Synthesis of Ecosystem Resources and Threats

Transcription

135
ECOSYSTEM RESTORATION ON SANTA CATALINA ISLAND: A SYNTHESIS OF
RESOURCES AND THREATS
Denise A. Knapp
University of California, Santa Barbara
Ecology, Evolution, and Marine Biology Department
Santa Barbara, CA 93106-9610
[email protected]
ABSTRACT: Catalina Island‘s oaks provide habitat for a diverse array of plants and animals, many of
them rare. The oak ecosystem encompasses a large majority of the island, yet the foundation of this
system, the oaks themselves, appear to be in decline. Seaver Institute funds allowed the Catalina Island
Conservancy and its partners to initiate a variety of ecological research projects and restoration initiatives
using an ecosystem-level approach. In this paper, the natural resources of the island are discussed (with an
emphasis on those depending on oaks), along with the threats to those resources and natural processes.
Oak habitat dominated by Quercus pacifica has declined by as much as 31 percent over the past 60+
years on the island, with no apparent recruitment into the canopy. Low moisture, poor dispersal,
trampling by bison, and browsing by mule deer appear to be limiting regeneration of this species, the
adults of which are likely dying due senescence (old age) hastened by stressors such as browsing by
introduced ungulates. Quercus tomentella is restricted to seven locations on the island, yet maintains
relatively high genetic diversity; the greatest threats to this species are likely small population sizes,
fragmentation, deer browsing, and root exposure caused by erosion. A multitude of invasive, transformer
species threaten the island‘s native and endemic species, along with hydrologic alteration, roads, and
increased fire frequency. Climate change is expected to exacerbate these existing threats, which
themselves interact, thus compounding and complicating their impacts. Although important steps have
already been taken towards protecting Catalina‘s ecosystem, such as feral goat and pig removal and
transformer plant control, much more action will be necessary if structure and function are to be
improved. Future restoration priorities should take into account the degree of the threat, the chances of
success, and the benefits to be derived from such action. The groundwork for setting those priorities has
been laid here; management recommendations will be presented elsewhere. The ecosystem framework
and collaborative approach taken here has had a multitude of benefits, including efficient use of funds and
the creation of partnerships; the result has been the true integration of science and management.
KEYWORDS: Catalina Island, dieback, ecosystem, oak regeneration, Quercus pacifica, Quercus
tomentella, restoration, transformer species
TABLE OF CONTENTS
INTRODUCTION………………………………………………………………………………………136
Background……………………………………………...………………………………………………136
Defining the oak ecosystem on Catalina…………………………………………………...……… .137
Setting and island history…………………………………………………………………….…….…..137
BIOLOGICAL RESOURCES……………………….……………………………………….……… ..143
Oak species………………………………………………………………………………………………143
Associated flora………………………………………………………………………………………… 144
Associated fauna…………………………………………………………………………………….….. 147
THREATS TO THE ISLAND’S OAKS…………………………………………………………..…. 150
Quercus pacifica……………………………………………………………………………………..…. 150
Oak ecosystem restoration on Santa Catalina Island, California: Proceedings of an on-island workshop,
February 2-4, 2007. Edited by D.A. Knapp. 2010. Catalina Island Conservancy, Avalon, CA.
136
Quercus tomentella……………………………………………………………………………………. ..153
THREATS TO THE ISLAND ECOSYSTEM…………………………………………………….…..153
Non-native, transformer species…………………………………………………………………..…... 154
Ungulates…………………………………………………………………………………………...…… 155
Feral goats………………………………………………………………………………………………..157
Feral pigs……………………………………………………………………………………………...….157
American bison……………………………………………………………………………………..…… 159
Mule deer…………………………………………………………………………………………….….. 160
Carnivores and rodents………………………………………………………………………………….. 163
Feral cats…………………………………………………………………………………………………163
Rats……………………………………………………………………………………………………….165
House mouse…………………………………………………………………………………………..… 168
Birds………………………………………………………………………………………………...…… 169
European starling…………………………………………………………………………………..…….169
Wild turkey…………………………………………………………………………………………..….. 171
Brown-headed cowbird………………………………………………………………………………….. 172
Amphibians…………………………………………………………………………………………….…172
American bullfrog…………………………………………………………………………………..…… 172
Invertebrates…………………………………………………………………………………………….. 174
European honeybee……………………………………………………………………………………… 174
Argentine ant………………………………………………………………………………………….… 175
Transformer Plants………………………………………………………………….……………...…... .177
Roads……………………………………………………………………………………………………..180
Hydrologic alteration…………………………………………………………………………………... 183
Fire……………………………………………………………………………………………………….185
Fire history………………………………………………………………………………………………. 185
CONCLUSIONS………………………………………………………………………………………. 187
ACKNOWLEDGMENTS……………………………………………………………………………... 188
LITERATURE CITED…………………………………………………………………………………189
INTRODUCTION
Background
Oak trees provide food, shelter, and habitat for a diverse array of native plants and animals in California
(Block et al. 1990; Pavlik et al. 1991), including over half of the state‘s native terrestrial vertebrates
(Tietje and Vreeland 1997). Oak woodland is more species rich than any other vegetation type in the state
(Barrett 1980; Verner 1980; Garrison 1996), and oak trees play an important role in maintaining water
quality, stabilizing slopes, and increasing soil fertility (Jackson et al. 1990; Dahlgren et al. 1997;
McCreary 2004). Yet over half of California‘s oak woodland has been lost due to clearing for rangeland,
agriculture and viticulture, and fuelwood, and conversion to residences and industry (Burcham 1957;
McCreary 2004). Simultaneously, existing stands of several oak species in California are in decline due to
a wide variety of postulated factors including introduced species, altered disturbance regimes such as fire
and flood, acorn and seedling predation, and fragmentation (Griffin 1971; McClaran 1986; Borchert et al.
1989; Brown and Davis 1991; Davis et al. 1991; Meyer 2002; Tyler et al. 2006; Zavaleta et al. 2007).
Many of the species showing stand level decline also show poor regeneration at the individual and stand
levels, and little is known about what limits both their regeneration and their stand level dynamics.
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On Santa Catalina Island (hereafter Catalina), oak woodlands are important both ecologically and
culturally. The two predominant oaks are Island scrub oak (Quercus pacifica) and Island oak (Quercus
tomentella). Together, they cover over 23% of Catalina (Figure 1), and comprise ecosystems that
encompass a large number of associated plant and animal species. It is thus alarming that oaks on Catalina
appear to be declining, with the island‘s most prominent oak species, Quercus pacifica, exhibiting standlevel dieback (Knapp 2002), primarily on the Channel side of the island (Figure 2), and little recruitment
noted for either species over the past several decades.
The Catalina Island Conservancy (hereafter Conservancy) owns and manages 88% of Catalina. In 2003,
the Conservancy received a grant from the Seaver Institute to begin restoration of the island‘s oak
ecosystem. A variety of ecological research projects were then initiated on the Island between 2003 and
2007 in order to learn more about the condition and dynamics of the oak ecosystem and to inform
restoration and management. At the same time, a variety of projects have catalyzed oak ecosystem
restoration on the island.
The purposes of this paper are to describe the natural resources of Catalina with specific reference to oaks
and to discuss the factors that threaten those resources and the natural processes on the island. The
accomplishments of the recent oak ecosystem-related research and restoration activities will also be
discussed. Recommendations for ecosystem-level restoration have been provided to the Conservancy but
are not presented here, at the Conservancy‘s request.
Defining the Oak Ecosystem on Catalina
The Society for Ecological Restoration (SER International 2004) defines an ecosystem as consisting of
―the biota (plants, animals, microorganisms) within a given area, the environment that sustains it, and
their interactions.‖ Thus, the ecosystem approach is holistic rather than species-specific, and focuses on
the maintenance of ecological processes and community mosaics at multiple scales (Ehrenfeld 2000). The
ecosystem, from such a holistic viewpoint, includes ecological processes such as the movement of
individuals as well as the maintenance of sustainable populations across landscapes.
Catalina‘s oak trees occupy a broad spectrum of plant community types, including woodland, riparian,
chaparral, coastal scrub, and grassland. They can comprise the dominant species in some woodland types,
or they can be part of a more diverse riparian or chaparral assemblage. They can be found from the tops
of ridges to the edges of streams, and from the shoreline to the most interior slopes (Figure 3). The
wildlife that are found within their influence also commonly depend on adjacent habitat. Therefore, a
single ‗oak ecosystem‘ is difficult to define – yet it is clear that it encompasses a large majority of the
island because of the interspersion of oaks with non-oak species and the gradation of oak-dominated
woodland into other vegetation types. Because oaks are so prevalent on Catalina but the boundaries of the
oak ecosystem are difficult to define, this paper will address the island ecosystem as a whole, with a
strong emphasis on the oaks and their closely associated species.
Setting and island history
Catalina is the third largest of the eight California Channel Islands, and is located approximately 32
kilometers (20 miles) from the southern California coast (Schoenherr et al. 1999). It is 34 kilometers (21
miles) long, and 13 kilometers (8 miles) wide, and has an area of 194 square kilometers (about 48,000
acres). An isthmus only 0.73 kilometers (less than half a mile) in width effectively isolates the western
19% of the island (Figure 4). Elevations range from sea level to 640 meters (2,100 feet) at Mount
Orizaba, near the island‘s center. Catalina is bisected by numerous steep canyons, and is dominated by a
north-west to south-east trending mountain range. Perennial streams are limited to a few dominant
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Figure 1. Distribution of areas dominated by Quercus tomentella (Island oak) and Quercus
pacifica (Island scrub oak) on Santa Catalina Island, California.
Oak ecosystem restoration on Santa Catalina Island, California: Proceedings of an on-island workshop, February 2-4, 2007. Edited by D.A.
Knapp. 2010. Catalina Island Conservancy, Avalon, CA.
139
Drainages
Quercus pacifica dieback
Healthy Quercus pacifica
Shoreline
N
8
0
8
16 Kilometers
Figure 2. Above: extent of a large-scale dieback of Island scrub oak (Quercus pacifica) on
Catalina Island as of 2005. Below: images of the dieback.
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Figure 3. Oaks occupy a diversity of habitat types on Catalina. Photos by Bill Bushing (lower middle
two) and the author.
canyons, including Middle and Cottonwood Canyons, which are the two largest drainages on the island
and are located near its center. Fifteen different vegetation types as defined by Knapp (2005a) are found
on the island. Most are rare and the three most predominant are coastal sage scrub (38%), island chaparral
(29%), and grassland (19%). While Quercus tomentella-dominated habitat is clearly woodland, Quercus
pacifica-dominated habitat can be both scrubby and arborescent, and therefore resembles both chaparral
and woodland; in this paper, oak-dominated habitat will cumulatively be called ―woodland‖ for
simplicity.
Catalina has a Mediterranean climate, with dry, warm summers and cool, wet winters, but, typical of
islands, its high and low temperatures are moderated by the ocean‘s influence (Schoenherr et al. 1999).
Between 1948 and 2005, an average annual (calendar year) rainfall of 29.7 centimeters (11.7 inches) was
recorded across four to six measurement stations across the island (Catalina Island Conservancy,
unpublished data). The minimum annual rainfall during that time was 9.7 cm (3.8 in) and the maximum
rainfall was 72.9 cm (28.7 in).
Human disturbance and land uses are an important part of the ecosystem threats discussed in this paper.
Therefore, a brief human history of the island is provided in the following paragraphs in order to provide
some perspective on how use and management of the landscape has changed over time, and some context
for how long certain changes have been in effect.
There is evidence of Native American settlements on Catalina as early as 7,000 years ago (Moore 2009).
These early residents, who called the island Pimu and themselves Pimugnans, subsisted on marine life
and terrestrial plants (Schoenherr et al. 1999). Based on practices observed on the nearby mainland, they
may have periodically burned the landscape to encourage the growth of grasses and herbs (Timbrook et
al. 1993; Keeley 2006a).
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Figure 4. Santa Catalina Island, California: Place names, habitation, and hydrologic alterations
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Although Juan Rodríguez Cabrillo stepped foot on the island in 1542, he stayed for less than a day
(Moore 2009), and there was little foreign influence between that time and the start of the Spanish
mission period in the late 1700‘s (White and Tice 1997; Moore 2009). The establishment of the missions
combined with the arrival of Russian, American, and Aleut otter hunters in the early 1800‘s caused the
decline of Native American populations due to introduced diseases such as measles, conflict with the
hunters, and active removal by Spanish missionaries (White and Tice 1997; Moore 2009). By the 1820‘s,
no Pimugnans remained on Catalina (Moore 2009).
In the early 1800‘s, Spanish missionaries and Yankee traders brought goats to Catalina (Capra hircus;
Coblentz 1977). As the gold rush hit mid-century, squatters took up residence and began running sheep
and cattle (Moore 2009). Miners arrived in 1863 and extracted some silver, mostly on the west end of the
island (Moore 2009). In 1864, Union soldiers constructed an outpost and barracks at the island‘s isthmus,
as part of a survey for a potential Indian reservation on the island (Moore 2009).
Shortly after the city of Avalon‘s founding in 1887, roads, buildings, and dams were constructed in the
island‘s interior, and remote coves were developed (Catalina Island Conservancy, undated-a; Propst 1997;
White and Tice 1997). A stagecoach road from Avalon to the Isthmus was completed in 1898, which was
extended to the middle of the west end of the island the following year (White and Tice 1997). In 1919,
Catalina was acquired by chewing gum magnate William Wrigley, Jr., and by the 1920‘s, Avalon was a
popular resort town and retreat for the rich and famous (Schoenherr et al. 1999). As visitation increased,
development of the interior of the island (outside of Avalon‘s influence) continued. Wrigley further
developed Pebbly Beach, just southeast of Avalon, and expanded existing rock quarries at Pebbly Beach
and Empire Landing (White and Tice 1997; Guhan 2009). Silver, zinc, and lead mines were established at
Black Jack Mountain and Renton Canyon in the 1920‘s, and the ore was processed at White‘s Landing
(White and Tice 1997). Wrigley‘s son, Philip, established a 1,500-acre horse ranch, El Rancho Escondido,
near Cottonwood Canyon in 1930 (White and Tice 1997). Camp Cactus, a remote outpost between
Eagle‘s Nest Peak and Bulrush Canyon, was occupied by as many as 600 men during World War II
(White 2002). Following the war, Philip Wrigley completed the ―Airport in the Sky‖ near Blackjack
Mountain in 1946 by leveling two mountain peaks and filling three canyons (White and Tice 1997).
These locations and others discussed throughout the document are presented in Figure 4.
Construction of earthen dams on Catalina began as the city of Avalon was being developed (Figure 4;
Propst 1997). The first dams to be constructed (prior to 1924) were in the Haypress area of the island, in
order to provide water to the city (Propst 1997). Additional dams were later constructed to further service
livestock, wildlife, and irrigated fields (Propst 1997). Thompson Reservoir was constructed at Middle
Ranch in 1924; water from this large reservoir, which includes a cement spillway and pump station,
services Avalon as well as other residents (Propst 1997), and was later used to irrigate nearby hay fields,
established in 1953 (Catalina Island Conservancy, undated-a). Cement dams were constructed in lower
Cottonwood and Sweetwater canyons shortly after Thompson Reservoir was built, but these quickly silted
in (Propst 1997). The Sweetwater dam was abandoned, whereas the lower Cottonwood dam and pumping
system still provided water to irrigate horse pastures at El Rancho Escondido (Propst 1997). Other dams
were constructed by the Catalina Rock and Ranch Company (by design of the Soil Conservation Service)
between 1953 and 1972, including those in Bulrush Canyon, middle and upper Cottonwood Canyon,
Cape Canyon, and smaller Middle Canyon tributaries (Propst 1997). Southern California Edison assumed
control of Catalina‘s water and all utilities in 1962; they enlarged Thompson Reservoir and constructed
additional dams in the island‘s interior shortly thereafter (Catalina Island Conservancy, undated-a).
Philip Wrigley and his sister Dorothy founded the Catalina Conservancy in 1972, and today the
Conservancy owns and manages 88% of the island. Its mission is to be a responsible steward of its land
through a balance of conservation, education and recreation. It maintains an easement agreement with Los
Angeles County to allow recreational access to a majority of its land. Farming was discontinued in
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Middle Canyon and a comprehensive restoration program was begun in 1998 (Catalina Island
Conservancy, undated-a), reflecting a more active vision for the island‘s conservation than that of
previous years.
The island‘s official population was 3,696 in 2000, with approximately 85 percent living in the
incorporated city of Avalon and 298 living at the village of Two Harbors (U.S. Census Bureau 2009).
Additional small settlements are found at Middle Ranch, Empire Landing, and El Rancho Escondido
(Figure 4).
BIOLOGICAL RESOURCES
Oak species
Seven species of oaks have been recorded on Catalina. Quercus pacifica (Island scrub oak) is a Channel
Island endemic species with a scrubby to treelike growth form (Figure 5). It is the predominant oak on the
island, currently occupying 4,425 ha.(23% of the island), as shown in Figure 2. Quercus tomentella
(Island oak), infrequent on Catalina, is a tree endemic to the California islands and the rarest oak in the
state (Tucker 1983; Pavlik et al. 1991; Figure 6). It forms dense, but somewhat small stands distributed in
the eastern portion of the island (Figure 7); the largest of these groves is 1.1 hectare (2.7 acres; McCune
2005). The insular populations of this tree are thought to be relicts of an ancient species that once
inhabited much of the mainland (Muller 1967). Many of the Quercus tomentella groves display some
Figure 5. Quercus pacifica can be shrubby to tree-like on Catalina Island. Photos by Bill Bushing (upper
right) and the author.
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Figure 6. Quercus tomentella often forms dense groves (left), but can be scattered and shrublike in difficult growing conditions such as this rocky slope at Twin Rocks. Photos by Jenny
McCune and John Knapp, respectively.
degree of introgression with Quercus chrysolepis (canyon live oak), a species which is widespread on the
mainland (Thorne 1967; McCune 2005). In general, Q. chrysolepis type morphology is more common on
the upper slopes, and Q. tomentella morphology is more common on the lower slopes and canyon
bottoms (Kevin Nixon, Cornell University, presentation at the 2007 workshop).
