Jul 14, 2000 - interactions for the intertidal benthic community of Yaquina Bay suggested reduced ...... 0,. 40, and 80 cm high-tide level. Core samples from each transect and tide are .... similarly between estuaries (Table 2.1; r = 0.98, p < 0.001, n = 6), and it ...... communities Quisset Harbor, Massachusetts. p. 191-227.
AN ABSTRACT OF THE THESIS OF
Gonzalo C. Castillo for the degree of Doctor of Philosophy in Fisheries Science presented on July 14, 2000 Title: Benthic Biological Invasions in Two Temperate Estuaries and Their Effects on Trophic Relations of Native Fish and Community Stability.
Redacted for Privacy Abstract approved: Hiram W. Li
The extent of biological invasions, their role on the feeding of native fishes and their impact on community stability were investigated in Alsea Bay and Yaquina Bay, two estuaries on the central Oregon coast, USA. Most nonindigenous species (NIS) introduced in these intermediately invaded estuaries are considered byproducts of culturing introduced Atlantic and Pacific oysters. Secondary potential vectors of NIS in Yaguina Bay are external fouling of ship hulls and ballast water. Native benthic invertebrates and native fishes dominate in density, catch per unit effort (CPUE) and richness in both estuaries.
Three of the 11 benthic NIS of invertebrates in Yaquina Bay and one of the eight NIS in Alsea Bay are among the 10 most dominant benthic invertebrate species. The NIS of invertebrates are concentrated in habitats with above average water temperature, salinity, and macrophyte density at high-tide. The CPUE of fishes and decapod crustaceans are associated wi.th above average
water temperature, salinity and macrophyte density but are not consistently correlated with invertebrate density in sediments. Biological invasions have caused significant prey shifts in intertidal food webs of Yaquina Bay. Diets of two species of native juvenile flatfishes
stellatus) included mainly polychaetes, crustaceans and bivalves and each of these taxa are represented in the diet by native species and NIS in each estuary. Both flatfish species are
generalist predators and had no consistently higher selection
for either native species or MIS. Prey selection experiments indicated that two native and two introduced amphipod prey (Corophium spp.) are acceptable prey for juvenile English sole.
Thus, predator-prey coevolution plays no significant role on
prey selection. Interspecific prey selection may depend on prey exposure, water visibility, substratum type, and species diversity of available prey. Modeling of functional-group interactions for the intertidal benthic community of Yaquina Bay suggested reduced community response to invasions or removal of fish predators as indicated by the community tendency to zero overall-feedback. However, the increased risk of stability decline of invaded community models implies that further humanmediated biological invasions should be avoided.
Benthic Biological Invasions in Two Temperate Estuaries and Their Effects on Trophic Relations of Native Fish and Community Stability
Gonzalo C. Castillo
A THESIS submitted to
Oregon State University
in partial fulfillment of the requirements for the degree of Doctor of Philosophy
Presented July 14, 2000 Commencement June 2001
Doctor of Philosophy thesis of Gonzalo C. Castillo presented on July 14, 2000
Redacted for Privacy Major Professor, representing Fisheries Science
Redacted for Privacy Chair of Department ofisheries and Wildlife
Redacted for Privacy Dean of G. .'e School
I understand that my thesis will become part of the permanent collection of Oregon State University libraries. My signature below authorizes release of my thesis to any reader upon request.
Redacted for Privacy hnzalo C. Castillo, Author
I would like to thank my advisor Dr. Hiram Li and Dr. John Chapman and Dr. Philippe Rossignol for their active involvement and support throughout my graduate program. I also thank other members of my committee: Dr. Susan Sogard; Dr. William Pearcy; Dr. Peter Bayley; Dr. Steven Rumrill; Dr. Eon
011a; Dr. Loren
Koller and Dr. Larry Curtis for their participation and comments on manuscripts. The taxonomic assistance of Dr. John Chapman;
Dr. James Carlton; Dr. Leslie Harris; Dr. Faith Cole; Dr. Eugene Kozloff; Dr. Les Watling; Dr. David Behrens and Dr. Jeffery Cordell is greatly appreciated. The active participation and substantial dedication of Todd Miller throughout this project was critical for conducting all the field sampling and the initial laboratory analyses. The help of Dr. Hiram Li; Dr.
Robert Olson; John Sewall and Scott Pozarycki was essential for the completion of experiments. The field assistance of Dr. John Chapman; Dr. Hiram Li; James Golden; John Johnson and Amy Chapman throughout the field season is greatly appreciated.
Gabriela Montaño and Jeffrey Dambacher provided generous assistance on software operation. I thank Jeremy Bonnichsen;
Kevin Crow; William Krueger; Terrin Ricehill; Peny Noland; Patty Gipson; James Archuleta; Orbi Danzuka; Liu Xin; Marcus Beck; Wilfrido Contreras and others for their field and/or laboratory assistance.
