When native species do go

extinct, dispersal failure could result, or the original native dispersers may be replaced by

other natives or newly introduced species that serve as ecological substitutes for

processes such as seed dispersal (Loiselle and Blake 2002; Hampe 2003).

39

In the Hawaiian Islands, the almost complete replacement of the native avifauna

with a suite of non-native generalist frugivores and game birds at lower elevations

provides an opportunity to explore the role that non-natives do play as potential

surrogates in seed dispersal and other ecological processes (Pratt 1997). Bird

observations on the four fleshy-fruited native study tree species and one non-native

species, although admittedly limited in scope, do provide insights into whether or not

ecological substitutions are occurring. Of the 16 non-native bird species documented in

KNAR (Medeiros et al. 1993), 12 were observed at some point during the 12-month

duration of this study (Table C.1, Appendix C), but only nine consume fruit pulp as a

regular part of their diet (Scott et al. 1986). Of these nine, only three were observed

visiting and consuming the pulp and/or seeds of some of the study trees. The most

abundant species at the study site was the Japanese white-eye, a now ubiquitous bird first

introduced to the islands in 1929 (Appendix C; Scott et al. 1986). Although this bird was

regularly observed swallowing the seeds and attached arils of Bocconia frutescens, it

predominantly pecked at the pulp of the two native species it also visited (Table 2.8).

Many of the berries of Reynoldsia sandwicensis were dislodged from the fruiting panicle

as white-eyes pecked at the pulp, causing them to fall directly under the parent canopy.

White-eyes are apparently dispersing Bocconia seeds and in limited instances may

disperse the seeds of native trees, but in general, species that do not provide the benefit of

seed dispersal when feeding on fleshy fruits have been referred to as “fruit predators”

(Herrera 1995). Of the other two non-native species observed feeding on fruits, both the

northern cardinal and the northern mockingbird swallowed whole fruits and seeds of both

B. frutescens and R. sandwicensis (Table 2.8). Either could be considered legitimate

40

dispersers if they pass viable seeds of these or other trees (Jordano 2001). Although both

species were among the more abundant recorded in KNAR, both are relatively

uncommon compared to the Japanese white-eye, and their roles as dispersal agents may

be limited (Table C1 and C2, Appendix C). In addition, rather than dispersing seeds of

Santalum ellipticum, northern cardinals act as pre-dispersal seed predators by clipping the

immature fruits in half to feed on seeds. This behavior, combined with the impacts of

post-dispersal seed predation by introduced rodents (Chapter 3), could severely limit or

entirely prevent the recruitment of this endemic species.

41

Of the other plant species recorded in the bird-dispersed seed rain (Table 2.5),

Japanese white-eyes were anecdotally observed feeding on the fruits of four (Lantana

camara, Nothocestrum latifolium, Osteomeles anthyllidifolia and Wikstroemia

monticola). Dispersal of these and other fleshy-fruited taxa, in addition to being related to

fruiting phenology, plant abundance and fruit crop size, may also be influenced by seed

size. Both fruit and seed size are considered to be important factors affecting the type and

number of frugivores capable of consuming fruits and potentially dispersing seeds (Snow

1981; Wheelwright 1985; Levey 1987). Smaller frugivores, limited by body size and

gape width, may in turn influence the evolution of fruit and seed size by selectively

feeding on relatively smaller fruits or seeds (Lord 2004). In KNAR, the most abundant

generalist frugivore, the Japanese white-eye, is a fairly small bird of between 11-12 cm

body length, 9.75-12.75 g weight, and gape width of between 5-8 mm (Corlett 1986; Van

Riper 2000). In Australia and New Zealand, the silvereye (Z. lateralis), a related species

of similar size and gape, does consume the pulp of fruits exceeding its gape width

(Williams and Karl 1996), and prefers relatively larger over smaller fruits, up to a certain

point (Stanley et al. 2002; Stansbury and Vivian-Smith 2003). Neither of these studies

indicates whether silvereyes are capable of dispersing the seeds of these larger fruit, but

in general, most birds have a size limit over which the handling costs outweigh the

benefits of fruit and seed consumption (Martin 1985; Levey 1987). In this study, there

was a negative relationship between seed width and number of seeds dispersed, and three

of the larger-seeded species were rarely, or never, consumed by birds or collected in the

bird-dispersed seed rain (Table 2.8-2.9). Most of the bird-dispersed seed sizes fell within

or below the recorded gape width of the Japanese white-eye, and presumably also the

much larger northern cardinal and mockingbird (Table 2.9). Therefore, although further

study is necessary to determine whether seed size limits the dispersal of larger-seeded

species, the relatively smaller sizes of the abundantly dispersed Bocconia and Lantana

seeds certainly did not hinder their consumption and dissemination.

42

Because bird-dispersed seeds in forested areas have heterogeneous spatial

distributions (Dalling et al. 1998; Ferguson and Drake 1999), it is possible that several

species that birds had dispersed were missed by chance and were therefore not

represented in the documented seed rain. Nevertheless, it is clear that a large proportion

of the fleshy-fruited seeds from the study species are not being dispersed and are instead

falling directly under parent trees. Several negative consequences could result from this

lack of dispersal. Seeds occurring at higher densities under parent trees may be subjected

to greater levels of predation (Howe et al. 1985; Chapman and Chapman 1996; Hulme

1997; Wenny 2000; Holl 2002). The presence of fruit pulp on seeds that are not

dispersed may even aid seed predators in finding and consuming seeds (Moles and Drake

1999, Nystrand and Granstrom 1997, Chapter 3). For those seeds that escape predation,

germination may be reduced due to the presence of pulp on seeds or lack of seed

scarification by birds or other dispersal agents (Fukui 1995, Yagihashi et al. 1998;

Traveset and Verdú 2002). Seedlings that germinate under parent trees may experience

higher rates of mortality due to density-dependent seedling predation, seedling

competition or greater exposure to seedling pathogens (Augspurger 1984; Chapman and

Chapman 1994; Kitamura et al. 2004). Finally, restoration of degraded areas may be

limited by lack of dispersal of seeds into these sites (Holl et al. 2000; Cordeiro and Howe

2001; Howe and Miriti 2004; Makana and Thomas 2004). Many of these factors may be

contributing to the relatively low frequency of seedlings of certain native fleshy-fruited

species under parent trees (Table 2.8), and, combined with other factors such as lack of

outcrossing and browsing and trampling by feral ungulates, may be hastening the

continued decline of dry forests at KNAR and throughout the Hawaiian Islands.

43

Plants that benefit from directed dispersal (e.g. seed dispersal by birds under

trees) tend to increase in abundance (Wenny 2001). For those species benefiting from the

effects of directed dispersal by the non-native frugivores at KNAR, seedling recruitment

under trees does not appear to be a problem (Table 2.8). Although the effects of dispersal

may prove beneficial to the recruitment and recovery of native species, directed dispersal

may also facilitate the spread of invasive, habitat modifying weeds (Richardson et al.

2000; Cordeiro et al. 2004). This appears to be the case for Bocconia frutescens and

Lantana camara, the two most common fleshy-fruited non-native plants in KNAR.

Lantana camara, first introduced to the Hawaiian Islands in 1858, is now widespread

throughout the low-elevation dry and mesic habitats of all the main islands (Wagner et al.

1999). Bocconia frutescens, a more recent invader, was first collected in Kanaio, Maui in

1920, but has rapidly spread throughout KNAR and the leeward slopes of the eastern half

of the island (Medeiros et al. 1993; Wagner et al. 1999; Appendix D). Habitat shaping,

the result of disperser and plant interactions, occurs when seed dispersal contributes to

plant composition and abundance over an area (Herrera 1995). This seed dispersal by

frugivorous birds can also lead to the homogenization of plant distributions over an entire

region (Debussche et al. 1982). The relative abundance of Bocconia and Lantana plants

in the study site, the frequency and abundance of their seeds in the seed rain under all

study species, the frequency of seedlings under and away from parent trees, and the

frequent interaction with Japanese white-eyes, all suggest that just such a homogenization

is occurring at KNAR and adjacent dry forest areas, to the likely detriment of the

remaining native species.