Three of the island‘s oak species are either rare or unverified, as follows. Quercus douglasii (blue oak)
was reportedly collected early in the 1900‘s by Blakley and identified by N. Mueller (Thorne 1967),
however these herbarium specimens have not been located, nor have the purported trees themselves of
this species (Steve Junak, Santa Barbara Botanic Garden, personal communication). Quercus engelmannii
(Engelmann oak) was recorded by Thorne (1967) as a single low, spreading tree in Bulrush Canyon,
however this individual has not been relocated. Lastly, Quercus lobata (valley oak) was collected at
Renton Mine, east of Avalon, and identified by N. Mueller (Thorne 1967), but this tree appears to no
longer be extant. The infrequent but widely distributed presence of Quercus Xmacdonaldii, a hybrid of
Quercus lobata and Quercus pacifica (Island scrub oak), suggests that Q. lobata was more frequent and
widely distributed in the past. Because Q. Xmacdonaldii is a hybrid with a species no longer extant, little
can be done to increase the extent of this species without introducing individuals from off of the island.
The foci of the research discussed in this paper will be Quercus pacifica and Quercus tomentella. The
reasoning for this is that these species are endemic, ecologically important, and genetically clearly
defined. They are not hybrids, are well-represented on the island, and surviving populations are in
confirmed locations.
Associated flora
Because oaks are part of a wide range of plant assemblages on Catalina, they have an array of plant
species associated with them (Figure 8). Common understory plants in Quercus pacifica-dominated
woodland include goldback fern (Pentagramma triangularis) and maidenhair fern (Adiantum jordanii),
perennial grasses such as leafy bentgrass (Agrostis pallens) and chaparral melic (Melica imperfecta),
annual herbs such as miner‘s lettuce (Claytonia perfoliata) and eucrypta (Eucrypta chrysanthemifolia),
perennial herbs such as California goosefoot (Chenopodium californicum) and bicolored everlasting
(Gnaphalium bicolor), and sub-shrubs such as the Channel Islands endemic Nuttall‘s bedstraw (Galium
Figure 7. Quercus tomentella grove locations, Santa Catalina Island, California. From McCune (2005).
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Figure 8. Common plants associated with oaks on Catalina include, clockwise from upper left:
maidenhair fern, chaparral melic, miner‘s lettuce, Island redberry, lemonadeberry, and Nuttall‘s bedstraw.
Photos by the author.
Figure 9. Rare, endemic plants associated with oaks on Catalina include, clockwise from left:
California dissanthelium, island rush-rose, and Catalina mountain mahogany. Photos by J. Travis
Columbus, Lauren Danner, and the author, respectively.
nuttallii). Common shrub and small tree associates include toyon (Heteromeles arbutifolia),
lemonadeberry (Rhus integrifolia), and island redberry (Rhamnus pirifolia).
Rare and endemic plants are commonly associated with oaks as well, including the island endemic grass,
California dissanthelium (Dissanthelium californicum), the Channel Islands endemic island rush-rose
(Helianthemum greenei), and the critically endangered Catalina mahogany (Cercocarpus traskiae), which
is found only on Catalina. Because these species are most commonly found in association with oaks and
are of conservation concern, these three plants will be discussed below in order to illustrate the variety of
ecosystem threats on the island. They are shown in Figure 9.
California dissanthelium is an annual grass which has only been recorded on Catalina, San Clemente, and
Guadalupe Islands. It had not been collected since 1903 and was presumed to be extinct, until its
rediscovery on Catalina in 2005 (McCune and Knapp 2008). This rediscovery was likely due a
combination of feral goat and pig removal, record rainfall in 2005, and increased botanical exploration of
the island (McCune and Knapp 2008). It was found in seven broadly distributed locations on the island, in
17 populations, and at five of those locations, Quercus pacifica was an associated species (McCune and
Knapp 2008).
Island rush-rose is endemic to Santa Cruz, San Miguel, Santa Rosa, and Catalina Islands (Junak et al.
1995), and is a Federally Threatened species. It is frequently associated with Quercus pacifica on the
island. It germinates following fire, but is degraded or eliminated by deer browsing if unprotected (Knapp
2005b; Sarah Ratay, Catalina Island Conservancy, personal communication).
The endemic Catalina mahogany is found naturally in only one remote drainage on the Catalina Island‘s
south side named Wild Boar Gully. There, it is found in sparse, open chaparral in association with
Quercus pacifica. Only seven pure individuals of this species remain, contributing to its status as
Federally Endangered (D. Knapp 2004). It is a relict species, having likely been more widely distributed
in the warmer, wetter conditions of the past (Axelrod 1958; Searcy 1969). Due to the severe threat of
introduced ungulates and the demonstrated benefits of protective fencing, the entire gully where it occurs
was fenced in 1999. Other risks include low genetic diversity and hybridization with the more common
Island mountain mahogany (Cercocarpus betuloides var. blancheae) (Rieseberg et al. 1989).
Associated fauna
Results from a survey of small mammals, reptiles and amphibians between 2002 and 2004 (Backlin et al.
2005) demonstrate the importance of oaks to the island‘s wildlife. Of the twenty sites sampled, the
highest diversity of small mammals, reptiles and amphibians was found at an oak-dominated site, and of
the the six sites with the highest species richness, five of them had oaks as a dominant plant species. All
twelve of the native vertebrate species captured by USGS‘s survey were found in oak habitat, with four of
the nine reptiles and amphibians captured (garden slender salamander Batrachoseps major, southern
alligator lizard Elgaria multicarinatus, gopher snake Pituophis catenifer, and western ringneck snake
Diadophis punctatus) more abundant in areas dominated by oak trees than any of the other three plant
communities sampled.
Avian abundance and richness are also highest in Island scrub oak dominated habitat, as indicated by land
bird monitoring data collected between 1999 and 2003 from 80 points distributed throughout the island
(Catalina Island Conservancy, unpublished data). Approximately 36% of all bird individuals observed in
those surveys were detected in Island scrub oak chaparral, a two-fold increase over the next-highest use
habitat, coastal sage scrub.
Eight vertebrate wildlife taxa are endemic only to Catalina Island (Catalina Island Conservancy 2004),
and all of these species utilize oaks for shelter, food, or both. This list includes such common species as
the Catalina ground squirrel (Spermophilus beecheyi nesioticus) and Catalina quail (Callipepla
californica catalinensis). It also includes rare endemics like the Catalina Island fox (Urocyon littoralis
catalinae), Catalina Hutton‘s vireo (Vireo huttoni unitti), and Catalina Island shrew (Sorex ornatus
willettii). Additionally, it includes the Santa Catalina Island deer mouse (Peromyscus maniculatus
catalinae) and Santa Catalina Island harvest mouse (Reithrodontomys megalotis catalinae), which are
common and important seed predators on the island, and occupy all of its predominant vegetation types,
including oak woodland (Backlin et al. 2005) Selected endemic vertebrate taxa are shown in Figure 10.
The island also supports 27 known Catalina endemic invertebrate taxa, such as the Catalina walking stick
(Neduba propsti) and Avalon hairstreak (Strymon avalona), as well as six Channel Island endemic species
(Sleeper 1989, updated with information from Backlin et al. 2005). Little is known about these taxa,
however undoubtedly many of them utilize oak woodland and its associated plant species. Selected
endemic invertebrates are shown in Figure 11. The Catalina Island fox and shrew will be discussed in
more detail below, as endemic species for which the most is known.
Figure 10. Endemic vertebrate taxa associated with Catalina‘s oak ecosystem include (clockwise,
from upper left) the Catalina quail, Catalina Island deer mouse, Catalina ground squirrel, Catalina
Hutton‘s vireo, Catalina Island shrew, and Catalina Island fox. Photos by Barbara Ezell, Sarah Ratay,
Barbara Ezell, Barbara Ezell, Frank Starkey, and Olivier Born, respectively.
The Catalina Island fox is one of six distinct subspecies of fox found on the Channel Islands (Laughrin
1980). It was listed as Federally Endangered following a 95% decline east of the Isthmus, which began in
1999. This decline was due to the introduction of canine distemper virus, likely transmitted from a
raccoon or domestic dog (Timm et al. 2009). Toxoplasmosis, which can be transmitted by feral cats, also
contributed to their mortality (Patronek 1998; Timm et al. 2009). A combination of translocations, captive
breeding, and vaccinations helped to bring this species back from the brink of extinction (King et al.
2008). On San Clemente Island, arthropods comprise the largest portion of their diet, followed by plants
and vertebrates (Phillips et al. 2007). The fox utilizes a variety of fruits which mature at different times of
year, from such species as manzanita (Arctostaphylos spp.), toyon, and cactus (Opuntia spp.) (Laughrin
1977). Woodland and chaparral habitats dominated by Island scrub oak and other chaparral species
support the densest island fox populations (Laughrin 1977, 1980). Interestingly, Catalina has historically
supported a lower density of foxes than the other Channel Islands (Laughrin 1980).
Figure 11. Endemic invertebrate taxa associated with oaks on
Catalina include (left to right) the Avalon hairstreak and Catalina
Island walking stick. Photos by the author and John Knapp,
respectively.
The Catalina Island shrew has been sighted or captured only fourteen times on the island since 1942,
despite several focused surveys (Aarhus 2005). This endemic shrew is one of the most genetically distinct
and endangered of the seven subspecies of ornate shrew in southern California and Baja California,
Mexico (Maldonado et al. 2001). Following the capture of two individuals in Cottonwood Canyon by the
U.S. Geological Survey (Biological Resources Division) in 2002, the Conservancy initiated more
intensive sampling for the shrew in maritime cactus scrub, chaparral, and riparian habitats (Aarhus 2005).
It was found to be closely associated with mesic riparian sites, some of which contain Island scrub oak,
and particularly utilizes lower Cottonwood and Shark Harbor drainages (Aarhus 2005). The leaf litter
produced by oaks provides ideal foraging habitat for this insectivore; the trees and their associated species
also appear to provide critical nesting cover and protection from predators (Aarhus 2005). Degradation of
the island‘s mesic habitats by feral ungulates and predation by feral cats are likely important factors
limiting the abundance of this species (Aarhus 2005).
The orange-crowned warbler (Vermivora celata sordida), a near-endemic to the Channel Islands (it also is
found on the Palos Verdes Peninsula), is an example of a species that depends on Catalina‘s oak habitat
(Yoon et al. 2008). This tiny green bird is confined to wooded habitats with a dense understory, and eats
primarily insects, some fruit and nectar (Scott Sillett, Smithsonian Institution, presentation at the 2007
workshop). They conduct more than 85% of their foraging on oaks during the breeding season, where
insect biomass has been found to be highest (Yoon et al. 2008). The Catalina Hutton‘s vireo (Vireo
huttoni unitti), a Catalina endemic species which is also a California Species of Special Concern, also
prefers oak habitat for foraging (Peterson 1990) and is rare on the island.
Other vertebrates of interest include acorn woodpeckers and ground squirrels. Acorn woodpeckers
(Melanerpes formicivorus) may be a particularly important species in the oak ecosystem, because they are
acorn predators as well as primary cavity-nesters. They excavate holes for nesting and these are later used
by other species (Block et al. 1990). The Santa Catalina Island ground squirrel is another species likely to
be important. The mainland form of this species is an acorn predator, generalist herbivore
(predominantly) and source of food for foxes and raptors, but its burrowing activities provide colonization
sites for plants as well as habitat for species such as the burrowing owl (Snow 1973; Lidicker 1989;
Schiffman 2007; Van Horne 2007). It is also important in distributing and burying acorns, some of which
escape predation and germinate (Griffin 1971). Although ground squirrels have been sporadically
captured in previous trapping efforts on the island, little is known about their distribution, abundance, and
population dynamics.
Catalina supports at least eleven native species of reptile and amphibian (Backlin et al. 2005), more than
any of the other California Channel Islands (Schoenherr et al. 1999). Although the California mountain
kingsnake (Lampropeltis zonata) has been recorded at least twice on the island, it is unclear whether this
species is native or an escaped pet (Backlin et al. 2005). The two rarest species are the two-striped garter
snake (Thamnophis hammondii) and arboreal salamander (Aneides lugubris), both of which utilize oak
habitat. The two-striped garter snake is a semi-aquatic species, and has been observed in only two
locations where water flows permanently: the lower reaches of Cottonwood Canyon (Brown 1980;
Backlin et al. 2005) and the mouth of Big Springs canyon at Little Harbor (Deb Jensen, formerly Catalina
Island Conservancy, personal communication). This species was not captured during surveys from 20022004, and is at risk of predation from introduced bullfrogs and feral cats (Backlin et al. 2005).
The arboreal salamander has only been collected on the island once, at Middle Ranch in 1941 (Hilton
1945, in Backlin et al. 2005). In California these salamanders live in moist places (mainly oak woodland),
are active primarily when soil moisture is high, and forage at night during wet weather (Stebbins 1985).
Their elusiveness on Catalina could very likely be because appropriate survey techniques for this species
have not been utilized in recent years. Because they are found under logs and the bark of dead trees, in
rock crevices, and in tree hollows, the most successful survey technique is to examine such locations at
night (Stebbins 1985).
THREATS TO THE ISLAND’S OAKS
Quercus pacifica
An analysis of historical aerial photographs reveals that oak stands dominated by Quercus pacifica have
declined by as much as 31 percent over the past 60+ years (D. Knapp, this volume). This indicates two
issues of concern: the widespread death of adult trees, and insufficient regeneration to replace those trees.
These will be discussed in turn below.
There are a variety of potential explanations for the dieback of mature Island scrub oaks on the island,
including disease, senescence, altered disturbance regimes such as fire and herbivory, and decreased
water availability. These will be addressed below.
In an outplanting experiment designed to explore the limiting factors to Q. pacifica regeneration on the
island, Stratton (this volume) did not find a significant difference in seedling survival between dieback
and healthy areas, suggesting that whatever is causing the dieback is only affecting older trees. In
addition, site-specific soil characteristics do not appear to be causing the dieback phenomenon (Stratton,
this volume). Although native oak root rot fungus (Phytopthora sp.) was detected on many of the dead
trees (Jerry Turney, Los Angeles County Agricultural Commissioner‘s office, personal communication
2002), this is unlikely to be their primary cause of death (Tedmund Swiecki, Phytosphere Research,
personal communication 2002). In Australian Eucalyptus forests, increased fungal pathogenicity and
insect herbivory appear to be the result of changed soil conditions and disturbance regimes, which stress
the roots of mature trees (Jurskis 2005); such may be the case here.
The dieback of Q. pacifica stands has been a decades-long, gradual process (D. Knapp, this volume),
suggesting that senescence (old age) is part of the cause. Interestingly, 12 years of intensive research on
dieback of ‗ohi‘a trees (Metrosideros polymorpha) in Hawai‘i‘s rainforest led to a similar conclusion
(Mueller-Dombois 1985). In order to determine what role senescence plays in the dieback of Q.pacifica, it
would be helpful to know the age distribution of the dead and dying groves. Are the dead trees all near the
maximum age found by de Gouvenain and Ansary (103 years)? Did the individuals in the dieback groves
all establish around the same time?
If the dieback groves are relatively even-aged, under what conditions did they establish? One possibility
is the occasions when a good masting year is followed by an exceptional year for germination and growth.
Another possibility, however, is fire, as even-aged stands of oaks can be created through mass post-fire
regeneration (Bartolome et al. 2002). In three Catalina burns over the past decade, however, post-fire
resprouting by Q. pacifica has been limited by both mule deer browsing and pre-burn dieback (Knapp
2008, personal observation). Even for unburned oaks, it is likely that repeated browsing of resprouts has
greatly reduced the below-ground carbon storage required for further growth (Langley et al. 2002;
Ramirez et al. 2008), therefore limiting their life span.
The adult trees likely utilize groundwater, as do other oaks in California (Lewis and Burghy 1964; Griffin
1973; Ogden 1975; Thomas 1980). They may thus be negatively affected if groundwater levels have been
reduced by water extraction or diversion for human use, which is likely given that wells are the primary
source of water for Catalina‘s 4,000+ residents and one million+ visitors a year. Recovery of a riparian
oak forest following levy removal (Bossard and Randall 2007) indicates that other forms of hydrologic
alteration (discussed in more detail later in the document) may be important stressors to oaks as well.
It is currently unclear which of the above factors is most important in the decline of adult Q. pacifica
observed over the last century. Spatial modeling of dieback data would provide further insight into the
relative causes of the dieback by identifying patterns associated with topographic position and other
environmental factors such as soil type and texture.
De Gouvenain and Ansary (this volume) conducted a demographic study of ten Q. pacifica stands
distributed throughout the island. Using both tree-ring and census data, they used matrix projection
modeling to estimate the growth rate of each population and thus determine if regeneration is sufficient to
replace existing stands. Their results suggest that tree survival and growth are not sufficient to ensure long
term population maintenance in at least two of the ten stands. It should be noted, however, that
regeneration rates were likely underestimated because resprouts from the root crown were not included in
the study (de Gouvenain and Ansary, this volume). The researchers were unable to determine whether
low regeneration is due to poor habitat conditions for germination, or low survival and growth rates.
De Gouvenain and Ansary‘s data (this volume) indicate that Q. pacifica individuals have established at
various times over the past century, despite the presence of introduced ungulates. However, as noted
earlier, this regeneration may not be enough to replace existing stands. Browsing by mule deer was found
to significantly reduce seedling survival and height by the third year following planting (Stratton, this
volume). In addition, sapling-sized resprouts in a burn near Empire Landing exhibited significantly
reduced survival (mean 12% reduction one year post-fire) and height (mean 44 cm [22%] reduction by
one year post-fire) when unprotected from mule deer browsing (Knapp 2008). Deer browsing poses a
greater threat to oak survival as seedlings grow taller (Tyler et al. 2008; Stratton, this volume), which
explains why Manuwal and Sweitzer (this volume) did not find a significant effect on the survival of
young oak seedlings in a two-year Catalina mule deer study. The latter authors did find, however, that
trampling by bison was a significant threat to Q. pacifica regeneration at the seedling stage.
Non-native annual grasses, including bromes (Bromus spp.), wild oats (Avena spp.), rye grass (Lolium
spp.), barley (Hordeum spp.), Arabian or Mediterranean grass (Schismus spp.), and fescue (Vulpia spp.),
are widespread in California and are responsible for a variety of severe ecosystem impacts (see discussion
under ―non-native annual grasses‖ later in this document). As a group, they are found in nearly all
vegetation types on the island, particularly mature stands of Quercus pacifica (unpublished data; Figure
12). These grasses have been implicated in poor recruitment for at least two oak species in California
Figure 12. Dense infestations of introduced annual grasses inhibit oak
regeneration. Photos by Jenny McCune (left) and the author (right).