I thank Patrick Clinton and Dr. Walt Nelson for
providing aerial images of Alsea Bay and Yaquina Bay. The Native Americans in Marine Sciences Program at Oregon State University and the Oregon Sea Grant College Program contributed with vital help and funding to accomplish this research.
CONTRIBUTION OF AUTHORS Hiram Li and John Chapman were involved in the design of field research (chapters 2 and 3) and laboratory experiments (chapter 4) .
They participated in the initial field surveys;
species identifications and manuscript reviews. Todd Miller collaborated in all field surveys and helped to collect samples for laboratory experiments. Hiram Li and Philippe Rossignol were involved in model construction and analyses (chapter 5)
TABLE OF CONTENTS Page
Chapter 1: General Introduction
Significance of Biological Invasions
Mechanisms of Biological Invasions
Community Susceptibility to Biological Invasions
Extent of Estuarine and Marine Invasions
Potential Impacts of Nonindigenous Species
Invasions in U.S. West Coast Estuaries
Focus of the Present Thesis
Chapter 2: Distribution and Habitat Use by Noncoevolved Assemblages of Macroinvertebrates and Fishes in Two Temperate Estuaries
Chapter 3: Trophic Contribution and Selection of Native and Nonindigenous Prey by Native Fishes in Estuarine Rearing-Habitats
TABLE OF CONTENTS (Continued) Page
Chapter 4: Predation on Native and Nonindigenous Aiuphipod Crustaceans by a Native Estuarine-Dependent Fish
Chapter 5: Absence of Overall Feedback in a Benthic Estuarine Community: A System Potentially Buffered from Impacts of Biological Invasions
Chapter 6: Conclusions
Recommendations for Future Research
Appendix A: Complement of Chapter 2
Appendix B: Complement of Chapter 3
Appendix C: Complement of Chapter 4
Appendix D: Complement of Chapter 5
LIST OF FIGURES Page
Alsea Bay and Yaquina Bay estuaries
Summer mean density of benthic invertebrates in sediment core samples from Alsea Bay and Yaquina Bay(bars) and percent of nonindigenous to native species density (circle)
Summer mean CPUS of fishes samples from the Alsea Bay (bars) and percent of CPUE relative to native species
and decapods in seine and Yaquina Bay estuaries of nonindigenous species (circle)
Clusters by taxa and sites based on invertebrate densities in sediment samples
Clusters by taxa and sites based on CPUS of fishes and decapods in seine samples
Mean summer densities of assemblages of native and nonindigenous invertebrates in sediment samples under various temperature-salinity combinations
Mean percent densities for summer assemblages of native and nonindigenous invertebrates under various temperature-salinity combinations
Mean percent richness for summer assemblages of native and nonindigenous invertebrates under various temperature-salinity combinations
Ordination of 35 benthic invertebrates from sediment samples and 12 high-tide intertidal sites along environmental gradients
2.10 Ordination of 12 fishes, three decapods (code in parenthesis) and 12 high-tide intertidal sites along environmental gradients 3.1
Fish and invertebrate collection sites in Alsea Bay and Yaquina Bay
Mean number and volume of prey species in the diets of English sole and starry flounder
Percent volume of major prey by origin in English sole and starry flounder diets
Percentages of prey frequency of occurrence, number and volume of main dietary items of English sole by estuary section
LIST OF FIGURES (Continued) Figure
Percentages of prey frequency of occurrence, number and volume of main dietary items of starry flounder by estuary section
Total prey volume of native and nonindigenous species and all taxa combined as a function of fish weight in intertidal areas
Mean number (A-D) and volume (E-H) of prey in the diet of juvenile English sole and starry flounder collected at low- and high-tide
Mean number of invertebrates in the flatfish diet (A-D), total densities for all invertebrates in the benthos (E-F) and CPUE of flatfish (G-H)
Johnson's selection index (line) and ranks of prey usage and availability (bars) for flatfish in Alsea Bay
3.10 Johnson's selection index (line) and ranks of prey usage and availability (bars) for flatfish in Yaquina Bay
Amphipod body length (from telson to eye, Y) with length of 4th articLe 2nd antenna (X) by species and sex
Mean activity (distance traveled in 5 s) by 10 males and 10 females of each Corophium species held at 14°C and at 24°C
Mean number of Corophium consumed by Pieuronectes vetulus, in single-species experiments
Strauss' selection index by prey size (4th article 2nd antenna) and Corophium species consumed by Pleuronectes vetuius
Percent of eaten and urieaten Corophium by size (4th article 2nd antenna) in 10 tanks with sand substratum
Mean number of Corophium consumed by Pieuronectes vetulus in mixed-species experiments
Percent of uneaten Corophium by size (4th article 2nd antenna) in sand and mud substrata
LIST OF FIGURES (Continued) Figure 5.1
Ecological interactions between guilds 1, and 3 in numbered circles and attendant community matrices
Basic guild structure of activity models for the benthic community of the Yaquina Bay estuary
Basic guild structure for trophic models of the benthic community in the Yaquina Bay estuary
Distribution of feedback in models for the pre-invaded and the invaded benthic community of Yaquina Bay
Percent of models with near-zero feedback