Conclusion

The much higher seed densities of fleshy-fruited and bird-dispersed seeds under

trees versus exposed sites emphasizes the importance of trees as dispersal and potential

recruitment foci for seeds and seedlings. In degraded ecosystems such as those of the

remaining Hawaiian dry forests, these sites may facilitate restoration and recovery

processes when desirable species are present in the seed rain (Otero-Arnaiz et al. 1999;

Holl et al. 2000; Guevara et al. 2004). Despite the loss of native dispersal agents, non-

native species will often serve as ecological substitutes and fill the roles left vacant by

extinct dispersers (Loiselle and Blake 2002). Although not investigated in this study, non-

native game birds may be scarifying and secondarily dispersing both native and non-

native seeds, as they have done in other Hawaiian ecosystems (Cole et al. 1995). Aside

44

from a few small-seeded natives, however, the majority of fleshy-fruited species being

dispersed by non-native frugivores at KNAR are relatively small-seeded non-native trees

and shrubs. Whether or not seed size and gape width limitations are contributing to this

disparity, it is clear that bird perches and other structures that enhance seed dispersal into

degraded habitats will only facilitate further invasion by these non-native species. New

introductions of larger frugivores could increase the number of large-seeded natives

being dispersed, but until the seed sources of invasive species are eliminated, only human

intervention will ensure the selective dispersal and perpetuation of desirable native

species over invasive non-native weeds.

45

CHAPTER 3: POST-DISPERSAL SEED PREDATION: LOCATION AND

EFFECT ON SPECIES

INTRODUCTION

The flora and fauna of Pacific island ecosystems, including the Hawaiian Islands,

evolved in the absence of native rodents, and presumably have fewer defenses against

them than species that evolved with rodents. Introduced rodents such as the Polynesian

rat (Rattus exulans), the black or roof rat (R. rattus), the Norway rat (R. norvegicus) and

the house mouse (Mus musculus) have been introduced to many islands throughout the

world, with devastating impacts on the insular biota (Tomich 1986). In the Hawaiian

Islands, rats have negatively impacted the native avifauna (Atkinson 1977; Scott et al.

1986), the endemic snail and arthropod fauna (Hadfield et al. 1993; Cole et al. 2000) and

the flora, including many rare and endangered species (Wirtz 1972; Baker and Allen

1978; Russel 1980; Scowcroft and Sakai 1984; Stone et al. 1984; Stone 1985; Sugihara

1997; Cole et al. 2000). Although introduced rodents are notorious plant and seed

predators in other insular ecosystems (Fall et al. 1971; Clark 1982; Campbell et al. 1984;

Allen et al. 1994; Moles and Drake 1999; Williams et al. 2000; Campbell and Atkinson

2002; Delgado Garcia 2002; McConkey et al. 2003), and have been anecdotally

implicated in the destruction of seeds in Hawaiian dry forests, (Medeiros et al. 1986;

Medeiros et al. 1993; Cabin et al. 2000), no studies have been published that

quantitatively document the impacts of these introduced rodents on the seeds of particular

dry forest taxa. As many of these species produce fruits with one to few larger seeds

(Wagner et al. 1999), and as rodents in general have been shown to be voracious seed

46

predators of intermediate and larger seeds (Stiles 2001), the effects that introduced

rodents have on seed mortality and subsequent levels of seedling recruitment could be

profound (Crawley 2001).

A troubling phenomenon that seriously affects the persistence as well as the

potential future restoration of dry forest communities is the almost complete lack of

natural reproduction of native tree species (Medeiros et al. 1986; Loope et al. 1995;

Loope 1998; Cabin et al. 2000). Janzen (1988) refers to adult trees that fail to reproduce

as “the living dead” and indicates that this problem is symptomatic of tropical dry forests

worldwide. Several factors may contribute to the lack of seedling recruitment in tropical

dry forests. These include browsing and trampling by ungulates, competition with weeds,

the loss of native pollinators, the impacts of non-native invertebrates and pathogens on

seed and seedling survival, modifications in microclimate and microhabitats suitable for

seed germination and seedling survival, loss of native birds that scarified and dispersed

seeds, and seed predation by introduced rodents (Janzen 1988; Aide et al. 2000; Cabin et

al. 2000; Holl et al. 2000; Zimmerman et al. 2000). The effect of rodent predation on

seeds is one factor that could be influencing community composition and structure of

Hawaiian dry forests.

The scattered spatial distribution of dry forest trees could also influence the

deposition of seeds throughout an area. At the Kanaio Natural Area Reserve and other

remnant Hawaiian dry forests, isolated trees and groves are surrounded by lower-statured

native shrublands, non-native grasslands and largely barren lava flows (Medeiros et al.

1993). Previous studies suggest that seed deposition and predation rates can differ

markedly between open habitats and under trees. Isolated trees or groves of trees in

47

otherwise open areas may serve as recruitment foci for bird-disseminated species

(McDonnell and Stiles 1983; Ferguson and Drake 1999). This dispersal results in a non-

random seed rain due to the preferential movements of birds to these feeding and perch

sites (Jordano and Schupp 2000; Clark et al. 2004). These recruitment foci also

sometimes serve as suitable microsites for the growth of seedlings and may play

important roles in the succession and regeneration of disturbed or degraded areas (Uhl

1998; Wijdeven and Kuzee 2000).

Despite the potential benefits, however, this non-random deposition under trees,

combined with the higher seed rain predicted by the seed shadow under parent trees,

could also influence seed predation rates (Hulme 1998). Several authors have

documented higher seed predation rates under trees and vegetation as opposed to more

open areas (Howe et al. 1985; Callaway 1992; Aide and Cavelier 1994; Osunkoya 1994;

Chapman and Chapman 1996; Hau 1997; Hulme 1997; Wenny 2000; Holl 2002). Others

have found no difference (DeSteven and Putz 1984; Terborgh et al. 1993), or higher rates

of predation in exposed areas or grasslands (Hay and Fuller 1981; Uhl 1998; Wijdeven

and Kuzee 2000). One study reported lower predation rates under isolated pasture trees as

opposed to either open pastures or intact forests (Holl and Lulow 1997). Seed predators

may avoid exposed sites to minimize risks from aerial predators such as owls (Murúa and

González 1982; Sánchez-Cordeiro and Martínez-Gallardo 1998), a factor that could

influence seed removal rates in Hawaiian dry forests. Lower seed densities in exposed

sites may also contribute to lower predation rates, as some seed predators may

concentrate their activity in optimal foraging areas of higher seed densities closer to the

seed source (Janzen 1971; Hulme 1998).

48

Therefore, to better understand the interaction between the distribution of dry

forest trees and seed predation in different microhabitats of the Kanaio Natural Area

Reserve, the following questions are addressed in this paper:

1) Are rodents removing and destroying seeds of dry forest species?

2) Is there a difference in removal rates of seeds under trees and in exposed

sites?

3) Does the presence of fruit pulp or fleshy arils affect rates of removal?

4) Is there a relationship between seed size and removal rates for dry forest tree

species?

METHODS

Study Site

This study was conducted in the Kanaio Natural Area Reserve (KNAR), on the

leeward side of East Maui, Hawai‘i, at an elevation between 750-850 meters (20º36’N,

156º20’W; Figure 2.1). The climate in the reserve is hot and dry, with annual

temperatures between 20-30° C and with mean annual rainfall of approximately 750 mm,

mostly falling between the months of October to April (Giambelluca et al. 1986). The

substrate is estimated to be less than 10,000 years old and consists mostly of `a`a lava

with some accumulation of overlying soils in some areas (Crandell 1983). Medeiros et al.

(1986; 1993) have provided a thorough description of the community composition of the

dry forest of KNAR and the leeward slopes of Haleakala, Maui. The 354-hectare reserve

was established in 1990 to protect exemplary representatives of Hawaiian dry forest

species and plant communities, including Dodonaea (`A`ali`i) lowland shrublands,

49

Diospyros (Lama) forest and Erythrina (Wiliwili) forests (Gagne and Cuddihy 1990). In

addition to these communities, the reserve also contains individuals and groves of at least

twenty-two native tree species. Anecdotal observations suggest that only eight of these

trees may be reproducing (Medeiros et al. 1993). Certain non-native trees, however,

including the widespread Bocconia frutescens (Papaveraceae), are able to reproduce

without the apparent difficulties experienced by native species (Chimera pers. obs.). At

least two species of non-native rodents, the black or roof rat (Rattus rattus), and the

house mouse (Mus musculus) currently occur in KNAR (Medeiros et al. 1993).