(Danielsen and Halvorson 1991; Davis et al. 1991; Swiecki et al. 1997; Gordon and Rice 2000). On
Catalina, Manuwal and Sweitzer (this volume) found that the presence of non-native, annual grasses was
one of the most significant variables hindering oak seedling success. In addition, oak planting sites at
which annual grasses (primarily) and other species were treated with herbicide had higher germination
and seedling survival (Stratton, this volume).
Moisture may also be limiting oak regeneration on the island. This hypothesis is supported by the greater
success of seedlings in three situations. Firstly, Island scrub oak seedlings exhibited greater height and
higher survival when they were protected from herbivory with tree shelters than when they were protected
by fencing (Stratton, this volume). These tree shelters provide greater moisture and more shade (and tend
to have higher temperatures and CO2), which may favor oak establishment (McCreary 2004). Secondly,
acorn germination and seedling survival were greater at sites where annual grasses were controlled and
soil moisture was consequently higher (Stratton, this volume). Lastly, acorn germination and seedling
survivorhip were significantly higher in open areas than at the canopy edge, which had somewhat greater
soil moisture (Stratton, this volume). This higher moisture may be due to fog drip, which Evola and
Sandquist (this volume) found to be higher in open locations. Greater oak success in the open may also be
due to higher light levels, however. In contrast to these findings that oaks perform better with greater soil
moisture, acorn germination and seedling survival did not differ significantly between oak dieback areas,
which had significantly higher soil moisture, than corresponding sites where oaks were healthy (Stratton,
this volume).
Oak germination and survivorship were just as high or higher in eroded and open areas as they were
adjacent to existing oak canopy (Stratton, this volume), yet oak seedlings are almost never observed
occurring naturally in such exposed areas. This suggests that acorn dispersal and/or burial may be a
limiting factor to oak regeneration on the island (Stratton, this volume). Although ground squirrels are an
important means of acorn dispersal (Griffin 1971), they are not accompanied on Catalina by scrub jays,
other typical oak woodland inhabitants which are important and efficient acorn dispersers (Scott 2010).
Spatial modeling of Island scrub oak data revealed a positive association between juvenile Quercus
pacifica abundance and intermediate canopy cover (Franklin and Knapp, this volume). Sites with
intermediate canopy cover would appear to provide a greater source of acorns than the open sites, yet
more light and moisture than dense stands.
The various limiting factors to oak regeneration on the island identified here likely vary in importance by
life stage. For example, in a study of recruitment in two mainland oaks, Tyler et al. (2008) found that
yearly weather conditions were most important to germination and survival in the first six months, while
site, mammal browsing, and cattle grazing became more important to oak survival over time. The sapling
stage has been identified as particularly limiting for regeneration of other oak species in California
(McClaran 1986; Davis et al. 1991; McCreary 2004). At the same time, it was determined to be of
relatively high importance in determining population growth rates (de Gouvenain and Ansary, this
volume). Because of these characteristics, it would be of considerable interest to distinguish between the
seedlings versus sapling stage at each of the stands in de Gouvenain and Ansary‘s study (this volume), to
determine if the sapling stage is limiting the Quercus pacifica regeneration.
Quercus tomentella
A census and survey performed in 2003 and 2004 (McCune 2005) confirmed seven general locations of
Quercus tomentella on the island, in a total of 95 groves (a grove was defined as a group of trees whose
canopies were not separated by more than 15 meters). A grove in Bulrush Canyon shown in historical
photographs (Millspaugh and Nuttall 1923) is no longer extant. In 2007, a new location with two groves
was discovered by Lauren Danner, John Knapp, and Sarah Ratay, for a total of eight locations and over
1500 individuals (Figure 7). Two of the eight locations (Mount Orizaba area and Swain‘s Canyon)
contain a large number of groves and individuals, while the remainder contain fewer, smaller, more
isolated groves. Two of the seven surveyed locations, at Pebbly Beach and Lone Tree, contained only 10
and 12 individuals, respectively. Of the 95 groves delineated on the island, 62 of them displayed strong Q.
tomentella morphology, two of them appeared to be Q. chrysolepis, and the remainder (31) had hybrid
qualities (McCune 2005).
The greatest threats to the groves noted in the census and survey included deer browsing and root
exposure caused by erosion (McCune 2005). Many of the groves also exhibited notable amounts of dead
wood (McCune 2005). Browsing, particularly on resprouts, was noted at 55% of the groves. Five of seven
of the surveyed groves had relatively low regeneration, with zero to 12 percent of the stems represented
by a seedling or sapling in the vicinity of the grove. Fourty-seven percent of the groves had no seedlings
or saplings at all (McCune 2005). In contrast, 18% of the mature trees had at least one seedling or sapling
associated with them at the largest metapopulation, in the vicinity of Mount Orizaba and Fern Canyon
(McCune 2005).
Despite obvious fragmentation of Quercus tomentella populations, overall genetic diversity for this
species on Catalina was found to be generally high (Ashley et al., this volume). Yet genetic differentiation
of isolated stands indicates that pollination and/or seed dispersal among populations is limited, both on
Catalina and on the other Channel Islands where this species occurs (Ashley et al. 2008).
It appears that overall, the largest threats to Quercus tomentella are small population sizes, fragmentation
(and the resulting genetic isolation), deer browsing, and root exposure caused by erosion. This species
appears to have a wide environmental tolerance despite a relatively fragmented current distribution.
Twenty nine percent of the island was predicted to be suitable habitat for Quercus tomentella on Catalina
on the basis of microclimatic and topographic features in locations that it occupies today (Franklin and
Knapp, this volume). A similarly large area was found to be potential habitat on Santa Rosa Island (48%;
Kindsvater 2006).
THREATS TO THE ISLAND ECOSYSTEM
The major threats to Catalina‘s ecosystem come from four main sources: transformer species, roads,
hydrologic alteration, and increased fire frequency. These threats often interact and facilitate one another;
each will be addressed in turn below. Climate change is also expected to introduce changes to the system;
in California, hotter and generally drier conditions are projected, while sea-level rise will likely
accelerate, and a greater frequency of extreme events such as droughts, floods, and wildfires is expected
(California Natural Resources Agency 2009). These changes will likely exacerbate many of the existing
threats such as fragmentation and invasive, transformer species (California Natural Resources Agency
2009). Climate change will be discussed here only when it is known to affect particular species,
processes, or interactions.
Non-native, transformer species
In this paper, the term transformer will be used to refer to those species which not only disperse widely
and compete successfully (and are thus invaders), but also change the character, condition, form, or nature
of ecosystems (Richardson et al. 2001). These species are often referred to as ‗invasive species‘ and are
the cause of significant habitat degradation and species extinctions (Vitousek et al. 1996; Mack et al.
2000). Their transport throughout the world has been vastly accelerated by human travel and commerce,
and their constant influx has exceeded the adaptive ability of native species. Transformers are second only
to habitat loss as a threat to biodiversity (Vitousek et al. 1997), and can also negatively affect agriculture,
forestry, public health, and recreational values.
Island ecosystems are particularly vulnerable to the effects of transformer species, which have been a
major cause of extinctions on these islands (e.g., Coblentz 1978; Vitousek 1988; Adsersen 1989;
D‘Antonio and Dudley 1995; Atkinson 1989; Schofield 1989; Waithman et al. 1999; Oberbauer 2005).
Island wildlife populations are typically restricted in both size and density (Nogales et al. 2004), and
individuals are naïve to predators (Courchamp and Sugihara 1999; Crooks and Soule 1999; Nogales et al.
2004). Island endemic plants that have evolved without large herbivores have lost defenses against
herbivory, such as chemicals, spines, and tough leaves, and therefore are particularly vulnerable to the
grazing and browsing pressures of introduced herbivores (Atkinson 1989; Bowen and Van Vuren 1997).
Catalina Island has been the recipient of numerous introduced animal and plant species, including many
that are considered among the world‘s top 100 invasive species (i.e., causing the greatest impacts; Lowe
et al. 2000). This includes goats, pigs, cats (Felis catus), black rats (Rattus rattus), house mice (Mus
musculus), European starlings (Sturnus vulgaris), American bullfrogs (Rana catesbeiana), Argentine ants
(Linepithema humile), tamarisk (Tamarix ramosissima), and giant reed (Arundo donax). Catalina hosts
more animal invaders than any of the other California Channel Islands (Schoenherr et al. 1999). This is
undoubtedly due to the greater extent and diversity of Euro-American land use, as well as more extensive
development and human visitation.
Transformer species can have impacts on the native ecosystem through a wide variety of mechanisms,
including predation, parasitism, herbivory, competition, hybridization with native species, spread of
parasites and disease, and alteration of ecological resources, plant reproductive mutualisms, or
disturbance regimes (D‘Antonio and Vitousek 1992; Feare 1994; Mack et al. 2000; Levine et al. 2003;
Tomkins and Poulin 2006; Traveset and Richardson 2006; Kenis et al. 2009). Interactions between
transformers can compound and complicate their impacts, rippling throughout food and pollinator webs,
and involving multiple ecological relationships (such as predator-prey, various types of competition, and
mutualisms). For example, disturbance by pigs has facilitated the spread of transformer plants (Aplet et al.
1991), introduced frugivorous birds disperse transformer plants (Vitousek and Walker 1989), and a
combination of goat browsing and seed predation by rats may inhibit forest reproduction (Pickard 1982).
In another locally pertinent example, Roemer et al. (2002) believe that introduced pigs indirectly led to
the near extinction of the island fox (Urocyon littoralis) on several of the northern Channel Islands, by
providing the food needed for colonization of a non-native predator, golden eagles.
Novel pathogens vectored by introduced species can have devastating consequences (Mack et al. 2000).
Dramatic losses due to such pathogens have, for example, been recorded for oak trees (Meentemeyer et
al. 2008), amphibians (Laurance et al. 1996; Blaustein et al. 2004), birds (van Riper et al. 1986; Lanciotti
et al. 1999), rodents (Harris 2009), and larger mammals (Timm et al. 2009). Although both native and
non-native species can carry parasites and other pathogens, invaders can introduce new taxa to a system
and are a major threat to wildlife conservation (Tomkins and Poulin 2006). Insular species are particularly
vulnerable to introduced parasites and predators, and transmission of ectoparasites from non-native to
native fauna has been demonstrated for several of the Channel Islands (Crooks et al. 2001, 2004).
The effects and status of many of Catalina‘s worst known transformers will be discussed below. What are
perceived to be the most destructive transformer species have been chosen for detailed discussion, in
order to demonstrate their wide range of impacts, the strength of evidence for those impacts, and their
interactions amongst each other, and thus to inform the prioritization of their management. More detail
will be given for those species which remain on the island and for which the most is known.
Ungulates
Much of the management focus on the Channel Islands to date has been on introduced ungulates, or
hooved mammals. On Catalina, the predominant introduced ungulates have included feral goats, feral
pigs, American bison (Bison bison), and mule deer, discussed in turn below. Approximately 620 head of
cattle (Bos taurus), 22,000 sheep, and 15,000 goats were present on the island in 1864, according to a
military census (Catalina Island Conservancy, undated-a), and the goat population may have reached
30,000 in the 1930‘s (Schuyler et al. 2002a). Excessive grazing from the mid-1800‘s into the 1900‘s
caused deterioration of habitat (Coblentz 1980). Sheep and cattle were the first non-native animals to be
eliminated from the island, by the 1920‘s and 1958, respectively (Catalina Island Conservancy, undateda). Additionally, black-buck antelope (Antelope cervicapra) were introduced to Catalina in 1967, as an
attempt to begin a big-game preserve on the island (Catalina Island Chamber of Commerce 2009). These
antelope are rarely seen, and a small herd is thought to be limited to Cottonwood and Sweetwater
Canyons. Because very little is known about these elusive animals, their impacts will not be discussed in
this document. However, future research on this species would be beneficial.
Introduced ungulates can alter ecosystem processes including nitrogen cycling, net primary production,
and fire regimes (Hobbs 1996), facilitate the dispersal of alien plants (Constible et al. 2005; Walter and
Levin 2008), and alter plant species composition (Donlan et al. 2002). Browsing species such as feral
goats and mule deer have reduced the extent of shrubland on the island, which has been replaced with
introduced annual grasses (Minnich 1982; Figure 13). Island scrub oak seedling densities have been found
to be highest in zones of the island where introduced goats, pigs, and bison have been removed the
longest (Manuwal and Sweitzer, this volume). Introduced ungulates have also undoubtedly played a large
part in causing the open understory of much of the island‘s oak stands (Figure 13). The consequent
reduction of structure and diversity has likely reduced the abundance and diversity of native vertebrates,
which favor a dense, shrubby understory (Tietje and Vreeland 1997), as well as insects, whose diversity is
correlated positively with increasing vegetation complexity (Lawton 1983). Such structural changes
appear to have been an important factor in the decline of native birds and colonization of non-native birds
in New Zealand (Diamond and Veitch 1981).
Substantial erosion is evident on Catalina (Figure 14), and although some of this may be due to the
erosion-prone nature of the soils (Eric McDonald and Todd Caldwell, Desert Research Institute,
presentation at the 2007 workshop), much of it is also likely due to the grazing activities of introduced
animals (Roger Poff, Professional Soil Scientists of America, personal communication). On nearby Santa
Cruz Island, vegetation cover was reduced to barren ground or converted to grassland by introduced
animals and severe soil erosion ensued (Brumbaugh 1980). Elsewhere, conversion of chaparral vegetation
to grassland has been found to significantly increase the amount of sediment loss in a watershed (Rice et
al. 1969).
Recovery has been shown on the Channel Islands and elsewhere following exclusion, removal, or
reduction of introduced animals, including goats (Parkes 1984) sheep (Wehtje 1994; Klinger et al. 2002;
Figure 13. Feral animals such as goats (Capra hircus), pictured on the upper left, have created a
dense network of trails on the island, and facilitated a transition to non-native annual grassland.
Native shrubs have often been protected only by clumps of prickly-pear cactus (Opuntia littoralis),
right. The remaining shrublands often have reduced density and open understory (lower left). Photos
by (clockwise from upper left) Bill Bushing, the author, and John Knapp.
Figure 14. Evidence of erosion is visible throughout the island. Left, Catalina cherry (Prunus
ilicifolia lyonii) roots are exposed in Silver Canyon; Center, oaks grow in an eroded drainage near
Wild Boar Gully; Right, deep gullies compromise oak habitat on the Channel side of the island.
Photos by (left to right): John Knapp (far left) and the author.
Pinter and Vestal 2005); rabbits (North et al. 1994); pigs (Peart et al. 1994; Chess et al. 2000); cats
(Rauzon et al. 2002); rats (Pascal et al. 2005); and multiple of these species plus cattle, elk, cats, and deer
(Laughrin et al. 1994; Christian et al. 2008; Corry and McEachern 2008; Dow et al. 2008; Drost et al.
2008). For example, greater cover and species richness were observed following initiation of goat, pig,
and bison removal from the west end of Catalina (Laughrin et al. 1994; Figure 15), and a dramatic
increase in Catalina cherry and Catalina Island ironwood regeneration was evident after only a few years
(Figure 16). An increase in native perennial grasses and shrubs, and a decrease in non-native annual
grasses was found on Santa Rosa Island following removal of cattle and reduction of both Roosevelt elk
and mule deer herds, although effects varied by taxon, functional group, and physical environment
(Christian et al. 2008, Corry and McEachern 2008, Dow et al. 2008).
Feral goats
Feral goats (Capra hircus) were present on the island from the early 1800s (Coblentz 1977) until 2003,
when their removal was complete, and their effects are still evident. Their ability to utilize an unusually
wide variety of forage (including both woody and herbaceous species) and their destructive feeding habits
allow them to cause significant ecological damage through denudation, compaction, and gully formation
(Coblentz 1980). Their browsing activities can drastically alter plant cover and composition and reduce
woody species regeneration (Atkinson 1964; Yocum 1967; Sykes 1969; Spatz and Mueller-Dombois
1973; Hamann 1975; Coblentz 1977; Coblentz and Van Vuren 1987; Schofield 1989; de la Luz et al.
2003), as well as facilitate plant invasion (Merlin and Juvik 1992). On Catalina, goat-inhabited areas
supported only 60% of the plant cover of goat-free areas, which also exhibited higher stability; California
sagebrush (Artemisia californica) was virtually eliminated and endemic vegetation was severely impacted
where they were present, with consequent impacts to native wildlife such as snakes, quail, and Island
foxes (Coblentz 1978, 1980). In addition, goats virtually eliminated the mulch layer on the soil surface,
which is important for nutrient cycling and slope stability (Coblentz 1980). In the Galápagos Islands,
goats were responsible for transforming forest and scrub into open grassland vegetation (Hamann 1975)
and contributed to the decline of the Galapágos giant tortoise Geochelone elephantus (MacFarland
1974a,b; Daly and Goriup 1987 cited in Courchamp et al. 2003). In Hawaii, they greatly altered plant
composition (Spatz and Mueller-Dombois 1973), and their competition led to the exclusion of the
endangered nene goose (Branta sandvicensis; Yocum 1967). On Guadalupe Island, goats have threatened
numerous native plant species, including Quercus tomentella, which exhibited low regeneration, trunk
damage, and erosion (de la Luz et al. 2003).
Feral pigs
Feral pigs (Sus scrofa) were introduced to the island in the mid 1930‘s (Baber and Coblentz 1986). They
reduce plant biomass by rooting, feeding, and trampling vegetation and inhibit the regeneration of woody
species, and directly consume wildlife such as invertebrates and amphibians (Breuer 1987; Drake and
Pratt 2001; Sweitzer and Van Vuren 2002; Carroll et al. 2003). Acorns are a preferred food (Baber 1985;
Breuer 1987), which combined with their rooting has been found to decrease oak regeneration (Groot
Bruinderink and Hazebroek 1996; Loggins et al. 2002; Sweitzer and Van Vuren 2002). Their wallowing
and rooting create disturbed areas (Figure 17), which were found to cover 65% of the land in areas of
central and northern California with high pig densities (Sweitzer and van Vuren 2002). This disturbance
may encourage erosion and invasion of non-native plant species (Breuer 1987; Aplet 1991; Cushman et
al. 2004; Kotanen 1997; Tierney and Cushman 2006) and alter nutrient cycling (Singer et al. 1984;
although see Moody and Jones 2000). Feral pigs have contributed to the extinction of numerous oceanic
island species (Waithman et al. 1999). The majority of the feral pigs on Catalina were eradicated by the
end of 2004, although at least one animal may still be at large.
Figure 15. Increased vegetation cover and plant species richness were evident within a few years
following initiation of goat, pig, and bison eradication on the west end of the island. Shown above is
Laughrin et al. (1994)‘s west end monitoring transect 14 in (left to right) 1990, 1992, and 1996.