LIST OF TABLES Table 2.1
Substrata, vegetation types and macrophyte density of intertidal sites in the Alsea Bay and Yaquina Bay during summer 1993
Summer density and occurrence (OC) of intertidal benthic invertebrates in core samples from Alsea Bay and Yaquina Bay
Summer catch per unit effort (CPUE) and occurrence (OC) of fishes and decapods in the Alsea Bay and Yaquina Bay
Percent of community variance explained, eigenvalues and correlations for the three main axes of CCA ordinations
Total number, mean total length and total weight of juvenile English sole and starry flounder
Species richness and frequency of occurrence of native and nonindigenous (NI) invertebrates in the environment and the diet of English sole and starry flounder
Frequency of species occurrence in the diet of juvenile English sole (E) and starry flounder (S) in Alsea Bay and Yaquina Bay
Percent of number of prey by taxa and species origin for two size classes of juvenile English sole in Alsea Bay and Yaquina Bay
Percent of number of prey by taxa and species origin for three size classes of juvenile starry flounder in Alsea Bay and Yaquina Bay
Walking (WA), swimming (SW) and partially visible (PV) Corophium spp. in sand substrate
Walking (WA), swimming (SW) and partially visible (PV) Corophium spp. in sand and mud substrata
Activity and trophic invertebrate guilds assigned to qualitative models of Yaquina Bay
LIST OF TABLES (Continued)
Guild structure of activity models and assumptions; number of guilds; number of alternative models and invasion status of the community for each community structure
Guild structure of trophic guild models and assumptions; number of guilds and invasion status of each community structure
Alternative activity guild models for the intertidal benthic community of Yaquina Bay
Alternative trophic guild models for the intertidal benthic community of Yaquina Bay
LIST OF APPENDIX FIGURES Figure B.1
Total weight (W) and total length (L) of English sole and starry flounder
Fulton's condition factor of English sole and starry flounder
Percent of dietary overlap (DO±) and trophic breadth (B1 )for flatfish
Number of native and nonindigenous species by volume of individual prey in the diet of juvenile English sole and starry flounder in the Alsea and Yaquina estuaries
mphipod dry weight (Y) with length of 4th article 2nd antenna (X) by species and sex
Mean number of surviving Corophium in single-species predation treatments and in controls without fish .. Mean number of surviving Corophiurn in mixed-species predation experiments in sand and mud treatments and in controls without fish
LIST OF APPENDIX TABLES Table A.1
Summer mean density and overall occurrence of intertidal invertebrates in sediment samples from Alsea Bay
Summer mean density and overall occurrence of intertidal invertebrates in sediment samples from Yaquina Bay
Summer mean catch per unit effort (CPUE) and occurrence of fishes and decapods in Alsea Bay as determined from seine sampling
Summer mean catch per unit effort (CPUE) and occurrence of fishes and decapods in Yaquina Bay as determined from seine sampling
Life-mode and functional-groups of nonindigenous invertebrates found in intertidal and subtidal areas of Alsea and Yaquina Bay
Ratio of English sole with prey (No. of fish with prey in their stomach / No. of fish analyzed); stomach fullness index; mean fish length and weight and mean prey richness (No. taxa) per fish
Ratio of starry flounder with prey (No. of fish with prey in their gut/total No. of fish analyzed); stomach fullness index; mean fish length and weight and mean prey richness (No. taxa) per fish ..
Frequency of prey occurrence and mean number and volume of prey consumed by juvenile English sole in intertidal-subtidal areas of Alsea Bay and Yaquina Bay during summer 1993
Frequency of prey occurrence and mean number and volume of prey consumed by juvenile starry flounder in intertidal-subtidal areas of Alsea Bay and Yaquina Bay during summer 1993
Density and number of Corophium salmonis in the benthos and the diet of juvenile English sole collected in intertidal areas at high tide
Number of prey and their percent frequency of occurrence in stomachs of juvenile staghorn sculpin (Leptocottus armatus)
Activity models derived from models in Figure 5.2
Trophic models derived from models in Figure 5.3
Benthic Biological Invasions in Two Temperate Estuaries and Their Effects on Trophic Relations of Native Fish and Community Stability
Chapter 1 General Introduction
Significance of Biological Invasions Throughout history, humans have moved and released plants, animals and other organisms. Both intended species introductions and inadvertent human activities have greatly increased the
distributional ranges of many aquatic and terrestrial organisms around the world (Elton 1958; Grosholz 1996)
Species moved by
humans into areas outside their natural geographic range are
referred to as nonindigenous species (NIS), non-native, alien or exotic species. Human-mediated biological invasions have caused many of the most dramatic effects on the world's natural communities (Elton 1958; Suter 1993) and are considered the
second most important threat factor after habitat destruction (Sandlund et al. 1999)
However, Crooks and Soulé (1999) state:
"biodiversity losses caused by NIS may soon surpass the damage done by habitat destruction and fragmentation".
