Seed Rain and Seed Predation in a Hawaiian Dry Forest Community

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

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

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.

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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

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

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

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

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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

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

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

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.

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;

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

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. 8

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.

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).

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;

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.

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

(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

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

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?

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 (Ààliì) 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.).

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

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),

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

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”.

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

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

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).

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 % Total Wind 595.6 (96.5) 1362.4 (283.6) 0.167 70 Gravity 321.0 (80.1) 42.5 (17.7) <0.001 13.0 Birds 237.9 (30.2) 7.7 (2.97) <0.001 8.8 Uncertain 230.5 (68.0) 1.3 (0.75) <0.001 8.3

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

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

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

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)

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

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 (*). Boc Dio Ple Rey San All Dead Exp. Seed Species fru san auw san ell Trees Non-natives 100.0 100.0 100.0 93.3 100.0 86.7 15.6 96.7 Natives 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 1.1 Bocconia frutescens* 100.0 93.3 100.0 86.7 93.3 86.7 5.6 93.3 Cocculus orbiculatus 0.0 0.0 13.3 6.7 0.0 6.7 0.0 4.4 Diospyros sandwicensis 0.0 0.0 100.0 0.0 13.3 0.0 0.0 18.9 Lantana camara* 73.3 86.7 66.7 73.3 93.3 53.3 11.1 74.4 Leptecophylla tameiameiae 0.0 0.0 20.0 0.0 6.7 0.0 0.0 4.4 Myoporum sandwicense 0.0 6.7 0.0 0.0 0.0 0.0 0.0 1.1 Nothocestrum latifolium 6.7 13.3 0.0 0.0 6.7 0.0 0.0 4.4 Osteomeles anthyllidifolia 13.3 33.3 20.0 20.0 13.3 0.0 0.0 16.7 Pleomele auwahiensis 0.0 0.0 0.0 46.7 0.0 0.0 0.0 7.8 Reynoldsia sandwicensis 0.0 0.0 0.0 0.0 80.0 0.0 0.0 13.3 Santalum ellipticum 0.0 0.0 0.0 0.0 0.0 46.7 0.0 7.8 Schinus terebinthifolius* 0.0 0.0 0.0 0.0 0.0 13.3 0.0 2.2 Wikstroemia monticola 13.3 20.0 6.7 6.7 0.0 0.0 0.0 7.8

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).

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).

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.

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 Trees Alyxia 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).

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

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. Greatest length Greatest width # bird- Seed species without pulp without pulp n dispersed (mm) (mm) 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

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

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

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).

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

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.

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 41

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.

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, 42

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.

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 43

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

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.

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

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

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).

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

(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 (Ààliì) lowland shrublands,

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.

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

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

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.

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

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. % of cleaned Greatest Greatest % of cleaned seeds length width seeds Species removed after without pulp without removed after 15 days (mm) pulp (mm) 15 days (tree) (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

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.

Source of SS Species d.f. MS F P variation (model III) 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

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; S= cleaned seeds in exposed sites; = seeds under trees (with pulp); z = 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.

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.

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).

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

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.

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

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

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

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

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

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.

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:

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

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

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

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.

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. Basal diameter Crown diameter Tree species Height (m) (cm) (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)

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).

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).

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)

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)

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).

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).

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.

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.

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

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. Rel. Rel. Species Common Name Family Freq Abund. Zosterops japonicus Japanese white-eye Zosteropidae 27.0 57.2 Cardinalis cardinalis northern cardinal Emberizidae 16.7 9.2 Mimus polyglottos northern mockingbird Mimidae 12.4 7.1 Carpodacus mexicanus housefinch Fringillidae 11.3 8.3 Geopelia striata zebra dove Columbidae 8.1 4.0 Francolinus francolinus black francolin Phasiandidae 6.8 3.4 Streptopelia chinensis spotted dove Columbidae 5.6 2.8 Francolinus pondicerianus gray francolin Phasiandidae 5.2 2.4 Acridotheres tristis common myna Sturnidae 3.8 2.4 Lonchura punctulata nutmeg mannikin Estrildidae 1.8 2.7 Phasianus colchicus ring-necked pheasant Phasiandidae 0.9 0.4 Cettia diphone Japanese bush-warbler Muscicapidae 0.5 0.2 Total 100 100

Figure C.1. Relative frequency of bird species recorded at 10 sampling stations over a 12-month period (April 2003 – March 2004) in KNAR.