Photos courtesy of Lyndal Laughrin.
Figure 16. Regeneration of Catalina ironwoods and Catalina cherries was evident within a few years
following initiation of goat, pig, and bison eradication on the west end of the island. Above, Laughrin
et al. (1994)‘s west end monitoring transect 16 (ironwood grove) in (left to right) 1990, 1991, and
1996, demonstrating growth below the browse line. Below, transect 5 (Catalina cherry grove), in
1990, 1992, and 2003, plus (below) new seedlings in 1993. Photos courtesy of Lyndal Laughrin.
Figure 17. Rooting by feral pigs (Sus scrofa) on Catalina created large bare areas prone to
erosion and invasion by transformer plants such as horehound (Marrubium vulgare) and
annual grasses, right. Photos by Bill Bushing.
American bison
American bison (Bison bison; hereafter bison), which evolved in the grass and forb-dominated, gently
rolling plains of North America, were introduced to the rugged, shrub-dominated hills of Catalina Island
in 1924 (Gingrich 1974). Bison are similar to cattle in some ways (Miller 2002), allowing a general
comparison to the effects of cattle grazing. In fact, more than half of the island‘s bison show evidence of
cattle DNA (Carpio et al. 2008). Although cattle have been implicated in the poor regeneration of oaks in
California (Tyler et al. 2007), the effect of livestock grazing varies by soil, climate, and perhaps
topographic position (Corbin et al. 2007; D‘Antonio et al. 2007; Jackson and Bartolome 2007). This
makes local ecological studies important for determining management actions. In addition, bison have
many important ecological distinctions from cattle, including greater use of low-quality forage such as
grasses, frequent movement, and more social behavioral patterns (Wuerthner 2002). Therefore, it is
important to understand the ecology of this species in a local context.
In a two-year study of bison ecology on the island, bison were found to reduce the species richness,
diversity, and cover of the island‘s grassland and riparian communities through grazing, trampling, and
wallowing. They also simplified the structure and reduced the tree diversity of scrub oak chaparral stands
and facilitated the dispersal of non-native plants (Sweitzer et al. 2003; Constible et al. 2005; Figure 18).
Open, relatively level grassland and riparian habitats are used heavily and persistently, resulting in
heavier impacts to those areas. Bison also negatively impact oak seedling survival through physical
damage (Manuwal and Sweitzer, this volume), and reduced the survival of bush mallow (Malacothamnus
fasciculatus) seedlings in a burned area near Empire Landing (Knapp 2008).
Although the bison graze non-native grasses on the island, which may have the positive effect of
controlling them, this effect may be short-lived or negated by consumption of native grasses and forbs
Figure 18. Bison are ideal vectors for plant invaders (left) such as annual grasses and
horehound (Marrubium vulgare). They tend to congregate in favored areas (such as
flat grassland and water sources), creating bare areas with dense piles of dung. Photos
by: John Knapp and the author, respectively.
and concentration of nutrients in preferred areas (Dyer et al. 1996; Sweitzer et al. 2003). The bison
themselves also exhibit nutritional stress and dehydration, and generally poor health (Sweitzer et al.
2003). Project scientists were asked to provide several management recommendations to the
Conservancy; however, the researchers recognized that two years was insufficient to give conclusive
results for vegetation response. At that time, it was decided to reduce the bison herd from over 300 to
~150-200 individuals but not restrict their range, in order to gather additional data before considering
more extensive measures. A third year of additional vegetation data was then found to be consistent with
the initial findings (Sweitzer 2003). Bison herds are currently managed on approximately half of the
island (although multiple animals have returned to the west end of the island).
Mule deer
Mule deer (Odocoileus hemionus californicus) were introduced to Catalina in the early 1930s (Coblentz
1977). This species consumes both oak acorns and seedlings (Griffin 1971, 1979; Menke and Fry 1979),
and may eat acorns exclusively when they are plentiful (Tietje 1990). Mule deer consumed over 90% of
the fallen acorns in one California location (Tietje 1990), thus reducing the oaks‘ regenerative capacity.
Deer in general can reduce plant species richness and cover, hinder woody species regeneration
(particularly post-fire), and negatively impact endangered plants (Okuda and Nakane 1990; Miller et al.
1992; Veblen et al. 1992; Shimoda et al. 1994; Patel and Rapport 2000; Russell and Fowler 2004). These
impacts to the vegetation have consequences for the native wildlife; for instance, reduction of understory
plant biomass and the resulting loss of habitat complexity can increase predation risk for birds (Ruscoe et
al. 2006). The impacts of deer may be especially severe where they have been introduced to islands, as
they have been found to lower species richness, change species composition, endanger rare plants, and
reduce wildlife populations of various trophic levels (Wiles et al. 1996, 1999; Wardle et al. 2001; Potvin
et al. 2003; Allombert et al. 2005a,b; Choinard and Filion 2005; Stockton et al. 2005; Gaston et al. 2006;
Spaggiari and De Garine-Wichatitsky 2006, Husheer 2007). For example, Sitka black-tailed deer
(Odocoileus hemionus sitkensis), were found to reduce plant cover, songbird populations, and
invertebrates on the island of Haida Gwaii, Canada (Stockton et al. 2005; Allombert et al. 2005a, b).
A focused study of the ecology and impact of mule deer on Catalina was conducted between 2005 and
2007 (Manuwal and Sweitzer, this volume); mule deer impacts were also part of a larger study on
limitations to oak regeneration (Stratton this volume). In addition to their effects on Island scrub oaks,
discussed previously, the abundance, height, and health of some oak-associated plant species have been
dramatically reduced by introduced mule deer. Rare island endemic shrubs (Ceanothus arboreus, feltleaf ceanothus, and Dendromecon harfordii, Island tree poppy) have been heavily impacted, to the point
of either complete defoliation below two meters or mortality (Manuwal and Sweitzer, this volume; Figure
19). These endemic shrubs germinated in abundance following fire in several locations, and would likely
have become dominant stand-level components of the oak ecosystem in those locations had they not been
nearly eliminated by deer browsing (Knapp 2005b, unpublished data). Mule deer also increased the
mortality of resprouting toyon (Heteromeles arbutifolia) shrubs – from 12% when they were protected
from browsing to 82% within two years post-fire (Ramirez et al. 2008). Resprouting of the endemic
Catalina ironwood tree (Lyonothamnus floribundus floribundus) was also inhibited by mule deer,
necessitating installation of protective fences to ensure their survival (Knapp 2005b; Figure 20). Seven
other regenerating shrub species were significantly shorter when unprotected from browsing (Knapp
2008; Table 1).
Figure 19. Mule deer (Odocoileus hemionus californicus) impacts to regenerating vegetation
following the Goat Harbor burn of 1999. Clockwise, from upper left: regeneration following fire
when protected with fencing; a close up of that fencing with Island tree poppy flowering; what
happened when a different fence failed and deer entered; and the same exclosure from the above left
photo following removal of the right half of the fence for the mule deer study. Photos by the author
(upper left, lower right) and Thad Manuwal (lower left, upper right).
Figure 20. A year following the Goat Harbor Burn of 1999, endemic Catalina ironwood trees
(Lyonothamnus floribundus floribundus) displayed vigorous resprouts (left) where they had been
fenced to exclude mule deer (Odocoileus hemionus californicus); those resprouts had been browsed to
the ground where the ironwoods were unprotected (right). Photos by Bruce Moore.
Table 1. Height comparison by species for plants in unfenced and fenced plots, Empire Burn
area, Catalina Island, California. Data are from October 2007, one year following the burn.
Student‘s t-tests were performed with JMP version 4.04 (SAS Institute 2001). Asterisks
denote significance.
Species
Regeneration
Type (n)
Height
Fenced (cm)
±SE
Height
Unfenced (cm)
±SE
p-value
Baccharis
pilularis
Encelia
californica
Malacothamnus
fasciculatus
Malosma
laurina
Heteromeles
arbutifolia
Quercus
pacifica
Rhamnus
pirifolia
Rhus
integrifolia
Resprouts (9)
37.0 ±8.3
23.4 ±7.4
0.243
Resprouts (8)
24.2 ±3.4
11.0 ±3.4
0.052 *
Seedlings (65)
24.0 ±3.1
4.2 ±6.4
Seedlings (16)
18.0 ±2.2
4.2 ±3.2
Resprouts (29)
66.5 ±5.2
9.4 ±8.9
Resprouts (94)
56.2 ±2.1
12.4 ±4.1
Resprouts (28)
37.0 ±8.3
5.6 ±5.2
Resprouts (147)
24.2 ±3.4
15.3 ±2.1
<0.001
***
<0.001
***
<0.001
***
<0.001
***
<0.001
***
<0.001
***
The documented impact of the deer to regenerating post-fire vegetation highlights their capacity to greatly
alter the structure and composition of island habitats, to the point of near-complete conversion to an open
landscape dominated by introduced annual grasses and forbs in some locations (Figure 19). Their impacts
are not confined to post-fire vegetation; in early trials performed on the island, Lisa Stratton (unpublished
data) found that restoration of plant communities such as coastal sage scrub and riparian woodland was
severely hindered by the browsing of mule deer. Subsequent restoration projects by necessity
incorporated some form of protection from these introduced ungulates.
Carnivores and rodents
Cats, rats, and mice are intimately tied ecologically, therefore a discussion of one should include a
discussion of the other. They are often found together on islands where extinctions have been
documented, therefore it is difficult to separate their impacts. Together, they are believed to be
responsible for the extinction of several rodents on the Galápagos islands; three of the four remaining
species are only found on islands where introduced cats, rats, and mice do not occur (Dowler et al. 2000).
The introduction of both a predator and its prey species (such as cats and rats) can lead to a phenomenon
called hyperpredation (Smith and Quin 1996). This is where the introduced prey is adapted to generally
escape predation, yet supports the predator population when preferred food (i.e., naïve native species) is
unavailable, thus leading to increased predation on native species. Although cats may limit rat and mouse
populations and thus their impacts (Courchamp et al. 1999; Fitzgerald and Gibb 2001; Donlan and
Wilcox 2008), introduced rodent populations can reach high densities even in the presence of predators,
as seen with stoats (Mustela erminea) in New Zealand (Blackwell et al. 2003). Regardless, the breeding
success of native species would likely be substantially greater with neither cats nor their introduced prey
present (Rayner et al. 2007).
The impacts of these multiple invaders are likely to vary markedly with the physical and biological
aspects of the island they have invaded (Martin et al. 2000; Donlan and Wilcox 2008; Quillfeldt et al.
2008; Angel et al. 2009; Ruffino et al. 2009). Several factors suggest a high risk to Catalina‘s bird
populations: the presence of small and/or endemic landbirds and seabirds, the presence of multiple nonnative mammalian predators including feral cats, and the lack of native rats (Atkinson 1985; Blackburn et
al. 2005).
Feral cats
The first written record of feral cats (Felis catus) on Catalina Island was in 1931, when more than 100
animals were removed in order to bolster Catalina quail populations (Anon. 1931). It is likely that cats
were introduced much earlier, however, with the first western European settlers on the island (Guttilla
2007).
Feral cats have a severe impact on native wildlife, due to their characteristics of being dietary generalists
(Barratt 1997; Fitzgerald and Turner 2000; Read and Brown 2001; Nogales et al. 2004) which continue to
hunt even when satiated (Adamec 1976; Crooks and Soulé 1999; Courchamp et al. 2003). They also have
a high reproductive rate (Griffin 2001), and can reach very high population densities (10-100 times higher
than comparable native predators), which are further augmented by human food subsidies (Liberg et al.
2000). Feral cats are believed to be responsible for the extinction of numerous native species, particularly
on islands (Whittaker 1998; Nogales et al. 2004). Nogales et al. (2004) reviewed the known species
extinctions (either localized or complete) due to feral cats, which includes endemic rodents in the
Caribbean, Galapagos, and Baja CA islands, and reptiles such as the New Zealand tuatara, and iguanas
and skinks in the Fiji and Caribbean Islands. It also includes at least 33 bird species, including endemic
landbirds such as the Stephens Island wren and Socorro Island dove, and endemic seabirds such as the
Guadalupe Island storm petrel. In addition, cats are implicated in the extinction of at least seven mammal
species in Australia (Dickman 1996). These extinctions are understandable when viewing the kill rates of
an individual rural cat (up to 1,127 animals per year [Coleman and Temple 1996]). or the estimated kill
rates of feral cat populations (about 455,119 seabirds per year on Marion Island [van Aarde 1980], 1.2
million seabirds per year on Kerguelen Island [Pascal 1980]), and 3 million petrels per year in the French
Subantarctic Islands [Chapuis 1995]). In some locations, their diet consists of predominantly introduced
animals (Nogales and Medina 2009), however their diet can shift dramatically (Peck et al. 2008).
The impacts of feral cats extend beyond predation; they also may affect native wildlife such as foxes,
owls, and hawks through competition for prey (George 1974; Erlinge et al. 1984; Dunn 1991; Coleman
and Temple 1996). Several pieces of evidence suggest that competition is strong between feral cats and
Island foxes. Feral cat diet overlapped by 80% with foxes on San Nicolas Island, and foxes are absent
there where cat densities are high (Kovach and Dow 1981). On San Clemente Island, overlap in dietary
composition of these two species was 93%, although the two partition prey resources differently (with
foxes relying more on arthropods; Phillips et al. 2007). In addition, there are numerous observations of
cats fighting with island foxes on Catalina (Guttilla 2007).
Lastly, feral cats may introduce disease and parasites to wildlife, pets and humans, and degrade water
quality. For example, cats were shown to be a predominant source of fecal coliform bacteria in an urban
Michigan watershed (Ram et al. 2007), and hookworms and roundworms from cats can be transmitted to
wildlife as well as humans (Anderson et al. 2003; Dubná et al. 2007). Transmitted diseases include rabies,
feline immunodeficiency virus (FIV), feline leukemia virus (FeLV), plague, and toxoplasmosis (Jessup et
al. 1993; Roelke et al. 1993; Patronek 1998; Moore et al. 2000; Conrad et al. 2005; Center for Disease
Control 2008; Miller et al. 2008). In addition, there is concern that cats may be vectors for avian flu
(H5N1) (Rimmelzwaan et al. 2006) and Canine Distemper Virus, which they carry on the Channel Islands
(Clifford et al. 2006). Cats exceeded dogs in the number of reported rabid animals between 2000 and
2006 in the U.S., and rabies can be transmitted to wildlife such as foxes (Urocyon spp.) and bats (Center
for Disease Control 2008). FIV and FeLV are potentially lethal feline retroviruses which suppress a cat‘s
immune system, and can be transmitted to wildlife and domestic cats (Jessup et al. 1993; Roelke et al.
1993). The protozoan Toxoplasma gondii can spread via cat feces, causing a disease called toxoplasmosis
which endangers marine mammals and Channel Island foxes, as well as pregnant women (Dubey and
Beattie 1988; Conrad et al. 2005; Clifford et al. 2006; Miller et al. 2008; Timm et al. 2009).
Guttilla (2007) conducted a study of Catalina Island feral cats between 2002 and 2004. Estimated mean
cat density ranged between 0.8 ± 0.4 cats/km2 and 0.3 ± 0.1 cats/km2 (depending on the type of trap
effort), which when extrapolated to the rest of the island produces an estimate of over 600-800 cats on the
island as a whole (Guttilla and Stapp 2010). Greater than 70% of these animals are associated with
human-populated areas, particularly colonies supported by a Trap-Neuter-Release (TNR) program
administered by the local chapter of the Humane Society (Guttilla and Stapp 2010). Under such a
program, cats are captured, sterilized, and returned to the location where they were captured, where they
are fed daily. The high-density conditions created by such food subsidies likely cause increased impacts
to wildlife and humans (Crooks and Soulé 1999; Schmidt et al. 2007) and foster higher incidence of
disease and parasites (Anderson and May 1979; May and Anderson 1979; Fromont et al. 1998). FIV and
FeLV infection prevalence was found to be much higher for Catalina feral cats (17.2% and 13.1%,
respectively) than reported elsewhere in the United States, while 41% of the animals trapped were found
to contain at least one endoparasite. Catalina cats travel a great distance (>10 km) away from the colony
as well, thus impacting interior wildlife: 26% of radio-collared animals were found to travel distances
exceeding 14 km from their colony (Guttilla 2007).
Riparian vegetation was used significantly more by feral cats during the rainy, breeding season (Guttilla
2007); this habitat is limited in distribution on the Island (Knapp 2005a) and is critical for multiple rare
and endemic species (Guttilla 2007). Native wildlife species on Catalina that are potentially threatened by
introduced cats, based on their habitat and size and the known diet of cats, include endemic species such
as the Santa Catalina Island fox, Catalina ground squirrel, the Catalina Island deer mouse, Catalina Island
shrew, Catalina quail, Catalina Hutton‘s vireo, and Bewick‘s wren (Thyromanes bewickii catalinae), as
well as the two-striped garter snake (Brown 1980; Backlin et al. 2005; Guttilla 2007). Diet analyses are
currently being conducted by Guttilla (personal communication).
Rats
Both black rats (Rattus rattus) and Norway rats (Rattus norvegicus) are found on Catalina. They have
likely been present for at least a century, since both of these species invaded Oceanic islands in the
Pacific during the eighteenth and nineteenth centuries, and are readily transported aboard ships (Atkinson
1985). Rats are opportunistic and adaptable predators which have adverse effects on a wide variety of
native wildlife including seabirds, land birds, small mammals, reptiles and amphibians, and invertebrates,
as well as on plant species and entire plant communities. Those effects have been reviewed by Atkinson
(1985), Courchamp et al. (2003), Towns et al. (2006), Jones et al. (2008), and Harris (2009). Their
combined presence likely also has synergistic effects, which have been little documented (Jones et al.
2008). In addition, rats may increase the effects of introduced predators such as feral cats by providing a
reliable source of food, a phenomenon termed hyperpredation (Smith and Quin 1996).
Rats are particularly detrimental to birds, by predating on eggs, chicks, and sometimes adults (Atkinson
1985; Jones et al. 2008). Rats appear to have been a contributory factor in the major decline or extinction
of at least 72 birds on 23 islands or island groups, 39 of those in association with feral cats and 33 of
those without that association (Atkinson 1985). Seabirds are particularly vulnerable to rat invasion,
especially the families Hydrobatidae (storm-petrels) and Alcidae (auks – incl. murres, auklets, puffins,
guillemots, auklets) (Jones et al. 2008). There is additional evidence that terrestrial birds such as wrens,
pipits, and flycatchers are negatively affected by introduced rats (Innes et al. 1999; references cited in
Towns et al. 2006; Traveset et al. 2009).