The increasing number of introduced aquatic species (e.g., Lachner et al. 1970; Baltz 1991; Li and Moyle 1993) and the apparent exponential rate of invasions in aquatic ecosystems (Cohen and Carlton 1998; Boudouresque 1999), impose unprecedented historical threats to the conservation of freshwater; estuarine; and marine ecosystems (e.g., Carlton and Geller 1993; Moyle 1999)
Concerns about the adverse impacts of NIS have been mostly
focused on short-term impacts, such as losses of marketable goods; the collapse of fisheries; and human-health problems (National Ocean Pollution Program 1991; Carey et al. 1996)
However, proactive management efforts to control transport of species are being increasingly addressed since the implementation of the Nonindigenous Aquatic Nuisance Prevention and Control Act
by the Federal Government in 1990 (e.g., Aquatic Nuisance Species Task Force 1994, Aquatic Nuisance Species Program 1994)
Because of the abiotic and biotic differences between the donor ecosystem (i.e., the source of NIS) and the invaded ecosystem, the effects of species introductions cannot be reasonably predicted, even after accounting for the species niche in the donor system (Nilsson 1985; Li and Moyle 1993) or the impacts of earlier introductions (Williamson and Fitter 1996)
Species introductions are largely irreversible processes (Moyle 1999) and control options for NIS entail further risks and costs (Lafferty and Kuris 1994; Oduor 1999)
The present level of species invasions resulting from natural dispersal mechanism (e.g., Edgpeth 1994) are dwarfed by the magnitude of human-mediated species introductions. Many humanmediated invasions of aquatic organisms can not be accounted for by natural dispersal mechanisms. Examples of the latter include species with life-cycles restricted to brackish-water systems
such as estuaries (e.g., Canton 1979, Cohen and Canton 1995) and enclosed seas (Carlton 1979, Leppakoski 1994)
Mechanisms of Biological Invasions The major recent phyletically and ecologically nonselective vector for the inadvertent dispersal of aquatic organisms is the release of ballast water from ships (Jones 1980; Carlton and Geller 1993). The world's fleet has at least 35,000 ships
transporting ballast water (Canton 1999)
The sheer scale and
magnitude of this vector are such that it has been referred to as "conveyor-belts" exchanging species among otherwise isolated
ecosystems around the world (Canton and Geller 1993)
The use of
ballast water dates from the 1850's and became significant by the 1880's (Stewart 1991). External fouling of ship's hulls also has been recognized as important vectors for the inadvertent introduction of many NIS (Elton 1958; Cohen and Carlton 1995)
The perceived benefit of intentional species introductions led to the spread of species at least over the last 3,000 years (Balon 1974)
Most of the fish introductions in the 19th century
in the United States resulted from the policy of the U.S. Fish Commission to populate the nations' waters with as many useful or valuable food species as possible (Hedgpeth 1980) .
authorized and illegal fish introductions account for 536 fish taxa (species, hybrids and unidentified forms) introduced in inland waters of the United States (Fuller at al. 1999)
other types of aquatic species were intentionally introduced since the 19th century by the aquaculture industry and fishing practices (Welconime 1986, Canton 1992) .
Such introductions have
in turn served as vectors for numerous inadvertent introductions of NIS, including pathogens, competitors, parasites and predators (Stewart 1991; Pillay 1992) . The results of all but a few
intentional aquatic introductions are a mixed blessing (Courtenay and Williams 1992; OTA 1993) and no unintentional aquatic introductions have been found beneficial (Steiner 1992)
Community Susceptibility to Biological Invasions
The type of NIS established in a given system depends on many factors, including: the ecological characteristic of inoculated species (Carlton 1979); their physiological tolerance (Chapman,
In press); their source-regions (i.e., donor-regions, Carlton 1996a); the available vector(s) or mechanism(s) of introduction (Cohen and Carlton 1995)
However, alternative human-mediated
mechanisms of introduction may exist for particular species within phyla ranging from microscopic organisms to conspicuous animals and plants.