Figure C.2. Relative abundance of bird species recorded at 10 sampling stations over a 12-month period (April 2003 – March 2004) in KNAR.

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.

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.

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. Absolute Rel. Category Form Status % Cover abundance Cover ROCK NA NA 358 35.8 30.4 Lantana camara Shrub Non-native 263 26.3 22.3 Melinis repens Graminoid Non-native 128 12.8 10.9 Bidens pilosa Forb Non-native 115 11.5 9.8 Dodonaea viscosa Shrub Native 61 6.1 5.2 Ipomoea indica Liana Native 54 5.4 4.6 Osteomeles anthyllidifolia Shrub Native 47 4.7 4.0 Melinis minutiflora Graminoid Non-native 44 4.4 3.7 LITTER NA NA 15 1.5 1.3 Wikstroemia monticola Shrub Native 11 1.1 0.9 Chamaecrista nictitans Forb Non-native 11 1.1 0.9 Santalum ellipticum Tree Native 6 0.6 0.5 Petroselinum crispum Forb Non-native 6 0.6 0.5 Emilia fosbergii Forb Non-native 5 0.5 0.4 Pennisetum clandestinum Graminoid Non-native 5 0.5 0.4 Plectranthus parviflorus Forb Native 4 0.4 0.3 Ageratina adenophora Forb Non-native 4 0.4 0.3 Neonotonia wightii Vine Non-native 3 0.3 0.3 Galinsoga parviflora Forb Non-native 3 0.3 0.3 Bocconia frutescens Tree Non-native 2 0.2 0.2 Cocculus orbiculatus Vine Native 2 0.2 0.2 Leptecophylla tameiameiae Shrub Native 2 0.2 0.2 Carex wahuensis Graminoid Native 2 0.2 0.2 Passiflora subpeltata Vine Non-native 2 0.2 0.2 Doryopteris decipiens Fern Native 2 0.2 0.2 Sonchus oleraceus Forb Non-native 2 0.2 0.2 Leucaena leucocephala Shrub/Tree Non-native 2 0.2 0.2 Salvia coccinea Forb Non-native 2 0.2 0.2 Cyperus hillebrandii Graminoid Native 1 0.1 0.1 Senecio madagascariensis Forb Non-native 1 0.1 0.1 Ageratina riparia Forb Non-native 1 0.1 0.1 Pellaea ternifolia Fern Native 1 0.1 0.1 Anagallis arvensis Forb Non-native 1 0.1 0.1 Nephrolepis multiflora Fern Non-native 1 0.1 0.1 Alyxia oliviformis Vine Native 1 0.1 0.1 Schinus terebinthifolius Tree Non-native 1 0.1 0.1 Lepisorus thunbergianus Fern Native 1 0.1 0.1 Chenopodium oahuense Shrub Native 1 0.1 0.1 Metrosideros 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. Absolute Category Form Status % Cover Rel. Cover abundance Solanum americanum Forb Non-native 1 0.1 0.1 Pityogramma austroamericana Fern Non-native 1 0.1 0.1 Peperomia blanda Forb Native 1 0.1 0.1 Indigofera suffruticosa Vine Non-native 1 0.1 0.1 Portulaca pilosa Forb Non-native 1 0.1 0.1 TOTAL 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 84.4 Pleomele auwahiensis Tree Native 48 4.8 Bocconia frutescens Tree Non-native 28 2.8 Diospyros sandwicensis Tree Native 26 2.6 Dodonaea viscosa Shrub Native 10 1.0 Santalum ellipticum Tree Native 10 1.0 Nothocestrum latifolium Tree Native 9 0.9 Wikstroemia monticola Shrub Native 6 0.6 Xylosma hawaiiense Tree Native 6 0.6 DEAD TREES Tree NA 4 0.4 Reynoldsia sandwicensis Tree Native 4 0.4 Myoporum sandwicense Tree Native 2 0.2 Opuntia ficus-indica Shrub Non-native 1 0.1 Myrsine lanaiensis Tree Native 1 0.1 Nestegis sandwicensis Tree Native 1 0.1 Total 1000 100.0

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. Rel. Rel. Canopy Category Ground Ground Canopy Cover 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

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