Rats are cited as the main cause of extinction for at least ten small mammal species on islands, including
native mice, rats, and bats (Harris 2009). Shrews are negatively impacted as well (Pascal et al. 2005).
Some of the best-documented extinctions include: Christmas Island (Indian Ocean), where black rats
appear to have caused the demise of two native rats, and multiple New Zealand islands, where black rats
are implicated in the loss of the greater short-tailed bat (Mystacina robusta) (Harris 2009). Rats thrive in
small habitat fragments, while endemic rodents decline in such fragments (Ganzhorn 2003).
Reptiles, amphibians, and invertebrates are also adversely affected. Rats are cited as a primary factor in
herpetofaunal extinctions on New Zealand islands, including frogs, tuatara, skinks, and geckos (Towns
and Daugherty 1994), and for the reduction of skinks, geckos, and racers on Caribbean islands (John
1999; Daltry et al. 2001; Bell 2002; Towns et al. 2006). Black rats also appear to have caused the nearextinction of the Lord Howe Island stick insect in Australia (Priddell et al. 2003), and decreased the
abundance and diversity of tenebrionid beetles, spiders, arboreal land snails, and other invertebrates on a
variety of other islands (Palmer and Pons 1996; Hadfield and Saufler 2009; Towns et al. 2009).
As omnivores which can eat virtually any part of a plant, rats can also have a strong impact on plant
regeneration, species abundance and distribution; this in turn has an effect on the wildlife that uses such
vegetation (Courchamp et al. 2003). These impacts can either be direct, through consumption, or indirect,
by reducing the numbers of seed-dispersing birds (Clark 1981). Black rats are thought to have prevented
tree and shrub regeneration on islands in both Australia (Pickard 1982) and New Zealand (Allen et al.
1994), and Pacific rats (Rattus exulans), despite being thought to have less severe impacts than Norway
and black rats (Atkinson 1985; Towns et al. 2006; Harris 2009), have been implicated in the loss of
lowland native forest in Hawai‘i (Athens 2009). Rats may, conversely, enhance the dispersal of some
plant species (Drake and McConkey 2001, cited in Courchamp et al. 2003).
The ecosystem-level consequences of seabird predation and extinction have recently been investigated.
Predation of seabirds on islands invaded by either black rat, Norway rat, or both was found to cause
further ecosystem-level effects by disrupting nutrient subsidies, which caused decreased fertility and soil
microbial CO2 production, and increased plant litter production and decomposition (Fukami et al. 2006;
Towns et al. 2009). Kurle et al. (2008) showed that Norway rats indirectly caused a shift in the intertidal
community of the Aleutian Islands from an algae-dominated to an invertebrate-dominated system, by
predating on marine birds. Rats are also direct predators of intertidal organisms such as crabs and limpets
(Navarrete and Castilla 1993).
Rat eradication efforts have had dramatic benefits for native biota on islands. A local example is the
Xantus‘ murrelet (Synthliboramphus hypoleucus scrippsi), which exhibited increased breeding, hatching,
and nesting on Anacapa Island following the eradication of black rats (Jones et al. 2005; Whitworth et al.
2005); the endemic Anacapa Island deer mouse (Peromyscus maniculatus anacapae) benefited from the
eradication as well (Gellerman 2007). Similarly, black rat control improved the breeding success of the
endangered San Clemente loggerhead shrike (Lanius ludovicianus mearnsi) (Heath et al. 2008).
Particularly impressive is when rare seabirds voluntarily colonize an island following rat eradication, such
as the rare Campbell Island snipe (Miskelly and Fraser 2006) or Cassin‘s auklet on Anacapa Island
(Howald et al. 2005). Shrew, skink, and invertebrate populations have been found to recover following rat
eradication as well (Thomas 2002; Pascal et al. 2005).
Rats carry a multitude of diseases and parasites which can be transmitted to both native wildlife as well as
humans, including bubonic plague, rabies, toxoplasmosis, scrub typhus, leptospirosis, lymphocytic
choriomeningitus, Weil‘s disease, Salmonella bacteria, and intestinal worms (Lever 1994). There is
convincing evidence that disease or parasite transmission has caused the extinction of native rodents on
more than one occasion (Harris 2009). In a local example of this potential, whipworms appear to have
been spread from black rats to native deer mice on Catalina as well as five of the other eight Channel
Islands (Smith and Carpenter 2006).
In a small mammal trapping effort during an ecological study of bison on the island, black rats were
similar in abundance to native deer mice, and lower in abundance than the native western harvest mouse
in both 2001 and 2002 (Sweitzer et al. 2003). Rats were found in oak habitat, but not in grassland or
coastal sage scrub (Sweitzer et al., unpublished data). Rats (species undetermined) were captured in a
variety of habitats across Catalina during fox trapping efforts, predominantly Island chaparral (oakdominated) and coastal sage scrub (Catalina Island Conservancy, unpublished data; Table 2, Figure 21).
The species of greatest concern for rat impacts on Catalina, due to their rarity, include the ornate shrew,
two-striped garter snake, and some seabirds. Rare arthropods such as the endemic Catalina Island
walkingstick (Pseudosermyle catalinae) could also be at risk, although little is known about them.
Mammals such as the endemic Santa Catalina Island deer mouse, Santa Catalina Island harvest mouse,
and Catalina ground squirrel are at-risk from introduced diseases and parasites vectored by rats, as well as
competition. Two species of storm-petrel and six species of auk have been recorded on the island, and all
but one (the common murre) are uncommon, rare, or have only been seen a few times on the island
(Catalina Island Conservancy 2002). Perhaps it‘s no coincidence that Catalina has fewer breeding
seabirds than any of the other California Channel Islands, and has low population densities of both stormpetrels and Xantus‘ murrelets, both of which are particularly susceptible to rat predation (Schoenherr et
al. 1999). The only known breeding location for seabirds on Catalina is an isolated islet called Bird Rock
(Schoenherr et al. 1999).
Table 2. Plant communities where non-native rats (Rattus
spp., species undetermined) were trapped on Catalina Island,
2002-2006. Rats were trapped during an extensive Catalina
Island fox survey which involved 16,729 trap nights across
1,219 trap locations. Species identifications were not
recorded.
Habitat
# Rats
Bare
Coastal sage scrub
Developed
Grassland
Island chaparral
Island woodland
Non-native herbaceous
Riparian
Total
3
12
1
7
13
2
1
3
42
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Figure 21. Incidental trap locations (dark green, n = 42) for non-native rats (Rattus spp.,
species undetermined) 2002-2006 relative to trap effort (in light green). Unpublished data
from fox trapping efforts, n=1219 trap locations and 16,729 trap nights.
House mouse
The house mouse (Mus musculus) likely originated on the Indian subcontinent, but spread to Europe at
least 6,000 years ago (Harris 2009), and is now one of the most widely distributed mammals on earth
(Bronson 1979; Rowe 1981). Although these mice are often closely associated with humans and their
structures, they can also exist as ―feral‖ populations independent of humans, and occupy a wide variety of
habitats, including areas with little or no open water (Bronson 1979; Rowe 1981; Ruscoe 2001; Traveset
et al. 2009). Their success in such habitats is largely dependent on the existing mammal fauna (Rowe
1981). House mice are omnivores, which feed primarily on a diversity of invertebrate and plant material
(Gleeson and van Rensburg 1982; Cole et al. 2000; Ruscoe 2001; Jones et al. 2003). They also have been
known to predate upon birds and their eggs (Mills et al. 2002; Wanless et al. 2007). House mice are more
commonly found in urban areas, fragments, habitat edges, and areas with lower vegetation cover than
other rodents (e.g., Boitani et al. 1985; Laurance and Laurance 1995; Bias and Morrison 2006; Clarke et
al. 2006; Chavez and Ceballos 2009; Idris 2009), and are more dense in areas with greater grass cover and
litter depth (Newman 1994; Brown et al. 2004; Bias and Morrison 2006).
House mice are relatively sensitive to competition from other rodents (Caldwell 1964; Sheppe 1967;
Dueser and Porter 1986), although they do appear to inhibit some other species (Orsini et al. 1982).
Because house mouse densities tend to be suppressed where other rodents and invasive mammals occur,
their impacts to native wildlife have been little noted until recently, when multiple islands have
experienced ecological release of this species upon removal of invasive rats (Innes et al. 1995; Caut et al.
2007; Witmer et al. 2007a; Angel et al 2009) or cats (Holdgate and Wace 1961; Huyser et al. 2000;
Wanless et al. 2007). As they have similar omnivorous habits to rats, whose detrimental impacts have
been well documented, it is likely that their effects have been underestimated (Angel et al. 2009;
Simberloff 2009).
House mice can have substantial impacts on seabirds, invertebrates, and plant communities (Angel et al.
2009). Their effects are particularly apparent on species-poor Marion Island, a sub-Antarctic island where
mice are the only herbivore, and where macro-invertebrates play a particularly important role in
decomposition and nutrient release (Smith and Steenkamp 1990). House mice have been implicated in the
reduced success of seabirds such as petrels, albatross, and shearwaters on multiple islands (Mills et al.
2002; Wanless et al. 2007; Zino et al. 2008). Their impact can be through both direct means, by predation
on eggs and chicks, and indirect means, by artificially increasing populations of key avian predators
(Mills et al. 2002) or consumption of a common prey item such as invertebrates (Huyser et al. 2000).
Seabird reductions can in turn have broader implications for ecosystem function, including reduced
nitrogen fertilization and thus lowered primary productivity (Angel et al. 2009).
Although data are sparse, there is some evidence that house mice impact terrestrial birds such as the
Gough Island bunting Rowettia goughensis (Cuthbert and Hilton 2004; Angel et al. 2009). There is also
some evidence that small mammals and herpetofauna are negatively impacted (Dickman 1992 in Harris
2009; Newman 1994), as well as invertebrates such as an endemic flightless moth Pringleophaga
kerguelensis on Marion Island (Rowe-Rowe et al. 1989; Crafford 1990; Smith and Steenkamp 1990), the
endemic giant weta (Deinacrida rugosa; Orthoptera) in New Zealand (Newman 1994), and the Lord
Howe Island stick insect (Dryococelus australis) in Australia (Carlile et al. 2009). Loss of such species
can have effects on nutrient cycling (Rowe-Rowe et al. 1989), or secondary prey items (Chown and Smith
1993). Intense invertebrate predation also indirectly affects native insectivores such as birds (Angel et al.
2009), or omnivores such as mice, shrews, and foxes.
House mice can impact native plants through either herbivory or burrowing (Angel et al. 2009), and have
the potential to significantly alter a plant community. They have halted regeneration of a sedge (Uncinia
compacta) on sub-Antarctic Marion Island, which is becoming increasingly important on a nearby mouse-
free island in response to climate change (Chown and Smith 1993). They have also caused extensive
damage to an important cushion plant species on that same island, likely with detrimental effects on the
epiphyte species that the cushion plants are nurse plants for (Phiri et al. 2009). They are believed to be
detrimental to Opuntia cactus in the Galápagos (Snell et al. 1994 in Harris and Macdonald 2007), and
together with the absence of a seed-burying native ant, are thought to be limiting the regeneration of cape
myrtle (Phylica arborea, Rhamnaceae) on Gough Island (Breytenbach 1986). In a potentially positive
effect, house mice may be an important consumer of non-native annual grass seed (Borchert and Jain
1978). Together with California voles (Microtus californicus), they reduced the abundance of wild oats
(Avena fatua) and barley (Hordeum leporinum) by 62% and 30%, respectively in a Davis, California
grassland (Borchert and Jain 1978).
House mice can also transmit a multitude of diseases and parasites, including rat-bite fever,
toxoplasmosis, tularaemia, murine typhus, scrub typhusrickettsial pox, leptospirosis (Weil‘s disease),
favus, lymphocytic choriomeningitis, food poisoning by Salmonella bacteria, fleas, roundworms, and
tapeworms (Chinchilla 1978; Rowe 1981; Lever 1994). Although this does not appear to be such a
problem for house mice as it is for rats (Rowe 1981), the introduction of such pathogens could affect
native wildlife in the same way as has been documented for rats (Harris 2009).
House mice have been observed on Catalina in Avalon (Darcee Guttilla, formerly Catalina Island
Conservancy, personal communication), but were not captured in the interior of the island during small
mammal trapping conducted between 2001 and 2002 (Sweitzer et al. 2003), or during pitfall trapping
efforts from 2002 to 2004 (Backlin et al. 2005). This apparent absence in interior regions could be due to
predation by species such as foxes, cats, and rats, or competition with native rodents such as deer mice
(Caldwell 1964; Newsome et al. 1976; Dueser and Porter 1986), or lack of detection.
Birds
Introduced birds have been implicated in the transmission of novel diseases in Hawaii, leading to the
extinction and decline of native birds there (Lever 1994). Diseases which may be spread by introduced
birds as well as mosquitoes include West Nile virus (Kramer et al. 2008), avian malaria, and birdpox
(Lever 1994). Although Catalina supports multiple introduced bird species, three are of particular
concern: the European starling (Sturnus vulgaris), wild turkey (Meleagris gallopavo), and brown-headed
cowbird (Molothrus ater); these will be discussed in turn below.
European starling
European starlings (hereafter starlings) were introduced to New York City in 1890 and 1891 and are now
ubiquitous throughout North America (Cabe 1993). They apparently immigrated to Catalina in the late
1960‘s or early 1970‘s (Jones 1975). They are generalists in habitat, food sources, and roosting sites,
which likely accounts for their success (Cabe 1993; Koenig 2003). Starlings typically favor disturbed,
homogeneous habitat (such as agricultural fields, fragments, and invaded monocultures) (Clergeau and
Fourcy 2005), and there is evidence that their presence is facilitated by introduced ungulates (Diamond
and Veitch 1981). They tend to avoid large undisturbed forest or scrub (Cabe 1993), but in California,
they have been found to utilize open woodlands of gentle topography with well-developed soil (Purcell et
al. 2002). They reach extremely high densities, and are widely considered agricultural pests (Lever 1994;
Clergeau and Fourcy 2005) as well as urban pests (Brough 1969; Garner 1978; Kirkpatrick and
Woolnough 2007; Seamans et al. 2007).
Starlings are secondary cavity nesters, which roost either in cavities excavated by other species such as
woodpeckers, or natural tree cavities (Troetschler 1976). Their impacts may include competition with and
displacement of other birds, particularly other cavity-nesting species such as woodpeckers and flickers,
facilitation of other transformer species, and introduction or spread of pathogens to other wildlife as well
as humans. These impacts are described below.
Starlings appear to have the greatest negative effect on some members of the woodpecker family. For
example, they reduced the fecundity of red-bellied woodpeckers (Melanerpes carolinus) in Mississippi by
usurping 52% of their nest cavities (Ingold 1989), while northern flickers (Colaptes auratus) produced
fewer nestling and fledglings in Ohio in their presence due to nest cavity loss and delayed clutch
completion (Ingold 1994, 1996, 1998). Starlings may also be the cause of a decrease in sapsuckers
(Sphyrapicus spp.) throughout North America (Koenig 2003) and a decrease in the number of nesting
Gila woodpeckers (Melanerpes uropygialis) in Arizona (Kerpez and Smith 1990).
For other woodpeckers, negative effects have not been observed. For example, breeding pairs of the
yellow-shafted flicker (Colaptes auratus auratus) and red-headed woodpecker (Melanerpes
erythrocephalus) did not appear to be negatively affected by starling introduction and increases in the
Midwest between the 1920‘s and 1970‘s (Troetschler 1976). And although starling attacks displace acorn
woodpeckers and elicit behavioral changes which likely have adverse effects, this woodpeckers‘
fecundity and group size do not appear to be adversely affected, perhaps because they are able to delay
their breeding season in order to avoid competition with starlings (Troetschler 1976). Nation-wide, acorn
woodpeckers do not show starling-associated declines (Koenig 2003). Starlings were also not found to be
a major factor in the decline of Lewis‘ woodpeckers (Melanerpes lewis) in Colorado (Vierling 1998).
Many other bird species that are not in the woodpecker family may be impacted by starlings, however
reports vary greatly. In the most damning account, complete nest displacement by starlings was observed
on one property in Nevada for a variety of native bird species, including American kestrels (Falco
sparverius) mourning doves (Zenaida macroura), house finches (Carpodacus mexicanus), and Northern
flickers (Colaptes auratus); these species only resumed breeding following starling eradication (Weitzel
1988). Starlings also predate heavily (73.1 to 90.2% population reduction) on terns on Santa Maria Island
in the Azores (Neves et al. 2006). Other data are more equivocal. In an analysis of long-term data sets for
27 different native cavity-nesting bird densities throughout North America, Koenig (2003) found that,
although nest-site competition undoubtedly displaces multiple native bird species, no species outside the
woodpecker family showed a decrease potentially attributable to invasion by starlings. Similarly,
abundance and breeding success of a variety of cavity-nesting birds (including flycatchers, woodpeckers,
and warblers) was not affected by European starlings in Arizona riparian habitat, where nest sites were
not limiting, despite interspecific aggression (Brush 1983). Populations of some native birds like the
purple martin (Progne subis) and eastern bluebird (Sialia sialis) have been in decline, and although the
starling undoubtedly competes with them, the impact of this competition has not been convincingly
established (Jackson and Tate 1974; Brown 1981; Green 1983; Zerhusen 1984, 1994; Pell and Tidemann
1997; Airola and Grantham 2003).
Starlings may facilitate other transformer species, such as invasive plants and introduced ungulates
(Murphy 1981; Guix et al. 2001; Kaellander 2004). They appear to preferentially disperse seed of nonnative plant species over natives (Ferguson and Drake 1999; Lafleur et al. 2007) the germination of which
may be enhanced by passage through the starling gut (Smith 1981; Clergeau 1992). Symbiotic
interactions have also been observed between starlings and deer (Odocoileus virginianus and O.
hemionus) (Murphy 1981) or cattle (Kaellander 2004), where the birds forage for ectoparasites from the
mammals. Sweitzer et al. (2003) found that starlings were associated with at least 52 groups of bison
(3.7% of all groups) on Catalina, typically those groups that were large and associated with water sources
and human habitation.
Lastly, there has been some concern about starling transport of pathogens. Starlings are a potential
reservoir for avian influenza virus (H-7 N-7) (Lipkind et al. 1982) and West Nile virus (Loss et al 2009),
and vector and enhance the growth of a disease-causing fungus, Histoplasma capsulatum (Garner 1978).