Although many attributes of successful
invaders have been identified (e.g., Elton 1958; Ehrlich 1986;
Arthington and Mitchell 1986; Pimm 1989), the predictive capacity of invasion biology is limited. Anticipated invasions in
particular habitats (e.g., Chapman and Carlton 1991 and 1994) and their potential effects (e.g., Grosholz and Ruiz 1996) are still uncommon. Moreover, assessments of impacts of NIS in most cases
is prevented by the lack of appropriate baseline information prior to species invasions (Hedgpeth 1980)
Habitat degradation; pollution or natural environmental changes (e.g., droughts; floods; El Niño events) can lead to more local adaptation of NIS in comparison to many native species. In fact, environmental changes have preceded the detection and/or population expansion of some NIS in estuaries (e.g., Cohen and
Canton 1995; Canton 1996a; G.C. Castillo, personal observation)
The previous patterns are consistent with the
observation that successful invasions of fishes in streams and estuaries are determined by appropriate abiotic factors regardless of the biota already present (Moyle and Light 1996)
Extent of Estuarine and Marine Invasions
The total number of species introductions in aquatic systems is unknown as most research on nonindigenous species has focused on groups such as macroinvertebrates; vascular plants; macroalgae and fishes (e.g., Lachner et al. 1970; Ruiz et al. 1997).
Moreover, over 1,000 species of nearshore marine plants and animals regarded as naturally cosmopolitan may represent pre-1800 century invasions (Carlton 1999)
The latter estimate excludes
non-cosmopolitan NIS with unusual distribution (e.g., Chapman 1988; Chapman and Carlton 1994), many of which may remain unrecognized.
Approximately 400 NIS have been reported along the Pacific,
Atlantic and Gulf coasts of the United States and hundreds of marine and estuarine species are reported in other regions of the world (Ruiz et al. 1997)
Perhaps, the most invaded aquatic
system is the Mediterranean Sea, where at least 300 species from the Red Sea have entered through the Suez Canal since 1869 (Boudouresque 1999)
The decreasing order of reported species
invasions among the most well studied U.S. estuaries is: San Francisco Bay, California (n = 234, Cohen and Carlton 1998);
Chesapeake Bay, Maryland and Virginia (n = 116, Ruiz et al. 1997); Coos Bay, Oregon (n = 60, J.T. Carlton, unpublished data); Puget Sound, Washington (n = 52, Cohen et al. 1998)
Potential Impacts of Nonindigenous Species Despite the complex effects of both natural and anthropogenic disturbances on fish feeding and growth (e.g., Livingston 1980; Choat 1982; Sogard 1994), introduced species that become established alter food webs and possibly the functions of the invaded ecosystem (Li and Moyle 1981; Pirnm 1982; Li et al. 1999)
Biological invasions may be energetically significant as food chain efficiencies can vary over two or more orders of magnitude (May 1979) and the caloric content of species vary significantly among phyla (Thayer et al. 1973) and within phyla (Padian 1970)
Moreover, species that contribute the most to overall biomass may not be the most important food sources for higher trophic levels.
Invasions in U.S. West Coast Estuaries
Virtually no information exists on the numbers of NIS in most U.S. west coast estuaries and less information in available on the percentage of NIS that can be considered "nuisance species" (i.e., species that affect the abundance of native species by
competing or preying on them (National Ocean Pollution Program 1991)
Nevertheless, biological invasions may have dramatically
changed the densities of native species in some estuaries. For example, the NI (nonindigenous) bivalves Potamocorbula amurensis and Corbicula fluminea can filter large amounts of phytoplankton (Cohen et al. 1984; Nichols et al. 1990) . Zooplankton consumption
by P. amurensis may also be a direct cause for the significant declines in zooplankton of San Francisco Bay (Kimrnerer et al. 1994) .
These findings support the hypothesis that P. arnurensis
may have irreversibly changed the ecosystem dynamics of that estuary by displacing a predominantly planktonic community with a
predominantly benthic community (Nichols et al. 1990; L.W. Miller, personal communication 1990)
Other nuisance species include the cordgrass Spartina alterniflora which has dramatically reduced the extent of intertidal mudflat habitats by excluding other plants; invertebrates; fishes; and shorebirds in some U.S. west coast estuaries (Strong 1997) and the NI predatory snail Ocenebra japonica introduced with oyster spat caused the collapse of the oyster fishery in Netarts Bay, Oregon (Kreag 1979) . Cohen and
Canton (1995) reported many other NIS that may be considered nuisance species in San Francisco Bay. Some NIS may not be clearly considered nuisance species despite their substantial effects on the invaded habitats. Such seems to be the case of the eelgrass Zostera japonica, which has changed the physical habitat and increased both the richness and densities of fauna in the South Slough of Coos Bay, Oregon (Posey 1988) . The effects of
many other potential nuisance species are yet to be evaluated, including at least six
sian copepods (Cordell and Morrison 1996)
and the European green crab Carcinus maenas which invaded many Northeast Pacific estuaries since the mid 1990s (Cohen et al. 1995; Miller 1996; Beherens and Hunt, in press)
Estuarine fishes are typically assumed to rely on opportunistic use of prey (e.g., Barnes 1974; Day et al. 1989).