Measuring Seed Predation Levels

To examine to what degree rodents remove tree seeds in KNAR, seeds from four

native and one non-native tree species were used in a series of predation trials. Native

species were selected to represent a range of seed sizes. Native fleshy-fruited taxa used

for this study include Reynoldsia sandwicensis (Araliaceae), Diospyros sandwicensis

(Ebenaceae), Santalum ellipticum (Santalaceae), and Pleomele auwahiensis (Agavaceae),

as all are locally abundant at the site. One non-native tree with arillate seeds, Bocconia

frutescens (Papaveraceae), was also used. Mean tree heights ranged between

approximately three to six meters (Appendix A). Fifteen trees of each species were

selected from the northeastern portion of KNAR for the placement of one set of

treatments. Trees were selected along 12 parallel, east-west oriented transects (10-m

widths), ranging in length from 100 to 675 meters and spaced 50 meters apart. All

reproductively mature study trees were first recorded along the length of each transect.

50

Study trees were randomly selected from this subset of individuals such that no selected

trees were closer than 25 meters from each other.

Treatments

Four treatments were used, based on designs of Moles and Drake (1999): open

ground, a depression (11 cm x 11 cm x 1 cm deep) in the ground with a rim so that seeds

did not wash away; open pot, square black plastic flower pots (110 mm x 110 mm wide;

150 mm tall) with one side cut away and filled with soil such that the inside and outside

soil levels are the same; pot with rodent access, open pots covered with 12-mm steel

mesh with an opening (35 mm x 35 mm) large enough to allow access to rodents but

small enough to deter feeding by larger game birds; pot with rodent-proof mesh, open

pots with 12-mm mesh designed to prevent entry by rodents but not invertebrates.

Sets of four treatments were placed under the canopy of each tree species such

that a distance of at least one meter separated each treatment. These treatments were

generally shaded from direct sunlight for a good portion of daylight hours. Sets of four

treatments were also placed in exposed sites (i.e. not under trees), on a random compass

direction and at a minimum distance of five meters from the nearest tree crown. These

exposed sites consisted of barren rock, low-stature grasses, woody shrubs and annual or

perennial herbaceous species less than 50 cm tall. Treatments in exposed sites were not

placed under vegetation and were therefore exposed to direct sunlight. Experiments were

arranged in a 2x4 factorial block design for three species and a 2x5 factorial block design

for two species (Bocconia and Diospyros) with an additional treatment (see below). Each

51

tree constituted an individual block with four or five treatments (factor 1) and two

locations (under trees and in exposed sites, factor 2), per block (Zar 1999).

For each species, a pile of five seeds was placed in each of the four treatments

under each of the 15 conspecific fruiting trees and in each of the 15 exposed sites, for a

total of 600 seeds per species (300 under trees, 300 in exposed sites). Seeds were

collected from no fewer than five haphazardly-selected trees of each species as they

became available. Length and width were recorded for 100 haphazardly-selected seeds of

each of the five species used in the predation trials. Fruit pulp was removed from seeds

unless otherwise stated. Pulp was removed using water and paper towels, and seeds were

allowed to dry before being used. For two species, Diospyros sandwicensis and Bocconia

frutescens, additional piles of five intact berries for Diospyros (1-3 seeds/berry), or seeds

with arils still attached (“arillate seeds” for Bocconia) were also placed in open ground

treatments under trees and in exposed areas. For these species, removal rates were

compared between seeds and whole fruits to determine the effect of fruit pulp on

predation. For one species, Pleomele auwahiensis, a pile of five fresh berries (1-3

seeds/berry) with pulp intact was placed in each of the four treatments under a different

set of 15 trees and exposed sites, for a total of 600 fruits from this species (300 under

trees, 300 in exposed sites).

Removal of seeds and fruits was recorded over five 15-day periods in 2003,

commencing on 2 April for P. auwahiensis seeds and fruits, 4 April for S. ellipticum, 15

April for R. sandwicensis, 29 April for D. sandwicensis, and 4 June for B. frutescens. In

all trials, seeds were used within a few days after collection to mimic the removal of

seeds when they would normally be available to predators. Seeds were categorized as

52

removed if they were taken from the treatment area or if they had been destroyed at the

treatment site. It is possible that some of the removed seeds may have actually been

dispersed by rodents or game birds, and were not necessarily consumed. Measures of

seed removal may therefore overestimate levels of predation.

Analysis

To determine whether predation levels under parent trees differ from those in

exposed sites away from the parent tree, numbers of seeds remaining in the sets of

treatments under trees were compared with those in exposed sites. This analysis

investigated differences between removal levels under trees of each species versus

associated exposed areas, but did not compare levels between different species.

Differences between combinations of treatment and location (tree or exposed) were

compared using a general linear model, with mean number of seeds remaining as the

dependent variable, individual tree as the blocking or random factor, and treatment and

location as explanatory variables with one interaction term: treatment x location. Tukey

tests were used to determine which treatments were significantly different (Zar 1999;

Ryan and Joiner 2001). Because removal trials for Pleomele auwahiensis seeds and fruits

were conducted at the same time, mean numbers of seeds and fruits remaining in open

ground treatments were further compared between locations using a two-way analysis of

variance, with a Tukey test used to identify differences among treatments.

53

Spearman’s rank correlation was used to test for relationships between seed

length, width and the mean number of cleaned seeds remaining in open treatments after

15 days, both under trees and in exposed sites.

RESULTS

Seed removal rates after 15 days varied greatly per species and by location (Table

3.1; Figure 3.1), ranging from 100% of Santalum seeds under trees to 0% of Reynoldsia

seeds under trees or in exposed sites. Rodent tooth marks, droppings and seed husks

suggest that rodents are responsible for at least some of the seed removal recorded.

Bocconia and Pleomele seed husks remained in many of the treatments. Rodent bite

marks and partially consumed seeds were present in several of the Diospyros treatments.

Seed husks and fragments were also present in about half of the accessible Santalum

treatments in which seeds were removed.

There were significant differences in removal rates between seeds and fruits set

out in different treatments and in different locations (Table 3.2; Figures 3.2-3.5). For

Bocconia frutescens, there were significant differences in removal rates between

treatments, but not between locations (Table 3.2; Figure 3.2). Significantly more arillate

seeds were removed than cleaned seeds (Figure 3.2). Some of the remaining arillate seeds

had arils removed with no apparent damage to seeds. For Diospyros sandwicensis,

significant differences in removal rates among treatments were dependent on location,

and overall removal rates were higher under trees (Table 3.2; Figure 3.2). Significantly

more fruits than seeds were removed from trees and exposed sites (Figure 3.2). For

Pleomele auwahiensis, significantly more seeds were removed from trees than exposed

54

sites, but removal rates for fruits were not significantly different between locations (Table

3.2; Figure 3.3). Significantly more Pleomele fruits were removed than seeds, but

location was not significant (Table 3.3; Figure 3.4). For Reynoldsia sandwicensis, no

seeds were removed from any of the treatments during the 15-day duration of the trial

(Figure 3.5). For Santalum ellipticum, significantly more seeds were removed from trees

than exposed sites (Table 3.2; Figure 3.5). No seeds or fruits of any species were

removed from rodent-proof pots at any point during the trials, and no signs of

invertebrate damage were noted during the 15-day study duration.

No significant relationships were found between the number of cleaned seeds

remaining in open ground treatments after 15 days under trees or exposed sites and seed

length (P=1.0), or seed width (P=0.517)

Table 3.1. Means (± 1 SE) of seed length and width and mean removal rate under trees and in exposed sites for species used in the seed predation trials. Percentages represent open ground treatments only. n = 100 for seed measurements.

Species

Greatest length

without pulp (mm)

Greatest width

without pulp (mm)

% of cleaned seeds

removed after 15 days (tree)

% of cleaned seeds

removed after 15 days

(exposed) Bocconia frutescens 4.08 (0.03) 2.71 (0.02) 77.33 64.00 Diospyros sandwicensis 12.44 (0.15) 4.98 (0.09) 49.33 2.67 Pleomele auwahiensis 7.62 (0.06) 6.15 (0.06) 66.67 56.00 Reynoldsia sandwicensis 5.34 (0.05) 2.92 (0.03) 0.00 0.00 Santalum ellipticum 7.88 (0.05) 6.96 (0.04) 100.00 81.33

55

Table 3.2. General linear models for number of seeds and/or fruits remaining versus block, treatment, location (tree or exposed), and the interaction of treatment (Trt) and location (Loc) after 15 days for four dry forest trees in Kanaio Natural Area Reserve. Tests were not performed on Reynoldsia, as no seeds were removed from any treatments.