In 1968, shortly after their introduction to Catalina, starlings were reported to be common at the Isthmus,
and present in areas farther from human disturbance but still associated with it (Jared Diamond, letter to
the Catalina Island Conservancy, May 7, 1968). The locations that Mr. Diamond lists are reflected in
more recent island surveys, during which distributed among 19 out of 80 different monitoring locations
near centers of human activity (Catalina Island Conservancy, unpublished data; Figure 22). They were
found in nearly every habitat type sampled, but were particularly abundant around power lines and fences,
and non-native trees such as eucalyptus and palms. Starlings are considered relatively abundant on the
island (Catalina Island Conservancy 2002). Cavity nesting species on Catalina which could potentially be
impacted by starlings include the acorn woodpecker, Bewick‘s wren (Thryomanes bewickii catalinae),
northern flicker (Colaptes auratus), northern saw-whet owl (Aegolius acadicus), and American kestrel
(Falco sparverius). Northern flickers are the most likely to be at-risk (Ingold 1994, 1996, 1998; Koenig
2003).
West End Coves
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Figure 22. European Starlings observed (n = 270) during land bird monitoring efforts on
Catalina Island, 1999-2007. Size of dot represents relative abundance.
Wild turkey
Although native to other parts of North America, the wild turkey (Meleagris gallopavo, multiple
subspecies) is not native to California (Eaton 1992). After drastic North American declines in the 19th and
20th centuries, it was introduced beyond its native range by game management officials and has flourished
in areas with suitable habitat (Eaton 1992). The Rio Grande subspecies of wild turkey (M. gallopavo
intermedia) was introduced to Catalina by the California Department of Fish and Game between 1959 and
1969 (Harper and Smith 1970). Wild turkey impacts to California‘s native flora and fauna have not been
studied, and the Department of Fish and Game has recently agreed to a moratorium on further turkey
introductions in California (Gluesenkamp 2003). However, important impacts are suspected
(Gluesenkamp 2003). After introduction to Santa Cruz Island in 1975, the wild turkey population erupted
in the early 2000‘s, potentially due to vegetation recovery following feral ungulate removal (Morrison
2007). Land managers have since eradicated all unsterilized wild turkeys from that island (Morrison
2007).
The wild turkey roosts in trees at night but forages by day on ground, and generally prefers mixed forest
with trees such as oaks and pines (Eaton 1992). Suitable turkey habitat also includes chaparral and
woodland interspersed with grassland (Harper and Smith 1970). Its diet mainly consists of a wide variety
of plants, but also includes invertebrates, reptiles, and amphibians (Eaton 1992). It consumes such plant
genera as Quercus, Arctostaphylos, and Ribes (references in Gluesenkamp 2003), which contain species
both dominant and endemic on Catalina. In fact, acorns comprise between 12% and 53% of a turkey‘s
diet by volume (mean 26%), based on eight data sets from four U.S. states; acorns also appear in 13% to
94% of gut or dropping samples (mean 57%) (Sanderson and Schultz 1970).
Although the wild turkey has been known to reside in the vicinity of Thompson‘s reservoir at Middle
Ranch, it has not been sighted on Catalina in recent years (Darcee Guttilla, formerly Catalina Island
Conservancy, personal communication).
Brown-headed cowbird
Brown-headed cowbirds (Molothrus ater), like the European starling, are widely naturalized and are
favored by human-modified habitat, preferring fragments, edges, agriculture, pasture, and grasslands with
low or scattered trees (Lowther 1993). The brown-headed cowbird‘s most distinguishing feature is that it
is a brood parasite, meaning that it lays its eggs in other species‘ nests, which then raise the young
(Lowther 1993). This may have a substantial impact on rare birds if they are heavily parasitized, therefore
this species is considered a major conservation concern (Morrison et al. 1999). The species which the
brown-headed cowbird most commonly parasitizes include warblers, sparrows, and vireos, among others
(Lowther 1993). The brown-headed cowbird is facilitated by ungulates, particularly bison, which flush
insects for it to prey upon (Lowther 1993). It spread from its native range in the great plains as forests
were cleared (Lowther 1993). It eats primarily grass seeds, grasshoppers, and beetles (Lowther 1993).
A total of four brown-headed cowbirds were identified during land bird surveys between 1999 and 2007
(Catalina Island Conservancy, unpublished data). Those birds were found in two locations, in bare and
human-associated areas at El Rancho Escondido and Emerald Bay. Cowbirds were found associated with
21 out of 1,419 (1.5%) groups of bison on Catalina, typically larger groups associated with water or
human habitation (Sweitzer et al. 2003). On Catalina, this bird has the potential to impact such island
endemic birds as the orange-crowned warbler and Catalina Hutton‘s vireo, both of which are strongly
associated with oak trees. Although the orange-crowned warbler is considered relatively abundant on the
island, the Catalina Hutton‘s vireo is uncommon (Catalina Island Conservancy 2002). Hutton‘s vireos are
frequently parasitized by the brown-headed cowbird on the mainland (Stokes and Stokes 1996).
Amphibians
American bullfrog
The American bullfrog (Rana catesbeiana; hereafter bullfrog) is native to eastern North America, but is
now distributed around the world, after being both purposely and accidentally introduced (Lever 2003;
Adams and Pearl 2007). Bullfrogs are highly fecund, can reach high population densities, and have a wide
ecological tolerance (Bury and Whelan 1984; Schwalbe and Rosen 1988; Lever 2003); in addition, adults
can disperse long distances (>1200 m; Willis et al. 1956 in Adams and Pearl 2007). They have a broad
diet, opportunistically preying on a wide variety of aquatic to terrestrial species (Lever 2003). The prey of
this rather large frog include: amphibians such as other frogs and salamanders; snakes; invertebrates such
as crustaceans, insects, worms, and snails; birds as large as robins and orioles; small mammals (including
bats, shrews, and mice); and fish (Bury and Whelan 1984; Orchard 1999; Lever 2003).
Bullfrogs thrive in human-modified habitats such as impounded streams, reservoirs, and stock ponds with
reduced riparian vegetation, where they have a competitive edge over some native amphibians (Moyle
1973; Sartorius and Rosen 2000; Kiesecker et al 2001). They also appear to be favored over native
amphibians by disturbances such as trampling of stream banks by cattle, clear-cutting of forests, and
excess sediment deposition from roads and construction (Moyle 1973; Patrick et al. 2006). Non-native
predatory fish such as smallmouth bass (Micropterus dolomieui) and bluegill (Lepomis macrochirus)
facilitate the presence of bullfrogs, while also having their own impacts on native freshwater fauna
(Hayes and Jennings 1986; Kiesecker and Blaustein 1998; Sartorius and Rosen 2000; Adams et al. 2003;
Boone et al. 2007). Indeed, it is difficult to separate the effects of bullfrogs, non-native predatory fish, and
human disturbance or habitat modifications, but they each contribute to native amphibian declines and
often facilitate one another (Moyle1973; Fisher and Shaffer 1996; Sartorius and Rosen 2000; Kiesecker et
al. 2001; Adams et al. 2003; Porej and Hetherington 2005).
Bullfrogs are believed to be at least partly responsible for the decline or extinction of seven native frogs
(Rana spp.) in the southwestern United States, many of them endemic species (Lever 2003). For example,
introduced bullfrogs, combined with habitat degradation, are implicated in both the disappearance of the
red-legged frog (Rana aurora draytonii) and the range restriction of the yellow-legged frog (Rana boylii)
in the San Joaquin Valley of California (Moyle 1973). Introduced bullfrog impacts have been reported in
regions outside of the United States as well (Licht 1972, 1974 in Orchard 1999; Hecnar and M‘Closkey
1997; Lever 2003; Boone et al. 2004). For example, bullfrogs have been charged with eradicating all the
native frog and toad species in the Sonso Reserve in Columbia, and with exterminating multiple endemic
amphibians in a Venezuelan lagoon (Lever 2003). These effects are likely due to a combination of
competition and predation (Moyle 1973; Werner et al. 1995; Kiesecker and Blaustein 1997; Kupferburg
1997; Kiesecker et al. 2001; Wu et al. 2005). In the case of the red-legged frog, over-harvesting is likely
to have also been an important factor in the decline (Jennings and Hayes 1985).
Snake populations may be at-risk from bullfrogs as well. Bullfrogs are believed to be responsible for the
extinction or decline of the Mexican garter snake Thamnophis eques (Rosen and Schwalbe 1995, in
Kupferberg 1997 and Lever 2003), and their predation resulted in a 20% loss of recently hatched giant
garter snakes (Thamnophis gigas) in north-central California (Wylie et al. 2003).
Introduced bullfrogs may interact with other factors such as fungal disease, climate change, and
environmental contamination to cause catastrophic declines in amphibian populations (Blaustein et al.
2004; LaMarca et al. 2005). Indeed, bullfrogs can themselves introduce the causal agents of diseases such
as chytridiomycosis, which is implicated in global amphibian declines (Hanselmann et al. 2004;
Schloegel et al. 2009), or schistosomiasis, a parasitic disease which infects humans (Lever 2003). An
increase of years with low rainfall, as is predicted to occur in California, will be an important source of
stress to amphibian populations (Daszak et al. 2005).
Bullfrogs were introduced to Catalina in 1953 from the California mainland (Bushing 2000 in Lever
2003). They appeared at the Cottonwood Canyon Reservoir in 1977, and were present in Thompson‘s
Reservoir, Middle Canyon, some time before that (Brown 1980). They have been observed by the author
both adjacent to and below that reservoir in recent years, and both historical and current herpetological
surveys also confirm the bullfrog‘s presence at Haypress Reservoir (Backlin et al. 2005). Non-native fish
such as bluegill are stocked by local residents at reservoirs such as Thompson and Haypress (John Knapp,
formerly Catalina Island Conservancy, personal communication). The bullfrog has also been observed in
the upper Cottonwood watershed (John Knapp, personal communication), and is widely believed to
occupy reservoirs throughout the island (Juanita Constible, formerly University of North Dakota; Darcee
Guttilla, formerly Catalina Island Conservancy, personal communication; John Floberg, formerly Catalina
Island Conservancy, personal communication to Adam Backlin).
Bullfrogs likely impact the two-striped garter snake both through direct predation as well as through
competition, by predating on the Pacific treefrog (Pseudacris regilla) (Backlin et al. 2005). The endemic
Catalina shrew is also potential bullfrog prey (Lever 2003), along with more common wildlife on the
island. Pacific treefrogs are the only native frog occurring on the island, and appear to be less sensitive to
bullfrog competition than other native frogs (Kupferberg 1997; Monello et al. 2006) as well as less prone
to decline in general than other amphibians in California (Fisher and Shaffer 1996; Blaustein et al. 2004).
Invertebrates
European honeybee
The European honey bee (Apis mellifera, hereafter honey bee) has been introduced by humans to nearly
every country on earth, and is one of the most widespread and abundant insects (Goulson 2003). It is
possible that the honey bee was introduced to Catalina around the same time that it was introduced to
Santa Cruz Island, in the late 1800‘s (Wenner and Thorp 1994).
Honey bees, unlike most other bees, are social insects which form large colonies with high population
densities (Goulson 2003). They are generalists, which typically visit a hundred plus different species of
plant per region (references in Goulson 2003). As an important agricultural species, honey bees have been
well studied. Their effects on native flora and fauna have only relatively recently been investigated,
however. Reviews of this species‘ impacts have been conducted by Sugden et al. (1996), Butz Huryn
(1997), Goulson (2003), Paini (2004), and Kenis et al. (2009). Their potential and/or demonstrated
impacts include declines in native pollinators, decreased seed set of native plants, disproportionate
facilitation of non-native plant species, competition with native vertebrates and invertebrates for nesting
sites, introduction of parasites and other pathogens, and genetic pollution of native species (Goulson
2003; Kenis et al. 2009), as discussed below. Their impact is likely greatest on native bees (Goulson
2003), which are critical to the maintenance of plant biodiversity, as they are the dominant pollinators in
natural ecosystems and often have co-evolutionary relationships with native plants (Sugden et al. 1996).
However, they could potentially impact a diversity of other pollen- and nectar-feeding organisms, such as:
bats, birds, mammals, and other insects like butterflies and moths, beetles, ants, and flies (Schaffer et al.
1983; Goulson 2003).
There is some disagreement as to whether or not introduced honey bees are detrimental to native
pollinators (e.g. Sugden et al. 1996; Butz Huryn 1997; Goulson 2003; Moritz and Härtel 2005), but this
likely reflects the great difficulty of proving the negative consequences of competition (Sugden et al.
1996; Goulson 2003; Kenis et al. 2009). Some studies have shown declines in native insect and bird
pollinators with honey bee activity (e.g. Sugden and Pyke 1991; Kato et al. 1999; Hansen et al. 2002;
Thomson 2004), or increases in native bees following their removal (e.g. Schaffer et al. 1983; Wenner
and Thorp 1994). Yet other studies have found low overlap in resource use between honey bees and
native bees (Pedro and Camargo 1991) or no evidence of competition with native bees (Roubik 1983).
Sugden et al. (1996), after evaluating the results of multiple studies, conclude that the balance of evidence
points to a negative impact of honey bees on native bees. It has also been suggested that the biological
characteristics described previously, along with the abundance of evidence for displacement, resource
overlap, and heavy resource use, makes it likely that they have a detrimental impact on some species
(Goulson 2003; Kenis et al. 2009). The impact, however, likely depends not only on the species but on
whether or not floral resources are limiting in an area (Goulson 2003), and on levels of rainfall (Wenner
et al. 2000). Any negative effects of honeybees may also be exacerbated by habitat clearing and
fragmentation (Kearns et al. 1998; Paton 2001; but see Winfree et al. 2009); indeed, fragmentation may
itself increase the abundance of honey bees (Aizen and Feinsinger 1994).
There is also evidence that honey bees are inefficient pollinators (Westerkamp 1991), which reduce the
pollination of native plants. In some cases, they are considered floral parasites, where resources are
obtained without efficient pollen transfer in return (Goulson 2003). In Australia, they have been found to
remove more than 80% of the floral resources from a plant, which may displace native fauna and reduce
their densities (Paton 2001; but see Horskins and Turner 1999). Where they dominate in place of native
bee species, they have sometimes been found to decrease the seed set of native plants (e.g. Paton 1993;
Roubik 1996; Gross and Mackay 1998). In other studies, however, where pollen was not limiting to a
native plant, seed set was not reduced (e.g. Gross 2001). Because they extract relatively large volumes of
nectar and pollen and yet move between plants less frequently, they have been found to reduce crosspollination services, yet the genetic consequences of this for the plant would be difficult to demonstrate
(Butz Huryn 1997; Goulson 2003).
Honey bees may facilitate the invasion of other transformer species, as they have been found to
preferentially pollinate the flowers of non-natives (reviewed in Goulson 2003). In some systems, a
transformer plant species may rely solely on honey bees for pollination, such as Scotch broom (Cytisus
scoparius) in Australia (Simpson et al. 2005; although see Parker et al. 2002). This may give those
species an advantage; for example, seed set of yellow star thistle (Centaurea solstitialis) was increased by
honey bees (Barthell et al. 2001). Thorp et al. (1994) report that honey bees on Santa Cruz Island strongly
prefer non-native plant species, based on their relative abundance on those species versus natives. Longterm studies are needed to show whether or not this preference increases non-native plant populations
(Goulson 2003).
Introduced honey bees have likely brought with them pathogens and parasites which can affect native
species (Goulson 2003; Kenis et al. 2009). Honey bees can also compete with native bees, mammals, and
birds for nesting sites, usually in old tree cavities, however the effect of this has not been well explored
(Goulson 2003). Donovan (1980) suggests that nest site requirements of native bees and honey bees are
generally quite different, and that competition for these sites may not be important. There have been
genetic consequences of honey bee introduction as well; for example, hybridization with a different race
of honey bee has imperiled a native honey bee in north-western Europe and the Canary Islands
(references in Kenis et al. 2009).
The full distribution of honey bees on Catalina is unknown, although managed hives were maintained
―scattered around the island‖ until 2002; these were primarily in Middle Canyon and its tributaries (Frank
Starkey, Catalina Island Conservancy, personal communication). Unmanaged hives have recently been
mapped in Coffee Pot and Middle canyons (John Knapp, formerly Catalina Island Conservancy, personal
communication). Although Apis mellifera is the only member of the subfamily Apinae on Catalina, a
review of the genera reported to be negatively affected by honeybees (Paini 2004) reveals at least 25
susceptible species of native bees present on the island: three in the family Colletidae, 14 in the family
Megachilidae, and eight in the family Anthophoridae (Sleeper 1989).
Argentine ant
The argentine ant (Linepithema humile) has been dispersed unintentionally by humans around the world,
and has many features which make it highly invasive (Holway et al. 2002). The first known record of
Argentine ants on Catalina is in 1916 (Hebard and Heller 1999). Between 2002 and 2004, Argentine ants
were found at 11 of 20 sampling sites monitored by the U.S. Geological Survey, occupying each of the
four major habitat types that they surveyed (Figure 23; Backlin et al. 2005). They were the most abundant
Figure 23. Location of Argentine ants (Linepithema humile) at pitfall arrays on Santa
Catalina Island. Note arrays 12 and 13 are represented by a single point since they occur
within close proximity and both arrays are without Argentine ants. Reprinted from Backlin et
al. (2005) with permission.
ant sampled during that time, and were associated with a much reduced diversity of native ants,
particularly the endemic species (Backlin et al. 2005).
Argentine ants are omnivores, which can be scavengers, predators upon small invertebrates, and
herbivores, and often harvest exudates from plants and insects (Holway et al. 2002). They can establish
new populations even if the propagules do not contain a queen, and form enormous ―supercolonies‖
containing multiple queens and composed of interacting, cooperative nests (Holway et al. 2002; Heller et
al. 2008a). In fact, it may be only one ―supercolony‖ that occupies their entire introduced range in
California (Holway et al. 2002). They are very successful competitors due to such features as physical
aggression (including raiding the nests of native ant species), use of chemical defensive compounds, high
population densities, and activity at all times of day and year (references in Holway et al. 2002;
Buczkowski and Bennett 2008). Argentine ants are adapted to life in disturbed and fragmented
environments, particularly moist urban ones, however they readily invade natural areas, as evidenced on
Catalina (Holway et al. 2002; Bolger 2007; Menke et al. 2007). In hot, dry climates such as California,
invasion success is tied to areas with higher moisture (Holway 1998a; Menke et al. 2007).