However, juvenile salmonids may have adapted their spatiotemporal use of the Squamish estuary, British Columbia, to the production of Eogammarus confervicolus (Levings 1980) .
salmon (Oncorhynchus keta) at the Nanaimo River estuary, may be in near balance with its major prey Harpacticus uniremus (Healey 1979) .
Moreover, information on noncoevolved predator-prey
relations suggest a selective prey pattern. In the Sacramento-San Joaquin delta, California, Herbold (1987) found that when the native shrimp Neomysis mercedis migrates in the fall, or its density is reduced by fish predation, native fishes switch to other prey, while NI fishes continue to feed largely on the shrimp. In the latter system, larvae of the introduced striped
bass (Morone saxatilis) may be less selective on two introduced noncoevolved copepods in comparison to at least one introduced coevolved copepod species (e.g., Meng and Orsi 1991).
Focus of the Present Thesis
This thesis addresses ecological aspects of benthic biological invasions in intertidal areas of the Alsea Bay and Yaquina Bay, two estuaries in the central Oregon coast (USA) . No studies have
assessed the extent of NIS invasions; their trophic effects on native fishes and their potential impacts on community stability in these estuaries. Major questions addressed in the following four chapters are:
1) Are environmental characteristics of intertidal
habitats available to NIS different between estuaries?, 2) Are total densities and richness of native species and NIS different between estuaries?, 3) Are taxonomically close native species and NIS distributed in common assemblages?, 4) How do the total abundances and richness of native and NI invertebrates vary under various temperature-salinity combinations?, and 5) Are native species and NIS similarly distributed across environmental gradients?
The general ecological patterns of biological invasions in chapter 2 provide the necessary context to link all additional chapters.
Is the richness of native species and NIS in the
environment proportional to the richness of native species and NIS in the diets of native fishes?, 2) What is the contribution of native species and NIS to the food-base of native fishes?, and 3)
Is the overall prey selection by native fishes similar between
native and NI prey types?
The evidence of noncoevolved predator-prey relations presented in chapter 3 is further evaluated to determine factors
controlling prey selection (chapter 4), and community interactions of native benthic fishes (chapter 5) Chapter 4: 1) Are there differences in visibility and activity
among taxonomically close native and NI prey?, 2) Does prey consumption by native benthic fishes vary with species, size, or sex of prey?, and 3) Does predator consumption and selection of prey vary with prey origin or substratum type? Prey behavior and predator selection experiments in chapter 4 provide an independent evaluation of predator selection on noncoevolved prey, allowing comparison with the field data reported in chapter 3.
1) Have biological invasions induced changes in the
stability of benthic estuarine communities?, and 2) What is the potential role of native fish predators in maintaining community stability characteristics?
Information from the preceding chapters is synthesized here to address the two previous questions using alternative functionalgroup interactions models in the benthic community of Yaquina Bay.
Alsea Bay and Yaquina Bay are partially-mixed drowned river estuaries with similar morphological and physical characteristics (Bottom et al. 1979)
Yaquina Bay is the fourth largest estuary
in Oregon (c.a. 16 km2 at mean high tide) and it has a drainage basin of 655 kin2
(Percy et al. 1974) . Mean tidal range is 1.80 m
and the tidal prism on mean range (i.e., the volume between high and low water level) is 23.64 x 106 m3 (Johnson 1972) . Alsea Bay
is 25 km south of Yaquina Bay. It is the seventh largest estuary in Oregon (c.a.
at mean high tide) and has a drainage basin
of 1,228 km2 (Percy et al. 1974). Mean tidal range is 1.77 m (Johnson 1972) and its tidal prism on mean range is 14.16 x 106 m3 (Goodwin et al. 1970) . Unlike Alsea Bay, jetties and a dredged
main channel are maintained in Yaquina Bay (Cortright et al. 1987) and only Yaquina Bay has been exposed to ballast water
traffic. However, both estuaries have been used for culture of Atlantic and Pacific oysters since the late part of 19th century
Distribution and density of intertidal assemblages of macrobenthic invertebrates and fishes are described in chapter 2. The diet composition and prey selection of the native pleuronectids English sole (Pleuronectes vetulus) and starry flounder
(Chapter 3) .