Species Source of variation d.f. SS

(model III) MS F P

Bocconia frutescens Block 14 47.76 3.41 2.58 0.003 Treatment 4 481.76 120.44 90.92 0.000 Location 1 3.23 3.23 2.44 0.121 Trt x Loc 4 4.107 1.03 0.78 0.543 Error 126 166.91 1.33 Total 149 703.76 Diospyros sandwicensis Block 14 22.17 1.58 1.03 0.427 Treatment 4 380.17 95.04 61.91 0.000 Location 1 42.67 42.67 27.79 0.000 Trt x Loc 4 25.13 6.28 4.09 0.004 Error 126 193.43 1.54 Total 149 663.57 Pleomele auwahiensis Seeds Block 14 116.12 8.29 3.30 0.000 Treatment 3 144.03 48.01 19.11 0.000 Location 1 28.033 28.03 11.16 0.001 Trt x Loc 3 16.03 5.34 2.13 0.102 Error 98 246.15 2.51 Total 119 550.37 Pleomele auwahiensis Fruits Block 14 110.05 7.86 3.34 0.000 Treatment 3 336.17 112.06 47.65 0.000 Location 1 4.03 4.03 1.71 0.193 Trt x Loc 3 8.57 2.86 1.21 0.309 Error 98 230.48 2.35 Total 119 689.30 Santalum ellipticum Block 14 74.97 5.36 4.33 0.000 Treatment 3 440.57 146.86 118.78 0.000 Location 1 14.70 14.70 11.89 0.001 Trt x Loc 3 6.57 2.19 1.77 0.158 Error 98 121.17 1.24 Total 119 657.97

56

Figure 3.1. Removal rates of seeds and fruits from the Kanaio Natural Area Reserve. Percentage of seeds or fruits remaining ± 1 SEM for open ground treatments under trees and in exposed sites (n = 15 trees * 5 seeds per treatment). = cleaned seeds under trees; = cleaned seeds in exposed sites; = seeds under trees (with pulp); = seeds in exposed sites (with pulp). Boc fru = Bocconia frutescens; Dio san = Diospyros sandwicensis; Ple auw = Pleomele auwahiensis; Rey san = Reynoldsia sandwicensis; San ell = Santalum ellipticum.

57

Figure 3.2. Removal rates for Bocconia frutescens (Boc fru) and Diospyros sandwicensis (Dio san) seeds and fruits under trees and in exposed sites. Treatments include G: open ground; O: open pots; RA: pots with rodent access; RP: rodent-proof pots; F: fruits on open ground (n = 15 trees * 5 seeds/fruits per treatment). Bars represent the mean + 1 SEM. Letters at the base of bars show the results of the Tukey test. Treatments sharing the same letter are not significantly different (p > 0.05). Removal rates were not compared between Bocconia and Diospyros.

58

Figure 3.3. Removal rates for Pleomele auwahiensis (Ple auw) seeds and fruits under trees and in exposed sites. Treatments include G: open ground; O: open pots; RA: pots with rodent access; RP: rodent-proof pots (n = 15 trees * 5 seeds/fruits per treatment). Bars represent the mean + 1 SEM. Letters at the base of bars show the results of the Tukey test. Treatments sharing the same letter are not significantly different (p > 0.05).

59

Figure 3.4. Removal rates for Pleomele auwahiensis seeds and fruits on the open ground under trees and in exposed sites. (n = 15 trees * 5 seeds/fruits per treatment). Bars represent the mean + 1 SEM. Letters at the base of bars show the results of the Tukey test. Treatments sharing the same letter are not significantly different (p > 0.05).

Table 3.3. Two-way ANOVA results for the effects of treatment (seed or fruit) and location (tree or exposed) on removal rates for Pleomele auwahiensis

Source of variation d.f. SS MS F P Treatment (seed/fruit) 1 22.82 22.82 5.56 0.022 Location (tree/exposed) 1 12.15 12.15 2.96 0.091 Interaction 1 2.02 2.02 0.49 0.486 Error 56 230.00 4.11 Total 59 266.98

60

Figure 3.5. Removal rates for Reynoldsia sandwicensis (Rey san) and Santalum ellipticum (San ell) seeds under trees and in exposed sites. Treatments include G: open ground; O: open pots; RA: pots with rodent access; RP: rodent-proof pots (n = 15 trees * 5 seeds per treatment). Bars represent the mean + 1 SEM. Letters at the base of bars show the results of the Tukey test. Treatments sharing the same letter are not significantly different (p > 0.05). Removal rates were not compared between Reynoldsia and Santalum.

61

DISCUSSION

The highly variable seed removal rates for the five tree species used in this study

are not surprising, as differences between species are frequently documented in the seed

predation literature (Chapman and Chapman 1996; Hulme 1998; Figueroa et al. 2002;

Jones et al. 2003). The average levels of predation on cleaned seeds for all species

combined, whether under trees (59%) or in exposed sites (41%), are comparable to

relatively high levels documented in forests throughout the world. Rates of seed

predation by rodents typically exceed 50% for many species in a number of different

ecosystems (Hulme 1998). In a forest site in Spain, about 70% of Ilex aquifolium seeds

were removed after two 2-week periods (Ramón Obeso and Fernández-Calvo 2002). In

the northern territories of Hong Kong, between 73% and 86% of seeds were removed for

four species in shrublands, and between 33% and 83% were removed for six species in

grasslands after 60 days (Hau 1997). For a tropical dry forest shrub in Mexico, 64% of

seeds were removed (Gryj and Dominguez 1996). Excluding Reynoldsia sandwicensis,

which had no seeds removed during the 15-day duration of this study (Table 3.1; Figure

3.5), average levels of seed removal were 73% under trees and 51% in exposed sites.

Relatively high levels of seed predation might be expected for Hawaiian species

that evolved in the absence of rodent seed predators (Ziegler 2002). These rates are much

higher than the 10% removal rate reported by Moles and Drake (1999) for 11 species in

New Zealand, in an island biota similarly lacking native rodents. For this study, however,

evolutionary history may be inconsequential, as the non-native Neotropical tree Bocconia

62

frutescens had the second highest percentage of seeds removed both from under trees and

in exposed sites (Table 3.1). Other native Hawaiian dry forest species, including Nestegis

sandwicensis (Oleaceae), Alectryon macrococcus (Sapindaceae), Nesoluma polynesicum

(Sapotaceae) and Pouteria sandwicensis (Sapotaceae) all suffer from high levels of seed

predation, as evidenced by the large piles of seed husks observed under almost every

fruiting tree crown (Chimera pers. obs.). Different, and potentially much higher, removal

rates would be expected for every unique species combination used in similar predation

trials.

Location of seeds and fruits, whether under trees or in exposed sites, also resulted

in variable removal rates. In three trials, more seeds were removed from under trees than

exposed sites, but location was not significant for Bocconia frutescens seeds and fruits or

Pleomele auwahiensis fruits (Table 3.2; Figures 3.1-3.5). Several authors have found

microhabitat to be an important factor influencing levels of seed predation. Many studies

found lower levels of predation by small mammals in open areas (Aide and Cavelier

1994; Bustamante and Simonetti 2000; Hulme 1994; Holl 2002). Others report that

higher levels occur in open sites versus under trees (Uhl 1998; Wijdeven and Kuzee

2000). Because seed densities have been shown to influence seed predation (Hulme

1998), different removal rates were expected under trees and exposed sites due to the

much higher seed densities found under trees at KNAR (Chapter 2).

Some authors suggest that aerial predators such as owls may deter rodents from

foraging in exposed areas, reducing seed removal in these sites (Murúa and González

1982; Sánchez-Cordeiro and Martínez-Gallardo 1998). Two owl species occur in KNAR,

the native Hawaiian short-eared owl (Asio flammeus), and the common barn-owl (Tyto

63

alba), a species introduced from North America in 1958 for rodent control in sugar cane

fields (Tomich 1962). If seed densities or aerial predators do influence removal rates in

different locations, these effects were negligible in two of the trials (Table 3.2). It is

possible that owls occur at densities too low to noticeably affect seed predation in

exposed sites, and those differences that were significant may be attributed to some other

factor.