There has been a proliferation of studies on the impacts of the Argentine ants to native species and
systems in recent decades; they are reviewed in Holway et al. (2002). Their potential impacts involve a
broad diversity of native species and a variety of mechanisms. For invertebrates, they are associated with
the displacement of native ants (Wetterer et al. 2001; Carpintero et al. 2005; Holway and Suarez 2006),
reduced diversity and abundance of a variety of arthropods including flies, bees, wasps, spiders, and
beetles (with consequent impacts to ecosystem processes and services such as pollination and herbivory)
(Lach 2007; Liebherr and Krushelnycky 2007; Nygard et al. 2008; however, see Holway 1998b), and
disruption of parasitic relationships (Martinez-Ferrer et al. 2003).
Such effects have ecosystem-level consequences for food webs, seed dispersal, and pollination. For
example, Human and Gordon (1997) found that the entire trophic structure had been shifted in habitats
where Argentine ants had invaded in northern California, while Sanders et al. (2003) found altered
community structure and assembly following their invasion. Lach (2008) found a 75% reduction in native
bee foraging in the presence of Argentine ants in South Africa, with serious implications for insect
pollination. Some scientists also strongly suspect that Argentine ants alter soil erosion, chemistry, and
turnover (Kenis et al. 2009).
Because native ants are displaced by Argentine ants, this can change seed consumption and dispersal
patterns (and thus plant distribution and abundance patterns), as well as soil characteristics (Holway et al.
2002). For instance, Argentine ants are less effective dispersers of large-seeded plants than native ants,
which can decrease the success of individual rare plants or shape entire plant communities (Bond and
Slingsby 1984; Quilichini and Debussche 2000; Christian 2001; Carney et al. 2003; Gomez and Oliveras
2003; Blancafort and Gomez 2005; Witt and Giliomee 2005). Argentine ants may facilitate some
transformer plants, as was indicated for two introduced vetch species in California (Koptur 1979) and
Polygala myrtifolia in Australia (Rowles and O‘Dowd 2009).
Another interesting indirect effect on butterflies and pollination may occur as a consequence of native ant
displacement. Many of the gossamer-winged butterflies (Lycaenidae, containing the blues, hairstreaks,
and coppers) have a mutualistic relationship with a particular species of ant, which attend and protect
their larvae (Pierce et al. 2002). If those tending ants were to be displaced by Argentine ants which do not
have the same tending role, it would likely have negative consequences for the butterflies and thus the
plants that they pollinate (Holway et al. 2002). This has implications on Catalina, which has ten butterfly
species within the Lycaenidae, one of which (the Avalon hairstreak) is endemic only to that island
(Sleeper 1989), and is considered rare.
Impacts to native small mammals, birds, reptiles and amphibians are less clear, but some potential effects
are documented. Argentine ants are associated with reduced densities of gray shrews (Notiosorex
crawfordi) in southern California (Laakkonen et al. 2001) and nest failure in endangered California
gnatcatchers (Polioptila melanura) (Sockman 1997), although their exact role in these trends is unclear.
They are also confirmed predators on other birds such as the endangered Least Bell‘s Vireos (Vireo bellii
pusillus) (Peterson et al. 2004). Reptiles and amphibians may also be affected by Argentine ants, which
are associated with the absence or low density of coast horned lizards (Phrynosoma coronatum) and
appear to negatively affect western skinks (Fisher unpublished, referenced in Backlin et al. 2005).
Climate change may cause an increase in Argentine ant invasion at Catalina‘s latitude (Roura-Pascual et
al. 2004; Heller et al. 2008b). However, Argentine ant impacts may decrease with time, as abandonment
of invaded areas and consequent increase in native ant species richness has been documented (Heller et al.
2008b). Unsuitable environmental conditions, lack of disturbance, and competition from native ants may
also greatly limit this invader‘s distribution, as was found by Wetterer et al. (2006) on the island of
Madeira.
Transformer plants
Invasive plants disperse and become naturalized 1,000 times faster by humans than they would under
natural conditions (Babbitt 1998), and can greatly reduce the biodiversity and functioning of ecosystems.
Some of the most transformative invasive plant species have been shown to displace wildlife, alter
disturbance regimes and soil properties, degrade wildlife habitat, out-compete native plant species, and
hybridize with native plant species. Transformer plants which have been established on Catalina will be
used as examples to illustrate these effects in the discussion that follows.
Transformer plants can change the fire regime of a native system, altering characteristics such as fire
intensity, frequency, extent, type, and season (Brooks et al. 2004). Such changes can alter food webs and
affect ecosystem properties such as nutrient cycling and soil erosion (Brooks et al. 2004). Changes to the
fire regime can establish a positive-feedback cycle, where fire further promotes the invader, which then
further promotes fire (D‘Antonio and Vitousek 1992). Introduced annual grasses such as bromes (Bromus
spp.) are particularly notorious for such an effect (Zedler et al. 1983; D‘Antonio and Vitousek 1992;
Brooks et al. 2004), which can eliminate or reduce native shrub communities (Zedler et al. 1983;
Haidinger and Keeley 1993; Swiecki and Bernhardt 2002; Keeley 2006a,b; Witter et al. 2007). Other
species which could increase fire frequency and intensity include saltcedar (Zavaleta 2000; Hass 2002),
french broom (Teline monspessulana; Pauchard et al. 2008), eucalyptus (Eucalyptus spp.; Williams 2002;
Brooks et al. 2004), and giant reed (Arundo donax; Scott 1994, cited in Bell 1997; Hass 2002).
In addition to altering fire regimes, some transformer plants can affect both hydrological processes and
erosion rates. For example saltcedar (Tamarix spp.), which consumes water 35% more rapidly than native
vegetation on average, can lower water tables, lower stream flow rates, and dry lakes and streams (Loope
et al. 1988; Zavaleta 2000). It can also trap sediment, causing siltation of river channels and further
causing water flow reduction (Blackburn et al. 1982; Zavaleta 2000). Giant reed, which grows along
riparian corridors, can alter flow regimes by stabilizing stream banks (Zohary and Willis 1992, cited in
Bell 1997). It has a shallow root system, however, which is easily undercut by annual storms; when this
happens, the entire stream bank can collapse (Hass 2002).
Some transformer plants can alter the chemical composition and nutrient levels of the soil where they
dominate. Crystalline iceplant (Mesembryanthemum crystallinum), for example, accumulates salts which
are released when it dies, producing a stressful environment for ill-adapted native plants (Vivrette and
Muller 1977).
Some plant invaders can degrade wildlife habitat, either because they have not evolved with the native
fauna and may be unpalatable or unsuitable structurally, or because they are unfamiliar and thus avoided.
For example giant reed (Arundo donax), which is native to southeast Asia, has displaced willows and
cottonwoods in California‘s riparian zones, thus precluding birds that use those native trees for nesting
(Hass 2002; Bell 1997). Similarly, bird diversity and density is lower in areas dominated by Eurasian salt
cedar (Tamarix sp.) than in native cottonwood-willow vegetation (Brand et al. 2008; van Riper et al.
2008). Reduced herbivorous invertebrate loads are also often observed on introduced plants, such as giant
reed in California, on which aerial invertebrate abundance and richness were approximately half that of
native willow vegetation (Herrera and Dudley 2003).
Many transformer plants can out-compete natives for important resources such as light, water, space,
nutrients, and pollinators. This can alter entire plant assemblages and their structure, and threaten rare
plant species. Highway iceplant (Carpobrotus edulis), for example, displaces native plants by reducing
available water (D‘Antonio and Mahall 1991). Fennel (Foeniculum vulgare) invasion is also correlated
with lowered plant species richness (Beatty 1991).
Transformer plants can also exert a negative impact through hybridization with closely related native
species (Vila et al. 2000). Hybridization can not only push native species toward extinction (Ayres 2003),
but it can also enhance the transformer‘s invasiveness (Ayres 2003; Schierenbeck and Ellstrand 2009).
The invasive mission cactus (Opuntia ficus-indica) hybridizes with the native prickly-pear cactus
(Opuntia littoralis), for example, producing a hybrid species that appears to be more successful due to its
occupation of less fire-prone habitats (Benson 1969, referenced in Vila et al. 2000).
Of the 240 non-native plants on Catalina, a group of local and state botanists identified at least 76
potentially transformer species that are somewhat manageable, based on their observations of the species
in the state as well as the plants‘ ranking by the California Invasive Plant Council (J. Knapp 2004).
Following a systematic and comprehensive island survey for these plants in 2002-2003, and a thorough
literature review, these species were ranked for management priority (J. Knapp 2004) and a systematic
invasive plant program was begun (J. Knapp, this volume).
The list of species targeted includes some of those identified by the California Invasive Plant Council as
California‘s most damaging plant invaders, such as giant reed, artichoke thistle (Cynara cardunculus),
tamarisk, and yellow star thistle. Those four species, for example, have a very limited abundance on the
island, making their eradication both feasible and cost-effective. Other transformer plants which were
already abundant on the island, such as fennel, flax-leafed broom (Genista linifolia), and Harding grass
(Phalaris aquatica), have been controlled in priority areas. For example, flax-leafed broom, which covers
376 hectares (929 acres) on the island and predominates in the vicinity of Avalon, was targeted for
eradication in Cottonwood Canyon, where a disjunct population of the broom threatens some of the
island‘s rarest wildlife species. Selected plant invaders are pictured in Figure 24.
As of 2007, 25 of those species had been nearly eradicated from the island, and another 18 were being
reduced or controlled, many of which invade Quercus pacifica and Quercus tomentella habitats (J.
Knapp, this volume). This type of extensive, strategic transformer plant management appears to be the
most effective way to conduct an invasive plant control program (Randall 1996; Rejmanek and Pitcairn
2002; Bossard and Randall 2007; DiTomaso et al. 2007), and can minimizing the threat that these species
pose to biodiversity and ecosystem functioning.
One group of transformer species that is not discussed in depth in Knapp (2004) is introduced annual
grasses, because they are so widespread and extremely difficult to manage (Prober et al. 2005; Bossard
and Randall 2007; Di Tomaso et al. 2007; Stromberg et al. 2007; Cox and Allen 2008). They pose an
important threat to Catalina‘s oak ecosystem, however, and their impacts are discussed below.
The effect of non-native annual grasses on oak regeneration has been reviewed previously in this
document. These grasses have also been shown to suppress the growth of native shrubs, forbs, and
perennial grasses (Young and Evans 1973; Da Silva and Bartolome 1984; Gordon et al. 1989; Danielsen
and Halvorson 1991; Eliason and Allen 1997; Dyer and Rice 1999; Hamilton et al. 1999; Brooks 2000;
Brown and Rice 2000; Carlsen et al. 2000; Kolb et al. 2002; Seabloom et al. 2003; Yelenik 2008), alter
the phenology of water resources and ecosystem production (Gordon et al. 1989; D‘Antonio et al. 2007;
Morghan et al. 2007), shift nutrient cycling regimes (Yelenik 2008), change hydrologic and
geomorphologic processes (Keeley 2002, 2006b), reduce forage quality for small mammals such as
ground squirrels (Van Horne 2007), and facilitate disease and insect pests (Malmstrom et al 2005).
Furthermore, non-native annual grasses can increase the frequency of fire, which further promotes the
grasses (D‘Antonio and Vitousek 1992). Introduced annual grasses have been associated with short fire
return intervals in some southern California shrublands (Keeley 2006a; Witter et al. 2007).
Figure 24. Artichoke thistle, upper left, is one of 25 ecologically damaging plants which were
targeted for eradication. Other species such as: (clockwise from upper right) fennel, Harding
grass, and flax-leaf broom have been controlled in priority areas. Photos by Peter Schuyler, lower
left, and the author.
Roads
There are 342.5 kilometers (213 miles) of roads through Catalina‘s wildlands (not counting those within
the Avalon city limits), traversing nearly every part of the island (Figures 3, 26). In addition, the
permitting process is underway for construction of one more road, to one of the largest and most remote
beaches on the island, Ben Weston. Roads can impact ecosystem function and biodiversity in numerous
ways: by increasing runoff, erosion, and stream sedimentation, reducing and altering habitat, increasing
wildlife mortality, promoting invasion of non-native plants and animals, increasing pollution, promoting
fire ignition, increasing human access to remote areas, and forming a barrier to dispersal for smaller
species such as invertebrates, amphibians and reptiles. The adverse ecological effects of roads have been
reviewed by Spellerberg (2002), Trombulak and Frissell (2000), and Forman et al. (2003), and will be
summarized below.
Some of the most severe impacts of roads are those to aquatic ecology and water quality (Luce 2002).
Roads disrupt, alter, and concentrate water flow and circulation, often causing excessive siltation and
drying of streams and wetlands, or even complete wetland elimination (Jones et al. 2000; Luce 2002;
Forman et al. 2003). The impervious surface of roads reduces water infiltration and causes increased
runoff; that runoff is then redistributed along roads lateral to the slope, compromising slope stability
(Luce 2002). The exposed soil of unpaved roadways contributes substantially to both wind and water
erosion (Spellerberg 2002; Forman et al. 2003; Figure 26). This can cause the formation of large gullies
(Forman et al. 2003) and increase landslide risk (Swanson and Dyrness 1975). Roads which cut into
slopes can have particularly severe effects on runoff and erosion, sometimes cutting into subsurface water
flow (Luce 2002).
Figure 26. Roads have contributed to extensive erosion on Catalina, as well as thick dust deposition
on heavily traveled routes. Photos by the author.
The increased sedimentation from road-induced erosion, along with altered water flow, flow duration,
materials movement, and stream channel structure can substantially affect the food web in affected
streams, ponds, and oceans (Luce 2002; Forman et al. 2003). In addition, windblown dust from these
roadways (Figure 26) can reduce the photosynthesis and decrease the water-use efficiency of roadside
plants, lichens, and mosses, leading to lowered production or even complete elimination of some taxa
(Walker and Everett 1987; Thompson et al. 1984; Sharifi et al. 1997). Elevated levels of dust and
pollutants (the latter of which is discussed below) also have implications for human health and climate
change (Forman et al. 2003).
Vehicles also add a wide variety of pollutants to the ecosystem, including salts, mineral nutrients such as
nitrogen or phosphorous, heavy metals such as zinc and cadmium, and organic compounds such as oil,
hydraulic fluid, gasoline, and other contaminants (references in Angold 1997; Trombulak and Frissel
2000; Forman et al 2003). Such pollutants can be incorporated into the groundwater or transported long
distances via runoff, and find their way into streams even from remote hillsides; they also build up in the
soil and plants, and have cascading effects through the ecosystem (Forman et al. 2003). Vehicle exhaust
also adds carbon dioxide, sulphur dioxide, oxides of nitrogen, organic gases such as ethylene, and heavy
metals such as lead into the atmosphere, all with serious environmental effects (references in Angold
1997). Increased concentrations of elements such as calcium, potassium, sodium, and phosphorous in
road dust can stimulate the growth of certain plant species at the detriment of others, thus significantly
changing the composition of roadside plant communities (Forman et al. 2003). In addition, elevated
nitrogen levels along roadways can increase herbivory in plants or reduce microbial insect controls,
leading to insect outbreaks (Port and Thompson 1980; New 1984).
Road construction and maintenance activities directly destroy habitat for plants, animals, and other
organisms. Roads also alter the habitat structure and connectivity, visual and audial disturbance levels,
soil properties, and microclimate of the remaining habitat (Trombulak and Frissell 2000; Forman et al.
2003). This often reduces the productivity of wildlife, plants, and other primary producers such as lichens
and alters their population dynamics and community composition (e.g., Angold 1997; Auerbach et al.
1997; Forman et al. 2003; Houlahan and Findlay 2003).
Roads can be detrimental to wildlife in a variety of other ways. The most direct is vehicle-caused
mortality (Figure 27). Animals of all sizes are vulnerable to being road-kill; those which are attracted to
roads as basking sites, such as snakes, are particularly at-risk (Forman et al. 2003). Areas with substantial
habitat cover adjacent to the road, such as on Catalina, are especially deadly (Forman et al. 2003). For
some species, such as bald eagles and panthers in Florida, the death rate from vehicles exceeds that of
natural causes (Forman et al. 2003). Indeed, vehicles were the predominant source of known Island fox
mortalities between 2004 and 2006, compromising the population‘s recovery from a devastating outbreak
of canine distemper virus (Carlos de la Rosa, presentation to Conservancy Board of Directors 2006).
Wildlife mortality generally increases with increasing traffic and speed, but is not limited to high-traffic
areas; for instance, at least one of 15 island fox fatalities reported between 2004 and 2006 was on a
relatively low-use road south of Avalon (de la Rosa et al. 2006). Loss of an individual to road-kill can
cause further impacts to the population when a mate or parent is killed. Wildlife groups most at risk are,
in decreasing order, birds, mammals, reptiles and amphibians (Forman et al. 2003). Even at the lower end
of that spectrum, the loss can be substantial where traffic density is high; for instance, roads accounted for
about 10% of adult frog mortality in one Denmark study (Hels and Buchwald 2001).
Figure 27. Roads are a significant cause of wildlife mortality: at left, a gopher snake; at right, an
island fox. Photos by: Lauren Danner, and an unknown photographer, respectively.
Animals are often more vulnerable to predation along open road corridors (Anderson and Burgin 2008).
Behavioral changes can also be detrimental to wildlife, including responses such as avoidance or elevated
stress hormones; such effects can occur even with relatively few vehicles per day (Trombulak and Frissell
2000; Forman et al. 2003). The cumulative influence of the mechanisms discussed above can result in
decreased population sizes and densities for many taxa. For instance, 60% of grassland birds in Denmark
exhibited decreased population sizes adjacent to roads, even those that were relatively little-travelled
(Reijnen et al. 1995); population loss as high as 60% was observed in another study (van der Zande et al.
1980). A diverse range of taxa have exhibited an avoidance zone within which their population densities
are lowered, including large and small mammals, birds, reptiles, and amphibians (Forman et al. 2003).
The ecological impact of this avoidance could exceed that of either habitat destruction or roadside
mortalities (Forman et al. 2003).
Organisms which regularly move through different habitats for activities such as feeding and breeding,
including amphibians, insects, and other invertebrates, are adversely affected by the reduced landscape
connectivity resulting from roads (Forman et al. 2003). Dispersal is essential for maintenance of
metapopulations and genetic diversity; reduced connectivity can result in higher mortality, lower
reproduction, and lowered population sizes and viability (Forman et al. 2003). This can be true for small
mammals even when the road is unpaved and partially overgrown (Forman et al. 2003).