stellatus) are considered in field analyses
Behavior of two native and two NI amphipods
(Corophium spp.) and consumption and selection of the latter prey
by juvenile English sole is further considered in laboratory experiments (Chapter 3)
Stability patterns before and after NIS invasions in the benthic community of Yaquina Bay are estimated using two types of functional group interaction models. Namely, activity models,
which emphasize physical interactions among invertebrates, and trophic models, which emphasize direct and indirect trophic interactions among invertebrates. English sole, starry flounder and the native fish staghorn sculpin (Leptocottus armatus) are the major benthic predators included in all models along with both native and NI benthic macroinvertebrate prey (Chapter 5)
Aquatic Nuisance Species Program. 1994. Aquatic Nuisance Species Task Force. Washington, D.C., U.S. Government Printing Office 1996-508-0889. 60 pp. + Appendices A-G. Aquatic Nuisance Species Task Force. 1994. Findings, Conclusions, and Recommendations of the Intentional Introductions Policy Review. Report to Congress. Washington D.C. 53 pp. Arthington, A.H. and D.S. Mitchell. 1986. Invading aquatic species. In Ecology of Biological Invasions. An Australian Perspective eds. Groves, R.H. and J.J. Burdon, 34-53. Australian Academy of Science. Canberra Balon, E.K. 1974. Domestication of the carp, Cyprinus carpio L. Royal Ontario Museum. Miscellaneous Publications. Toronto. Baltz, D.M. 1991. Introduced fishes in marine systems and inland seas. Biological Conservation 56:151-177. Barnes, R.K. 1974. Estuarine biology. Studies in Biology 49. Edward Arnold, London, England.
Beherens, S.Y. and C. Hunt. In press. The arrival of the European green crab Carcinus maenas in the Pacific Northwest. Dreissena 11(1).
Boudouresque, C.F. 1999. The Red Sea - Mediterranean link: unwanted effects of canals. In Invasive Species and Biodiversity Management, eds. Sandlund, O.T., P.J. Schei and A. Viken, 213-228. Kiuwer Academic Publishers, Dordrecht, Netherlands. Carey, J.R., P.B. Moyle, M. Rejmánek and G. Vermeij. 1996. Preface. Biological Conservation 78:1-2.
Canton, J.T. 1979. History, biogeography and ecology of the introduced marine and estuarine invertebrates of the Pacific Coast of North Imerica. Ph.D. dissertation. University of California, Davis, 904 pp. Carlton, J.T. 1992. Dispersal of living organisms into aquatic environments as mediated by aquaculture and fisheries activities. In Dispersal of living organisms into aquatic ecosystems, eds. A. Rosenfield and R. Mann, 13-46. A Maryland Sea Grant Publication, College Park Maryland.
Canton, J.T. l996a. Pattern, process, and prediction in marine invasion ecology. Biological Conservation 78: 97-106.
Canton, J.T. 1999. The scale and ecological consequences of biological invasions in the world's oceans. In Invasive Species and Biodiversity Management, eds. Sandlund, O.T., P.J. Schei and A. Viken, 195-212. Kluwer Academic Publishers, Dordrecht, Netherlands. Canton, J.T. and J.B. Geller. 1993. Ecological roulette: The global transport of nonindigenous marine organisms. Science 261:78-82.
Chapman, J.W. 1988. Invasions of the Northeast Pacific by Asian and Atlantic garnmaridean amphipod crustaceans, including a new species of Corophium. Journal of Crustacean Biology 8:364-382. Chapman, J.W. In press. Climate and nonindigenous peracaridan crustaceans in northern hemisphere estuaries. In National Conference on Marine Bioinvasions, ed. J. Pederson, Proceedings, January 1999. Massachusetts Sea Grant. Massachusetts Institute of Technology. Cambridge, Massachusetts.
Chapman, J.W. and J.T. Canton. 1991. A test of criteria for introduced species: The global invasion by the isopod Synidotea laevidorsalis. Journal of Crustacean Biology 11:386400.
Chapman, J.W. and J.T. Canton. 1994. Predicted discoveries of the introduced isopod, Synidotea laevidorsalis (Miers, 1881) Journal of Crustacean Biology 14:700-714. Choat, J.H. 1982. Fish feeding and the structure of benthic communities in temperate waters. Annual Review of Ecology and Systematics 13:423-449.
Cohen, A.N. and J.T. Canton 1995. Nonindigenous aquatic species in a United States estuary: A case of the biological invasions of the San Francisco Bay and Delta. Biological Study. A Report for the U.S. Fish and Wildlife Service, Washington, D.C. and the National Sea Grant College Program, Connecticut Sea Grant. Cohen, A.N. and J.T. Canton. 1998. Accelerating invasion rate in a highly invaded estuary. Science 279:555-558. Cohen, A.N.,, J.T. Canton and M.C. Fountain. 1995. Introduction, dispersal and potential impacts of the green crab Carcinus maenas in San Francisco Bay, California. Marine Biology 122: 225-237.