For the trials that did not differ, the location of the exposed site treatments, within

five meters of the tree crown, may be too close to account for a truly separate

microhabitat, despite the differences in vegetative cover, exposure, amounts of leaf litter,

understory humidity or seed densities. In addition, placing five seeds in each treatment

results in much higher seed densities in exposed sites than would naturally occur away

from the tree crown. This increased density, intended to represent either a hypothetical

dispersal event by a frugivore or a supplementary seed-scattering technique for

restoration purposes, could contribute to unnaturally high predation levels, as rodents

have demonstrated the ability to more easily locate bigger seed piles (Hammond 1995).

None of the preceding explanations satisfactorily address why removal rates differed in

locations for three trials, but not for the other two, however, so yet another factor may be

responsible for the variability. It may simply be that the short duration of the study

accounted for differences detected by location (Figure 3.1), and keeping seeds out longer

may result in equivalent final predation levels for all trials.

Seed size is another factor thought to affect amounts and rates of seed predation

(Hulme 1998). Although no relationship was found between seed length and width and

rates of removal in either location, the range of sizes used in this study may be too small

64

to detect any size effects on predation levels. As the seeds used in this study were

relatively large (Table 3.1), future experiments should utilize a broader range of seed

sizes, along with different densities, to more accurately determine if any real relationships

exist between seed size or mass and removal rates under trees and exposed sites.

Higher removal rates for fruits than seeds were reported in New Zealand (Moles

and Drake 1999) and Spain (Hulme 1997). Presence of pulp on Diospyros seeds and the

fleshy aril on Bocconia seeds also resulted in higher rates of removal than those recorded

for most of the cleaned seed treatments (Figure 3.2). More Pleomele auwahiensis fruits

than seeds were removed, but their location was not significant (Table 3.3; Figure 3.4).

For these three species, it is unknown whether fruit removal results in destruction or

dispersal of the seed or seeds. As intact Reynoldsia seeds have previously been found in

rat droppings (Medeiros et al. 1986), the use of cleaned seeds in the various treatments

may explain the lack of removal of any Reynoldsia seeds (Figure 3.5). For this species,

rodents may be more attracted to fruits and could potentially benefit the tree by

legitimately dispersing seeds. In general, fruit pulp appears to contribute to higher overall

removal rates, but further tests with a wider range of Hawaiian dry forest species is

necessary before more definitive conclusions can be made about the effects of pulp on

seed removal or predation.

As very few fleshy-fruited dry forest species are dispersed to exposed sites

(Chapter 2), differences in removal rates under trees and in exposed sites may be

unimportant to the fitness and survival of these plants. Even if dispersal to exposed sites

did allow seeds to escape predation, the harsher microsite conditions would likely limit

seedling recruitment and survival (Kitajima and Fenner 2001). In addition, the species of

65

rodent removing seeds is probably also unimportant, as the non-native black rat (Rattus

rattus) and mouse (Mus musculus) recorded in KNAR (Medeiros et al. 1993), and

possibly the Polynesian rat (Rattus exulans), are all known to be important, generalist

seed predators on a range of seed sizes. No trapping was conducted during this study, so

it is unknown which rodent species is responsible for the seed removal levels reported

here. Differences in removal rates from some accessible treatments may be due to the

reluctance of rodents to enter these artificial environments, as other studies have also

reported (Inglis et al. 1996; Moles and Drake 1999). The high rates of removal recorded

for all rodent-accessible treatments of Santalum ellipticum (Figure 3.5) suggest that

rodents will enter artificial environments to attain highly desirable seeds of certain

species. Some of the removal of seeds and fruits from all trials may be due to the foraging

of non-native game birds, including the black (Francolinus francolinus) and gray (F.

pondicerianus) francolin, the ring-necked pheasant (Phasianus colchicus) and the

common peafowl (Pavo cristatus), and removal of seeds or fruits could therefore result in

both dispersal and predation. Nevertheless, presence of teeth marks on seeds and gnawed

husks suggest that rodents were also responsible for at least some of the seed and fruit

removal.

Seed predation of particular Hawaiian dry forest tree species appears to be one of

the many important factors contributing to lack of seedling recruitment of these species.

Seed predators such as rodents sometimes benefit seeds by moving them to sites

favorable for seedling establishment (Vander Wall and Longland 2004). Although

introduced rodents may have limited beneficial effects if removal results in dispersal

rather than seed destruction, for many Hawaiian species, the negative impacts probably

66

outweigh any benefits. Howe and Brown (1999) report that seed predators selectively

remove individuals of some species more than others, thereby giving a competitive edge

to less preferred species. For Hawaiian plants that evolved in the absence of rodent

predation, such as the heavily depredated Santalum ellipticum, this is particularly

relevant, but for other species experiencing moderate levels of predation, the effects may

be negligible. For invasive plants such as Bocconia frutescens that evolved with rodent

predators, seed removal and predation do not appear to be preventing the continued

spread of this species into native forest habitat. As seed predation can be a major

constraint for tree regeneration in neotropical forests (Guariguata and Pinard 1998), it

would be valuable to know to what degree predation limits recruitment of each species.

Future conservation and restoration efforts in degraded forest ecosystems of the Hawaiian

Islands and elsewhere could therefore begin to address and attempt to alleviate the

impacts of seed predation on those species identified as particularly vulnerable.

67

CHAPTER 4: CONCLUSION

Tropical dry forests throughout the world have been severely reduced in extent

and continue to be lost at an alarming rate. Several factors have contributed to this

decline, mostly related to anthropogenic alteration of ecosystems. Although the direct and

indirect impacts of humans have greatly modified the vegetation of Hawaiian dry forests,

native tree diversity and abundance remain relatively high. Ungulate exclusion and other

management strategies may address some of the problems associated with lack of native

tree reproduction, but other factors will continue to play an important role in the future

composition of these areas. Native dry forest trees are extremely important seed sources

and microsites for both natural recruitment and propagation programs, and non-native

trees are centers of continued invasion and habitat modification. Understanding the

dynamic interaction of trees with potential dispersers and seed predators not only allows

for comparisons with and evaluations of more general theories of seed ecology, but also

helps to guide management decisions aimed at preserving and restoring the remaining dry

forest ecosystems.

In this study, I attempted to address how seed dispersal patterns and seed

predation by rodents interact with tree distribution in a Hawaiian dry forest. As the native

avifauna has been entirely extirpated and replaced by an assemblage of generalist

frugivores, granivores and game birds, I expected that the distribution of seeds

throughout the area and the types of seeds being dispersed would be influenced by the

types of birds currently found at the site. To gain insights into the processes of seed

dispersal, I addressed the following hypotheses:

68

1) There is no difference in the seed rain of fleshy-fruited, bird-dispersed seeds

under trees versus in exposed areas.

2) Non-native birds such as the Japanese white-eye (Zosterops japonicus), the

common mynah (Acridotheres tristis), the zebra dove (Geopelia striata) the

northern cardinal (Cardinalis cardinalis) and the northern mockingbird (Mimus

polyglottos) are not dispersing the seeds of fleshy-fruited native and non-native

trees.

3) There is no relationship between seed size and amount of seeds dispersed for

fleshy-fruited species.

In all three cases involving seed dispersal, I rejected the null hypotheses. For hypothesis

1, seed rain of bird-dispersed seeds was significantly higher under all tree species versus

exposed sites. In particular, by foraging, perching and nesting in trees, birds are directing

the dispersal of certain species under their crowns. In exposed sites, bird-dispersed seeds

were almost entirely absent, and only seeds not dependent on animal vectors for

dispersal, such as wind-adapted grasses, were reaching these areas. For hypothesis 2,

indirect evidence from the seed rain and direct observations of foraging on fruiting trees

indicated that at least three species of non-native birds were dispersing the seeds of some

native and non-native plants. Among these bird species, the Japanese white-eye was by

far the most common at the site. Northern cardinals and northern mockingbirds were also

frequently observed, but at much lower numbers. The quality of frugivory differs among

the bird species, with white-eyes swallowing entire fruits and seeds of some tree species,

and pecking at the pulp of others. They are therefore acting as both legitimate dispersers

69

and as pulp predators. Mockingbirds and cardinals were legitimate dispersers of certain

species, but cardinals were also pre-dispersal seed predators of Santalum ellipticum and

possibly other trees. Furthermore, although a larger number of native species compared

to non-natives were collected in the bird-dispersed seed rain, non-natives far

outnumbered natives in total seed numbers and densities. In particular, Bocconia

frutescens and Lantana camara were the most abundant and widespread in all seed traps

under trees. For hypothesis 3, there was a significant negative relationship between seed

width and numbers of seeds dispersed by birds. This may be due to limitations on body

and gape size of the three most common frugivores observed at the study site.