The fragmentation caused by roads also creates extensive stretches of edge habitat which is unfavorable
to many species requiring more intact vegetation communities (Forman et al. 2003). Invasive wildlife
species are typically favored by this edge habitat; this was noted for rats, mice, and starlings earlier in this
document, but has also been found for invasive amphibians such as cane toads (e.g., Bias and Morrison
2006; Seabrook and Dettmann 1996; Antos and White 2003; Ganzhorn 2003). Roads promote alien plant
invasion and dispersal (Gelbard and Belnap 2003; Merriam et al. 2006) by removing native shrubs,
providing a source of constant disturbance, and forming long, interlinked corridors along which vectors
such as animals, people, and vehicles travel. Roads are also the place of origination for most humanignited fires (Keeley 2002).
All of the above effects of roads are exacerbated by increased human access both to the roadsides
themselves and to the remote destination, which lead to trampling and trailing, fire ignition, invasive plant
dispersal, resource collection, refuse, and pollutants.
An animal species‘ behavior, population density, and dispersal ability will determine the degree of road
impact, as will the traffic intensity, road width and topographic placement, and adjacent habitat
characteristics (Forman et al. 2003). Increased moisture along roadsides may benefit some species such as
rodents and invertebrates, while scavengers and predators may also benefit from increased hunting
efficiency or roadkill; this, of course, is balanced by direct mortality from cars and other negative effects
(Antos and White 2004; Forman et al. 2003). Although the presence of roads may benefit some species,
however, those appear to be mostly common species such as ravens, exotic birds, and invasive non-native
plants, and it seems the negative effects far outweigh any positive effects.
Increased erosion and gully formation, stream sedimentation, dust deposition, and wildlife mortality are
evident along roads throughout the island (Figure 27). Donaldson and Bennett (2004) suggest that a
road‘s impact may extend up to 300 meters on either side; for Catalina Island, this equates to 12,603
hectares (31,142 acres), or 65% of the island‘s habitat (Figure 28). Other studies place the impact for rural
roads at 500 to 600 meters (van der Zande et al. 1980), which would, at 500 meters, mean an impacted
area of 16,000 hectares (39,536 acres), or 82% of the island (Figure 28).
Hydrologic alteration
The island‘s hydrologic regime has been altered in at least four ways: damming, water withdrawal, roads
through streams, and overgrazing. These alterations have been found to have a diversity of effects, which
can impact oaks and their associated species. Altered hydrologic regimes and sediment loads are two of
the leading threats to freshwater fauna nationwide (Richter et al. 1997). Each of the four types of
hydrologic alteration will be reviewed below, starting with the situation on Catalina followed by the effect
this has been found to have elsewhere.
Roads
Primary dirt
Primary paved
Secondary dirt
Tertiary dirt
Abandoned
Road impact 300m
Road impact 500m
Shoreline
N
7
0
7
14 Kilometers
Figure 28. Roads and their potential impact areas, Catalina Island
The majority of Catalina‘s watercourses have dams; some, such as Bulrush and Cottonwood Canyons,
have several (Figure 3). Such impoundments alter flood flows as well as base flows, homogenizing river
flow regimes (Poff et al. 2007). They also trap sediment behind the dam (thus decreasing sediment loads
downstream), and alter water tables, water temperatures, and the incision of stream channels (Downs et
al. 2002; Poff et al. 2007). The ecological consequences of these effects include reduced cover,
connectivity, and diversity of riparian vegetation (Downs et al. 2002) and facilitation of non-native
species such as bullfrogs, tamarisk, and predatory fish (Everitt 1980; Sartorius and Rosen 2000; Bunn and
Arthington 2002). Indeed, entire food webs are altered, as natural river fluctuations are important in
regulating the life cycles of river biota (Power et al. 1996; Bunn and Arthington 2002).
Catalina‘s approximately 4,000 residents and over one million visitors a year use water that is drawn
predominantly from wells on the island, particularly those in Middle Canyon. It could be expected that
this would have an effect on the island‘s groundwater, although how much is not known. Oaks, which are
generally deeply rooted (Lewis and Burghy 1964; Griffin 1973; Ogden 1975; Thomas 1980), may be
affected by this human alteration. Lowered water tables and shallower aboveground water sources are
also particularly detrimental to species dependent on aquatic systems, such as amphibians (Sartorius and
Rosen 2000; Blaustein et al. 2004).
Roads cross or traverse the island‘s streams frequently, altering the size, distribution, and amount of
sediments — features which are critical to establishing plants and animals. This has been discussed
above, under the heading of road impacts.
Cattle, bison, goats, sheep, deer, and antelope have inhabited Catalina for varying lengths of time since
the mid 1800‘s, as discussed previously in this document. Overgrazing by livestock and other introduced
animals can alter the hydrologic regime in multiple ways, resulting in: increased water temperatures and
evaporation, lowered water tables, altered stream channel morphology, and increased nutrients and
erosion, as well as reduced or altered riparian vegetation density, diversity, and composition (Platts 1979,
1981; Kauffman and Krueger 1984; Armour et al. 1991; Fleischner 1994; Belsky et al. 2002; Jones 2002;
Kauffman 2002).
Fire
Fire is a natural source of disturbance in most Mediterranean ecosystems such as Catalina‘s, and the
presence of fire-dependent shrubs (obligate seeders) on the island as well as fire-following annual herbs
such as fire poppy (Papaver californicum) and whispering bells (Emmenanthe penduliflora) indicate that
fire has been a predictable presence on the island historically (Keeley 2006a). Within the normal
parameters of a region‘s fire regime (including fire type, season, size, frequency, and intensity), fire can
be a regenerative force. However, outside of these parameters, and depending on the biotic environment
where the burn occurs, it can be a significant threat to biodiversity. As described below, the positive or
negative effects of fire on Catalina will be heavily determined by the presence and density of introduced
ungulates, the nature of the seed bank present (i.e. native or transformer species), and the time since the
last burn.
Introduced mule deer have in some places completely altered the species composition of the post-fire
landscape on the island. Endemic obligate seeder species such as the Channel Islands tree poppy
(Dendromecon harfordii) and Island rush-rose (Helianthemum greenei) have germinated in abundance
following recent fires, but seedlings of these species are preferred by the introduced mule deer, which
have decimated them following recent island fires (Knapp 2005b; Sarah Ratay, Catalina Island
Conservancy, personal communication). Mature plants of these species are also at risk from the deer
(Manuwal and Sweitzer, this volume), as are resprouts of dominants such as oaks (Knapp 2008) and
toyon (Ramirez 2008). Those impacts have been discussed under the header of mule deer earlier in this
document.
Germination of transformer plant species such as brooms (Cytisus, Genista spp.), fennel, and yellow star
thistle may be promoted by fire at moderate temperatures (Terrega et al 1992; Bossard 2000; Alexander
and D‘Antonio 2003; Klinger and Brenton in review; DiTomaso and Gerlach 2000). Alexander and
D‘Antonio (2003) showed that a single fire could significantly reduce the French broom seedbank in part
due to stimulated germination. With proper contingency planning and a prompt post-fire response, this
feature may be successfully used to control plant invasions; however, without those actions
transformation of the landscape to one dominated by invasive species may be accelerated. In addition to
the impacts of the fire itself, fuel breaks may serve as establishment sites and dispersal corridors for
invasive, non-native species (Keeley 2006a).
Shrubs and trees of both resprouting and seeding species can be eliminated by too-frequent fire.
Resprouter species may be eliminated by fires occurring more frequently than ten years (Keeley 2006b).
Obligate seeders require up to 40 years between fires to build up a sufficient seedbank to replace
themselves after fire (Keeley 2006a), thus fires occurring more frequently than this can eliminate these
seeder species from a shrubland community. Frequent fire can also cause a type conversion to a non-
native annual grassland (Zedler et al. 1983; Haidinger and Keeley 1993; Swiecki and Bernhardt 2002;
Keeley 2006b; Witter et al. 2007). Annual grasses increase the rate of ignition and spread of fire, which in
turn favors the annual grassland in a feedback loop (D‘Antonio and Vitousek 1992; Mack and D‘Antonio
1998; Brooks et al. 2004). The resulting grassland fires are less intense than shrubland fires, and allow for
a high survival of annual grass seeds (Keeley et al. 2005). Such type conversion of native shrubland to
grassland can affect hydrological and morphological processes as well (Keeley 2002). In addition to
biotic impacts, increased fire frequency can promote erosion: for instance, in the San Gabriel Mountains
of southern California, recurrent human-ignited fires have increased sediment production by an average of
≥60%, and some debris basins have exhibited up to 400% increases in deposition (Lavé and Burbank
2004).
In order to place Catalina‘s current fire frequency into context, it is helpful to examine the fire history of
the previous centuries. A review is provided below.
Fire history
Fire frequencies on Catalina have likely fluctuated substantially in recent centuries. Keeley (2006a)
hypothesizes that during the Pleistocene fire was infrequent in coastal southern California (every 50-100
years); the rate may have been even lower on the California Islands, due to fire‘s inability to spread across
an ocean barrier and a naturally lower incidence of lightning-ignited wildfires in coastal regions (Carroll
et al. 1993; Keeley 2006a). Lightning-caused fires are still occasionally recorded; there have been six
recorded lightning-caused fires on the Island since 1967 (D. Knapp, unpublished data).
Purposeful burning by Native Americans apparently raised the natural fire frequency in pre-historic
coastal southern California (Keeley 2006a). Fire was used regularly by Native Americans as a
management tool to encourage seed plants, bulbs, and green shoots critical to their diet (Timbrook et al.
1993, Anderson 2006), and it is likely that it was used regularly on the California Islands as well (Keeley
2006a). This would have ceased with the elimination of Native American populations on the island,
which was complete by the 1820‘s (Schoenherr et al. 1999).
Settlers in early California, particularly nomadic sheepherders and miners, also set fires throughout the
state to improve forage and open the habitat up to exploration (Bauer 1930; Clar 1959; Keeley 2006a).
Given the historical presence of both of these groups on Catalina, it is possible that relatively high fire
frequencies were maintained through the early 1900‘s.
Despite widespread policies of fire suppression beginning in the late 1920‘s and early 1930‘s (Clar 1969a,
1969b; Rogers 1982), large fires continued to burn in coastal southern California throughout the 20 th
century (Moritz 1997). On Catalina Island, however, historic overgrazing combined with fire suppression
appear to have limited the number of major fires over the past century (Minnich 1982; Landis 2000). This
has favored plant species that can germinate between fire events, such as lemonadeberry (Rhus
integrifolia), laurel sumac (Malosma laurina), oaks, and toyon (Heteromeles arbutifolia). These species
respond to fire by resprouting from resources stored at their base (and are thus termed resprouter species).
Species which require fire to germinate (obligate seeder species), such as the endemic Catalina manzanita
(Arctostaphylos catalinae) and felt-leaf ceanothus (Ceanothus arboreus), are much more localized on the
island, and existing stands appear to have established quite some time ago (Landis 1997; D. Knapp,
unpublished data).
Fire suppression activities throughout the 20th century have generally offset an increase in human-caused
ignitions in California (McClaran and Bartolome 1989; Keeley et al. 1999; Keeley 2006a). However, in
recent years both the fire frequency and the area burned per decade have been increasing around urban
areas, paralleling growth in human populations (Keeley and Fotheringham 2001; Witter et al. 2007). The
fire frequency on Catalina Island appears to mirror this recent trend. There have been 229 known
ignitions in Catalina‘s natural areas within the last century, and 97% of them were human-caused (D.
Knapp, unpublished data). Prior to the year 2000, however, none of these fires had burned over 500 acres.
This changed with the Empire Burn in July 2006, which covered approximately 1,100 acres, and the
Island Fire in May 2007, which burned 4,700 acres.
In addition, the length of the fire season has increased from historical times in California due to human
ignitions, from primarily summer and fall fires (fall accounting for the majority of the area burned) to all
year long (Keeley 2006a); indeed, two of the last 5 major fires on Catalina have been outside of the
typical fire season (January and May). This fire season extension facilitates an increase in fire frequency
and may have important effects on the post-burn vegetation response.
CONCLUSIONS
Quercus pacifica has been declining on Catalina Island for at least sixty years. Although senescence
appears to be a major cause, and stressors such as browsing and trampling by introduced ungulates may
be hastening their demise. In addition, lack of recruitment to replace this loss can be in large part
attributed to the effects of non-native annual grasses and introduced animals. Populations of Quercus
tomentella are small and isolated are likely limited by low pollination and seed dispersal, and browsing by
introduced ungulates.
A multitude of invasive, transformer species with dramatic ecological impacts threaten the island‘s native
and endemic species. These threats are joined by: fragmentation due to roads; altered hydrology from
dams, groundwater withdrawal, and stream channel modifications; increased fire frequency; and other
human impacts such as development. These threats, which may compound and facilitate one another,
have left the ecosystem highly disturbed and modified.
The effects of fragmentation, wildfire, transformer species, and hydrologic alterations will all be
exacerbated by climate change (California Natural Resources Agency 2009). Under current climatic
projections, stream flow will be reduced, water temperatures will rise, and groundwater recharge will be
reduced, affecting both wildlife and humans (California Natural Resources Agency 2009). At the same
time, existing stressors such as dams, roads, and invasive species will reduce the capacity of plants and
animals to adapt to the effects of climate change (California Natural Resources Agency 2009). The
ecological tolerance of many endemic species will likely be exceeded, particularly aquatic species
(California Natural Resources Agency 2009). Relict species such as island oak, ironwoods, and Trask‘s
mahogany will be particularly at risk. Endemic oak ranges in California may shrink as much as 59%, and
will likely shift northward (Kueppers et al. 2005).
Although important steps have already been taken towards protecting Catalina‘s ecosystem, such as feral
goat and pig removal and transformer plant control, much more action will be necessary if structure and
function are to be restored. Restoration priorities should take into account the degree of the threat, the
chances of success, and the benefits to be derived from such action. The groundwork for setting those
priorities has been laid here, along with a ranking of watershed resources based on the rarity of their
vegetation, rare plant, and wildlife resources (Knapp and Knapp 2005 and unpublished improvements).
Recommendations for restoration of the ecosystem have been submitted to the Conservancy, but omitted
here at their request.
Biotic interactions should also be taken into account when devising a strategy, in order to avoid
unintended consequences. Removal of transformer species is complicated by their interactions with each
other and with pre-existing native species (Towns et al. 1997; Zavaleta et al. 2001; Mack and Lonsdale
2002; Courchamp et al. 2003). When multiple invaders are being targeted, the order of control can make
the difference between success or failure (Fan et al. 2005; Klinger 2007). Experimental, localized removal
of transformer species, dams, or roads would be a step toward achieving restoration goals while
contributing to the understanding of the ecosystem and illuminating any unintended consequences which
may result from such actions.
Restoring the ecosystem will not be straightforward, quick, or easy. Restoration requires a long-term
commitment (SER International 2004; McCreary 2004; DiTomaso et al. 2007), and this is only the very
beginning of a long process if function and self-sustainability are to be restored to Catalina‘s oak
ecosystem. The first step, researching the site and determining the problems, has been achieved; now
goals and objectives must be set, a plan must be implemented, and results assessed (Bossard and Randall
2007).
Although conservation efforts on Catalina are constrained by many factors, substantial restoration
progress is achievable. Some of the Island‘s constraints are also key assets. The high level of visitation to
Catalina makes it an ideal educational and outreach center for the southern Channel Islands. The
Conservancy‘s Nature Center and Botanical Garden are perfect venues to educate about the uniqueness of
the islands, the threats that face them, and the benefits of restoration.
The ecosystem framework and collaborative approach of this multi-year project were beneficial in a
number of ways. Approaching research and restoration from an ecosystem perspective has allowed the
Conservancy to seek answers to a variety of interdisciplinary questions pertinent to the health of the
island, while still maintaining focus on the island‘s oaks, which are a dominant and beloved feature. It
also enabled the use of funds for both gathering of baseline information and acquisition of multi-use tools,
which will continue to be useful for a variety of research and management projects on the island. The
collaborative approach taken here both formed and strengthened a multitude of partnerships which led to
more projects than the Catalina Island Conservancy could have supported alone, with the benefit of
extensive additional expertise. These projects were complementary, and thus tackled our questions more
thoroughly than otherwise possible. Through the efforts outlined here, science and management have
been truly integrated and much has been learned. Yet much remains to be done, and learned.
ACKNOWLEDGMENTS
Generous support by the Seaver Institute made this work possible through funding of workshops, baseline
data collection, and many of the research projects discussed here. The ecosystem approach was the
brainchild of Peter Schuyler, who wrote the grant that got this process going. A sincere thank you goes
out to all oak workshop participants: Mary Ashley, Todd Caldwell, Sarah Chaney, Rosi Dagit, Stephen
Davis, Carlos de la Rosa, Meriam Djelidi, Janet Franklin, Russell Greenberg, Darcee Guttilla, Michael
Herrera, Rodney Honeycutt, Elizabeth Kellogg, Laura Kindsvater, Julie King, Jo Kitz, John Knapp, Ed
Larson, Thad Manuwal, Eric McDonald, Kathryn McEachern, Tony Michaels, Aaron Morehouse, Ann
Muscat, Kevin Nixon, Bruce Pavlik, Sarah Ratay, Darren Sandquist, Tom Scott, Victoria Seaver-Dean,
Scott Sillett, Lisa Stratton, Kathryn Suding, Rick Sweitzer, Janet Takara, Claudia Tyler, Marti Witter,
Mariana Yazbek, and Jongmin Yoon. John Knapp and Darcee Guttilla were invaluable in providing
information, brainstorming sessions, and editorial comments which greatly improved the paper. Carla
D‘Antonio also gave excellent editorial comments which improved the paper. Vyki Englert and Sarah
Ratay investigated the introduction of various species to the island. In addition, members of the Integrated
Hardwood Range Management Program provided useful comments on the project: Reginald Barrett,
Sheila Barry, James Bartolome, Caroline Bledsoe, Ken Churches, Larry Costello, Morgan Doran, Sabrina
Drill, Barry Garrison, Robert Keiffer, Royce Larsen, Karl McArthur, Doug McCreary, Adina
Merenlender, Tom Scott, Bill Stewart, Bala Thiagarajan, Bill Tietje, and Claudia Tyler. Members of the
D‘Antonio lab group at UCSB also provided helpful feedback: Carla D‘Antonio, Gail Drus, Alice Levine,
Nicole Molinari, Karen Stahlheber, and Vivianne Vincent.
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Synthesis of Ecosystem Resources and Threats

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