The Hawaiian Islands, which lacked native rodents, now have four species that

are widely distributed in both modified and native ecosystems. As rodents are well-

documented seed predators, I expected that rodent seed predation at KNAR would be an

important factor influencing seedling recruitment and future tree distribution. To gain

insights into the processes and potential effects of seed predation, I addressed the

following null hypotheses:

4) Rodents are not removing and destroying seeds of dry forest species.

5) There is no difference in removal rates of seeds under trees versus exposed

sites.

6) There is no relationship between seed size and removal rates for dry forest tree

species.

I rejected null hypothesis 4 for certain dry forest species, as rodents were removing and

destroying the seeds of four study tree species. In particular, direct removal trials and

presence of seed husks indicated that rodents were removing and destroying seeds of

70

Bocconia frutescens, Diospyros sandwicensis, Pleomele auwahiensis, and Santalum

ellipticum. Of these tree species, S. ellipticum seeds appeared to suffer the highest rates of

removal and destruction. I did not reject the null hypothesis for Reynoldsia sandwicensis

seeds, as rodents removed no seeds during the entire duration of the study. I rejected null

hypothesis 5 for D. sandwicensis, P. auwahiensis seeds and S. ellipticum, as removal

rates for these species were significantly higher under trees versus exposed sites. I did not

reject the null hypothesis for Bocconia frutescens, P. auwahiensis fruits and R.

sandwicensis, as there were no significant differences in removal rates under trees versus

exposed sites. I did not reject null hypothesis 6, as no significant relationship was found

between seed size and removal rates for the dry forest species used in this study. As only

five species were used for the size analysis, however, small sample size may account for

the lack of any relationship. Further studies using more species with a broader range of

sizes will better detect whether or not a relationship between seed size and predation does

exist for dry forest trees.

In conclusion, current dispersal patterns indicate that a few readily disseminated

non-native species are being spread throughout KNAR. For fleshy-fruited species, non-

native frugivores are responsible for the dispersal of predominantly non-native invasive

species under standing trees, while the majority of native seeds are falling directly onto

the ground without the benefits of scarification or dispersal. This pattern of seed rain, the

distribution of seedlings under trees, and the high levels of predation on native seeds

suggest that, without management intervention, the diverse, native dry forest community

that currently exists will eventually be replaced by a homogenous landscape of a few

prolific invaders. For any semblance of a functioning native Hawaiian dry forest

71

ecosystem to persist into the future, immediate threats must be addressed by removing

non-native species dispersed by birds and by eliminating rodents depredating native

seeds. The critical issue will then be whether extinct birds can be replaced by truly

effective surrogates capable of scarifying and disseminating the wide range of native

species still extant in the remaining dry forests, or whether the conservation of these

species has become entirely dependent on human intervention.

72

APPENDIX A. TREE DIMENSIONS (BASAL DIAMETER, HEIGHT, CROWN

DIAMETER) OF STUDY SPECIES AND DEAD TREES

Table A.1. Tree dimensions of study species used in sampling of fleshy-fruited seeds and seed removal trials. Values are means (± 1 SE) for 15 of each species and 15 dead trees.

Tree species Basal diameter (cm) Height (m) Crown diameter

(m) Santalum ellipticum 17.62 (1.34) 2.76 (0.19) 3.70 (0.23)

Diospyros sandwicensis 22.02 (1.38) 4.72 (0.39) 4.41 (0.22)

Dead Trees 22.80 (2.48) 3.77 (0.29) 3.00 (0.23)

Bocconia frutescens 27.30 (1.94) 3.21 (0.12) 3.27 (0.12)

Pleomele auwahiensis 33.32 (1.63) 5.30 (0.22) 3.23 (0.17)

Reynoldsia sandwicensis 38.28 (2.43) 5.77 (0.32) 5.73 (0.34)

73

Figure A.1. Tree dimensions for study species and used in sampling of fleshy-fruited seeds. Bars represent means +1 SE (n = 15 per tree). Tree dimensions were compared with a one-way ANOVA. Letters above the bars show the results of the Tukey test. Treatments sharing the same letter are not significantly different (p > 0.05).

74

APPENDIX B. PHENOLOGY OF STUDY TREES AND COMMON FLESHY-

FRUITED SHRUBS OF THE KANAIO NATURAL AREA RESERVE

METHODS

The presence of immature and ripe fruits on sample trees and other taxa in the

area was recorded on a monthly basis from March 2003 to February 2004 to document

the potential pool of bird-dispersed seeds that could be disseminated in the area. For the

five fleshy-fruited tree species used in the seed dispersal study, percentages of stems with

immature and ripe fruit were estimated by categories (0 = none; 1 = 1-5 percent; 2 = 6-25

percent; 3 = 26-50 percent; 4 = 51-75 percent; 5 = 76-100 percent) on 15 to 30 trees per

species occurring along seed sampling transects. Means for trees were calculated using

the following values for each phenology category: 0 = 0; 1 = 3 percent; 2 = 15 percent; 3

= 37.5 percent; 4 = 62.5 percent; 5 = 87.5 percent (Tables B1-B3; Figures B.1. and B.3.).

For three widespread, fleshy-fruited shrubs, the non-native Lantana camara, and the

natives Osteomeles anthyllidifolia and Wikstroemia monticola, 25 plants of each species

were selected along the seed-sampling transects such that no conspecific plants were

within 25 meter of each other. For each plant, a 50-cm branch segment was tagged and

numbers of immature and ripe fruits counted along that length for each month (Table B.4;

Figures B.2. and B.4.). Phenology for these plants was recorded as they are common

throughout the reserve and possess relatively small, fleshy fruits with characteristics

attractive to either obligate or opportunistic frugivores (Howe 1986).

75

Table B.1. Monthly percentage of fleshy-fruited trees and shrubs with mature fruit. Boc fru: Bocconia frutescens (n = 28); Dio san: Diospyros sandwicensis (n = 25); Ple auw: Pleomele auwahiensis (n = 30); Rey san: Reynoldsia sandwicensis (n = 15); San ell: Santalum ellipticum (n = 28); Lan cam: Lantana camara (n = 25); Ost ant: Osteomeles anthyllidifolia (n = 25); Wik mon: Wikstroemia monticola (n = 25).

Date Boc fru Dio san Ple auw Rey san San ell Lan cam Ost ant Wik mon

Mar-03 0.00 72.00 63.33 40.00 21.43 20.00 0.00 0.00 Apr-03 0.00 72.00 60.00 0.00 28.57 84.00 4.00 4.00 May-03 96.43 64.00 43.33 0.00 17.86 92.00 0.00 24.00 Jun-03 100.00 36.00 13.33 0.00 3.57 80.00 4.00 0.00 Jul-03 100.00 16.00 0.00 0.00 7.14 12.00 8.00 0.00 Aug-03 53.57 52.00 6.67 0.00 3.57 48.00 20.00 8.00 Sep-03 0.00 64.00 6.67 0.00 3.57 4.00 52.00 0.00 Oct-03 0.00 96.00 6.67 0.00 3.57 0.00 64.00 0.00 Nov-03 0.00 100.00 3.33 0.00 0.00 0.00 60.00 0.00 Dec-03 0.00 92.00 13.33 46.67 10.71 0.00 28.00 0.00 Jan-04 0.00 80.00 13.33 80.00 0.00 12.00 0.00 0.00 Feb-04 0.00 80.00 13.33 20.00 42.86 84.00 0.00 32.00

Table B.2. Mean (± 1 SE) monthly percentage of stems with immature fruits by tree species. Means were calculated by assigning values to phenology categories 1-5 (0 = none; 1 = 2.5 percent; 2 = 15.5 percent; 3 = 37.5 percent; 4 = 62.5 percent; 5 = 87.5 percent). Boc fru: Bocconia frutescens (n = 28); Dio san: Diospyros sandwicensis (n = 25); Ple auw: Pleomele auwahiensis (n = 30); Rey san: Reynoldsia sandwicensis (n = 15); San ell: Santalum ellipticum (n = 28).

Date Boc fru Dio san Ple auw Rey san San ell

Mar-03 53.95 (4.50) 26.82 (6.12) 0.00 (0.00) 0.20 (0.20) 2.73 (1.41)Apr-03 71.88 (4.20) 12.58 (3.32) 0.00 (0.00) 0.00 (0.00) 1.71 (0.57)May-03 77.68 (2.68) 11.48 (3.68) 0.00 (0.00) 0.00 (0.00) 2.30 (1.33)Jun-03 59.02 (2.42) 25.28 (4.55) 0.10 (0.10) 0.00 (0.00) 2.36 (0.88)Jul-03 24.30 (2.58) 47.00 (4.83) 2.08 (2.08) 0.00 (0.00) 4.39 (1.92)Aug-03 2.04 (0.56) 55.90 (4.75) 0.00 (0.00) 0.00 (0.00) 1.18 (0.56)Sep-03 0.00 (0.00) 67.60 (4.28) 0.00 (0.00) 0.00 (0.00) 1.29 (0.57)Oct-03 0.00 (0.00) 67.60 (4.28) 0.00 (0.00) 3.70 (2.61) 1.39 (0.57)Nov-03 0.00 (0.00) 62.80 (4.88) 0.00 (0.00) 17.60 (6.97) 1.93 (0.89)Dec-03 0.00 (0.00) 52.52 (4.80) 0.00 (0.00) 22.80 (6.41) 7.41 (3.20)Jan-04 0.00 (0.00) 36.18 (5.07) 0.00 (0.00) 15.93 (5.89) 10.95 (4.71)Feb-04 0.00 (0.00) 20.68 (3.53) 0.00 (0.00) 2.20 (1.36) 15.54 (5.13)

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Table B.3. Mean (± 1 SE) monthly percentage of stems with mature fruits by tree species. Means were calculated by assigning values to phenology categories 1-5 (0 = none; 1 = 2.5 percent; 2 = 15.5 percent; 3 = 37.5 percent; 4 = 62.5 percent; 5 = 87.5 percent). Boc fru: Bocconia frutescens (n = 28); Dio san: Diospyros sandwicensis (n = 25); Ple auw: Pleomele auwahiensis (n = 30); Rey san: Reynoldsia sandwicensis (n = 15); San ell: Santalum ellipticum (n = 28).

Date Boc fru Dio san Ple auw Rey san San ell

Mar-03 0.00 (0.00) 4.02 (1.51) 24.57 (5.83) 1.20 (0.39) 0.64 (0.24)Apr-03 0.00 (0.00) 3.60 (0.90) 16.50 (4.57) 0.00 (0.00) 0.86 (0.26)May-03 8.36 (1.85) 1.92 (0.29) 4.88 (2.19) 0.00 (0.00) 0.54 (0.22)Jun-03 25.34 (2.62) 1.56 (0.63) 0.40 (0.19) 0.00 (0.00) 0.11 (0.11)Jul-03 11.89 (1.75) 0.48 (0.22) 0.00 (0.00) 0.00 (0.00) 0.21 (0.15)Aug-03 1.61 (0.29) 2.52 (0.81) 2.18 (2.08) 0.00 (0.00) 0.11 (0.11)Sep-03 0.00 (0.00) 1.92 (0.29) 0.20 (0.14) 0.00 (0.00) 0.11 (0.11)Oct-03 0.00 (0.00) 3.36 (0.50) 0.20 (0.14) 0.00 (0.00) 0.11 (0.11)Nov-03 0.00 (0.00) 11.88 (2.52) 1.25 (1.25) 0.00 (0.00) 0.00 (0.00)Dec-03 0.00 (0.00) 18.94 (3.14) 4.77 (2.91) 5.30 (2.65) 0.75 (0.55)Jan-04 0.00 (0.00) 8.04 (1.27) 2.78 (2.12) 7.10 (2.58) 0.00 (0.00)Feb-04 0.00 (0.00) 7.14 (1.74) 2.70 (1.73) 0.60 (0.32) 1.71 (0.57)

Table B.4. Mean (± 1 SE) monthly number of immature fruits (IF) and mature fruits (MF) on branches by shrub species. LC: Lantana camara (n = 25); OA: Osteomeles anthyllidifolia (n = 25); WM: Wikstroemia monticola (n = 25).

LC LC OA OA WM WM Date IF MF IF MF IF MF Mar-03 4.32 (0.71) 0.28 (0.12) 3.64 (0.90) 0.00 (0.00) 1.80 (0.92) 0.00 (0.00)Apr-03 1.40 (0.31) 2.04 (0.39) 7.36 (1.39) 0.20 (0.20) 1.24 (0.68) 0.40 (0.40)May-03 1.68 (0.56) 3.80 (0.67) 8.88 (1.68) 0.00 (0.00) 0.16 (0.16) 0.48 (0.22)Jun-03 0.52 (0.25) 3.36 (0.68) 8.96 (1.74) 0.04 (0.04) 0.00 (0.00) 0.00 (0.00)Jul-03 0.80 (0.32) 0.24 (0.15) 8.24 (1.56) 0.08 (0.06) 0.00 (0.00) 0.00 (0.00)Aug-03 0.04 (0.04) 0.88 (0.22) 6.48 (1.11) 0.40 (0.19) 0.44 (0.28) 0.08 (0.06)Sep-03 0.00 (0.00) 0.04 (0.04) 5.32 (0.97) 1.24 (0.35) 0.00 (0.00) 0.00 (0.00)Oct-03 0.00 (0.00) 0.00 (0.00) 2.08 (0.86) 1.76 (0.68) 0.00 (0.00) 0.00 (0.00)Nov-03 0.00 (0.00) 0.00 (0.00) 0.20 (0.16) 2.68 (0.73) 0.00 (0.00) 0.00 (0.00)Dec-03 0.76 (0.30) 0.00 (0.00) 0.00 (0.00) 0.80 (0.35) 1.44 (0.91) 0.00 (0.00)Jan-04 4.40 (0.75) 0.16 (0.09) 0.00 (0.00) 0.00 (0.00) 4.52 (1.63) 0.00 (0.00)Feb-04 2.00 (0.53) 6.20 (0.98) 0.00 (0.00) 0.00 (0.00) 1.08 (0.36) 3.16 (1.23)

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Figure B.1 Monthly percentage of fleshy-fruited trees with mature fruit. Phenology recorded from March 2003 to February 2004. Boc fru: Bocconia frutescens (n = 28); Dio san: Diospyros sandwicensis (n = 25); Ple auw: Pleomele auwahiensis (n = 30); Rey san: Reynoldsia sandwicensis (n = 15); San ell: Santalum ellipticum (n = 28).

78

Figure B.2. Monthly percentage of fleshy-fruited shrubs with mature fruit. Phenology recorded from March 2003 to February 2004. Lan cam: Lantana camara (n = 25); Ost ant: Osteomeles anthyllidifolia (n = 25); Wik mon: Wikstroemia monticola (n = 25).

79

Figure B.3. Mean (± 1 SE) monthly percentage of stems with immature and mature fruits by tree species. Means were calculated by assigning values to phenology categories 1-5 (0 = none; 1 = 2.5 percent; 2 = 15.5 percent; 3 = 37.5 percent; 4 = 62.5 percent; 5 = 87.5 percent). Dotted lines represent immature fruits and solid lines represent mature fruits. A: Bocconia frutescens (n = 28); B: Diospyros sandwicensis (n = 25); C: Pleomele auwahiensis (n = 30); D: Reynoldsia sandwicensis (n = 15); E: Santalum ellipticum (n = 28). Note that Bocconia and Diospyros (A & B) are scaled 0-100, and Pleomele, Reynoldsia and Santalum (C-E) are scaled 0-50. Phenology recorded from March 2003 to February 2004.

80

Figure B.4. Mean (± 1 SE) monthly number of immature and mature fruits on branches by shrub species. Dotted lines represent immature fruits and solid lines represent mature fruits. A: Lantana camara (n = 25); B: Osteomeles anthyllidifolia (n = 25); C: Wikstroemia monticola (n = 25). Phenology recorded from March 2003 to February 2004.

81

APPENDIX C. RELATIVE FREQUENCY AND ABUNDANCE OF NON-NATIVE

BIRDS IN KNAR STUDY SITE

METHODS

To assess the relative frequency and abundance of bird species at the study site in

KNAR, sampling was conducted at 10 stations along a north-south oriented transect

roughly bisecting the seed sampling transects and extending from approximately 650-830

meters elevation. Sampling stations were placed at 134-m intervals, following protocols

described by Scott et al. (1986). Each station was surveyed monthly on days of good

weather (i.e. no heavy rain or wind) from April 2003 to March 2004. Surveys began

shortly after sunrise and continued until all 10 stations were sampled. Estimates of bird

numbers at each station were made using the variable circular-plot method (Reynolds et

al. 1980). During an eight-minute count period at each station, each bird species both

heard and seen was recorded, and horizontal distances to each bird were estimated.

Because these counts were intended to give rough estimates of bird abundance, and were

not meant to derive estimates of densities for all birds in the dry forest habitat, data for

each species from all 10 stations per month were combined, and mean numbers of birds

of each species per month were calculated for the 12-month duration of the sampling.

The frequency of each bird species was calculated by recording the presence of a species

at each of the ten stations for the 12-month duration of the sampling, and dividing this

number by 120. The relative frequency and abundance of each species was then

calculated from these totals (Table C.1; Figures C.1-C.2). For three important non-native

82

frugivores, the Japanese white-eye (Zosterops japonicus), the northern cardinal

(Cardinalis cardinalis), and the northern mockingbird (Mimus polyglottos), mean number

of birds at the 10 sampling stations was also calculated for each sampling month, as these

species have been observed consuming fruits of some of the study species (Figure C.3).

Table C.1. Relative frequency and abundance of bird species recorded at 10 sampling stations over a 12-month period (April 2003 – March 2004) in KNAR.

Species Common Name Family Rel. Freq

Rel. Abund.

Zosterops japonicus Japanese white-eye Zosteropidae 27.0 57.2Cardinalis cardinalis northern cardinal Emberizidae 16.7 9.2Mimus polyglottos northern mockingbird Mimidae 12.4 7.1Carpodacus mexicanus housefinch Fringillidae 11.3 8.3Geopelia striata zebra dove Columbidae 8.1 4.0Francolinus francolinus black francolin Phasiandidae 6.8 3.4Streptopelia chinensis spotted dove Columbidae 5.6 2.8Francolinus pondicerianus gray francolin Phasiandidae 5.2 2.4Acridotheres tristis common myna Sturnidae 3.8 2.4Lonchura punctulata nutmeg mannikin Estrildidae 1.8 2.7Phasianus colchicus ring-necked pheasant Phasiandidae 0.9 0.4Cettia diphone Japanese bush-warbler Muscicapidae 0.5 0.2 Total 100 100

83

Figure C.1. Relative frequency of bird species recorded at 10 sampling stations over a 12-month period (April 2003 – March 2004) in KNAR.

84

Figure C.2. Relative abundance of bird species recorded at 10 sampling stations over a 12-month period (April 2003 – March 2004) in KNAR.

85

Figure C.3. Mean (± SE) number of three important non-native frugivores at 10 sampling stations over a 12-month period (April 2003 - March 2004) in KNAR.

86

APPENDIX D. CHARACTERIZATION OF THE VEGETATION OF KNAR

STUDY SITE

METHODS

To supplement the vegetation descriptions provided by Medeiros et al. (1986;

1993), plant cover in the study site was sampled at 1000 points using the point-intercept

method (Bonham 1989). Samples were taken at a random point in each 2-m segment of

ten 200-m transects placed midway between and parallel to the seed dispersal transects.

Points recorded under two meters height were classified as ground cover, and points over

two meters height were classified as canopy cover. For each species, percent cover for

ground and canopy was calculated separately by counting the number of points at which a

vertically-projected line intercepted living plant material and dividing this value by 1000.

More than one different species could be sampled per point, but no single species was

counted more than once per point.

87

Table D.1. Abundance (absolute, % cover and relative cover) of ground vegetation (< 2-m height) at KNAR study site. Data presented in order of decreasing abundance.

Category Form Status Absolute abundance % Cover Rel.

CoverROCK NA NA 358 35.8 30.4Lantana camara Shrub Non-native 263 26.3 22.3Melinis repens Graminoid Non-native 128 12.8 10.9Bidens pilosa Forb Non-native 115 11.5 9.8Dodonaea viscosa Shrub Native 61 6.1 5.2Ipomoea indica Liana Native 54 5.4 4.6Osteomeles anthyllidifolia Shrub Native 47 4.7 4.0Melinis minutiflora Graminoid Non-native 44 4.4 3.7LITTER NA NA 15 1.5 1.3Wikstroemia monticola Shrub Native 11 1.1 0.9Chamaecrista nictitans Forb Non-native 11 1.1 0.9Santalum ellipticum Tree Native 6 0.6 0.5Petroselinum crispum Forb Non-native 6 0.6 0.5Emilia fosbergii Forb Non-native 5 0.5 0.4Pennisetum clandestinum Graminoid Non-native 5 0.5 0.4Plectranthus parviflorus Forb Native 4 0.4 0.3Ageratina adenophora Forb Non-native 4 0.4 0.3Neonotonia wightii Vine Non-native 3 0.3 0.3Galinsoga parviflora Forb Non-native 3 0.3 0.3Bocconia frutescens Tree Non-native 2 0.2 0.2Cocculus orbiculatus Vine Native 2 0.2 0.2Leptecophylla tameiameiae Shrub Native 2 0.2 0.2Carex wahuensis Graminoid Native 2 0.2 0.2Passiflora subpeltata Vine Non-native 2 0.2 0.2Doryopteris decipiens Fern Native 2 0.2 0.2Sonchus oleraceus Forb Non-native 2 0.2 0.2Leucaena leucocephala Shrub/Tree Non-native 2 0.2 0.2Salvia coccinea Forb Non-native 2 0.2 0.2Cyperus hillebrandii Graminoid Native 1 0.1 0.1Senecio madagascariensis Forb Non-native 1 0.1 0.1Ageratina riparia Forb Non-native 1 0.1 0.1Pellaea ternifolia Fern Native 1 0.1 0.1Anagallis arvensis Forb Non-native 1 0.1 0.1Nephrolepis multiflora Fern Non-native 1 0.1 0.1Alyxia oliviformis Vine Native 1 0.1 0.1Schinus terebinthifolius Tree Non-native 1 0.1 0.1Lepisorus thunbergianus Fern Native 1 0.1 0.1Chenopodium oahuense Shrub Native 1 0.1 0.1Metrosideros polymorpha Tree Native 1 0.1 0.1

88

Table D.1 (Continued) Abundance (absolute, % cover and relative cover) of ground vegetation (< 2-m height) at KNAR study site.

Category Form Status Absolute abundance % Cover Rel. Cover

Solanum americanum Forb Non-native 1 0.1 0.1Pityogramma austroamericana Fern Non-native 1 0.1 0.1Peperomia blanda Forb Native 1 0.1 0.1Indigofera suffruticosa Vine Non-native 1 0.1 0.1Portulaca pilosa Forb Non-native 1 0.1 0.1TOTAL 1177 117.7 100.0

Table D.2. Abundance (absolute, and relative cover) of canopy vegetation (> 2-m height) at KNAR study site. Data presented in order of decreasing abundance.

Category Form Status Absolute Rel.Cover

OPEN NA NA 844 84.4Pleomele auwahiensis Tree Native 48 4.8Bocconia frutescens Tree Non-native 28 2.8Diospyros sandwicensis Tree Native 26 2.6Dodonaea viscosa Shrub Native 10 1.0Santalum ellipticum Tree Native 10 1.0Nothocestrum latifolium Tree Native 9 0.9Wikstroemia monticola Shrub Native 6 0.6Xylosma hawaiiense Tree Native 6 0.6DEAD TREES Tree NA 4 0.4Reynoldsia sandwicensis Tree Native 4 0.4Myoporum sandwicense Tree Native 2 0.2Opuntia ficus-indica Shrub Non-native 1 0.1Myrsine lanaiensis Tree Native 1 0.1Nestegis sandwicensis Tree Native 1 0.1Total 1000 100.0

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Table D.3. Abundance (absolute and relative cover) of native and non-native vegetation, dead trees, bare ground, and open sky at KNAR study site.

Category Ground Rel.

Ground Cover

Canopy Rel. Canopy Cover

Native Plants 198 16.8 123 12.3 Non-native Plants 606 51.5 29 2.9 Dead trees 0 0.0 4 0.4 Bare Ground 373 31.7 NA NA Open NA NA 844 84.4 Total 1177 100.0 1000 100.0

90

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