Cohen, A., C. Mills, H. Berry, M. Wonham, B. Bingham, B. Bookheim, J. Canton, J. Chapman, J. Cordell, L. Harris, T. Klinger, A. Kohn, C. Lambert, G. Lambert, K. Li, D. Secord and J. Toft. 1998. Report of the Puget Sound Expedition. September 8-16, 1998. A rapid assessment survey of non-indigenous species in the shallow waters of Puget Sound. Washington Department of Natural Resources, Olympia, WA. U.S. Fish and Wildlife Service, Lacey, WA. 37 pp. Cohen, R.R.H., P.V. Dresler, E.J.P. Phillips, and R.L. Cory. 1984. The effect of the Asiatic clam, Corbicula fluminea, on phytoplankton of the Potomac River, Maryland. Limnology and Oceanography 29:170-180. Cordell, J.R. and S.M. Morrison. 1996. The invasive Asian copepod Pseudodiaptomus inopinus in Oregon, Washington, and British Columbia estuaries. Estuaries 19:629-638. Cortright, R., J. Weber and R. Bailey. 1987. The Oregon estuary plan book. Oregon Department of Land Conservation and Development, 126 pp. Courtenay, W.R. Jr. and J.D. Williams. 1992. Dispersal of exotic species from aquaculture sources, with emphasis on freshwater fishes. In Dispersal of living organisms into aquatic ecosystems, eds. Rosenfield., A. and R. Mann, 49-81. A Maryland Sea Grant Publication, College Park, Maryland. Crooks, J.A. and N.E. Soulé. 1999. Lag times in population explosions of invasive species: causes and implications. In Invasive Species and Biodiversity Management, eds. Sandlund, O.T., P.J. Schei and A. Viken, 103-125. Kiuwer Academic Publishers, Dordrecht, Netherlands. Day, J.W.Jr., C.A.S. Hall, W.M. Kemp and A. Yañez-Arancibia. 1989. Estuarine Ecology. John Wiley & Sons, New York. 558 pp.
Ehrlich, P.R. 1986. Which animal will invade? In Ecology of biological invasions of North America and Hawaii, eds. Mooney, H.A. and J.A. Drake, 79-95. Springer, New York. Elton, C.S. 1958. The ecology of invasions by animals and plants. Reprint 1972, Chapman & Hall, London, 181 pp. Fuller, P..L., L.G. Nico and J.D. Williams. 1999. Nonindigenous fishes introduced into inland waters of the United States. U.S. Geological Survey, Biological Resources Division. Florida Caribbean Science Center. Bethesda, Maryland. 613 pp.
Goodwin, C.R., E.W. Emmet and B. Glenne. 1970. Tidal study of three Oregon estuaries, Engineering Experiment Station. Bulletin 45. Oregon State University, Corvallis, Oregon. 33 pp.
Grosholz, E.D. 1996. Contrasting rates of spread for introduced species in terrestrial and marine systems. Ecology 77: 16801686.
Grosholz, E.D. and G. Ruiz. 1996. Predicting the impact of introduced marine species: lessons from the multiple invasions of the European Green crab Carcinus maenas. Biological Conservation 78:59-66. Healey, M.C. 1979. Detritus and juvenile salmon production in the Nanaimo estuary. I. Production and feeding rates of juvenile churn salmon (Oncorhynchus keta) Journal of the Fisheries Research Board of Canada. 36:488-496. .
Hedgpeth, J.W. 1980. The problem of introduced species in management arid, mitigation. Helgo1nder Meeresuntersuchungen 33:662-673. Hedgpeth, J.W. 1994. Nonanthropogenic dispersals and colonization in the sea. In Nonindigenous Estuarine & Marine Organisms (NEMO), 45-62. Proceedings of the Conference & Workshop. April 1993. U.S. Department of Commerce. Seattle, Washington. Herbold, B. 1987. Resource partitioning with a non-coevolved assemblages of fishes. Ph.D. Dissertation. University of California, Davis. Johnson, J.W. 1972. Tidal inlets on the California, Oregon, and Washington coasts. Hydraulic Engineering Laboratory HEL 24-12. University of California, Berkeley, California, 56 pp. Jones, M.M. 1991. Marine organisms transported in ballast water. A review of the Australian Scientific Position. Bureau of Rural Resources. Australian Government Publishing Service. Canberra. Bulletin No. 11, 48 pp.
Kimmerer, W.J., E. Gartside and J.J. Orsi. 1994. Predation by an introduced clam as the likely cause of substantial declines in zooplankton of San Francisco Bay. Marine Ecology Progress Series 113: 81-93. Kreag, R.A. 1979. Natural resources of Netarts estuary. Final Report. Estuary Inventory Project. Oregon Department of Fish and Wildlife, Portland, OR, 45 pp. Lachner, E.A., C.R. Robins, and W.R. Courtenay. 1970. Alien fishes and other aquatic organisms introduced into North Pmerica. Smithsonian contributions to Zoology 59:1-29. Lafferty, K.D. and A.M. Kuris. 1994. Potential uses of biological control of alien marine species. Proc. Nonindigenous Estuarine and marine Organisms. U.S. Department of Commerce, NOAA office of the chief Scientist. pp. 129-150.
Leppakoski, E. 1994. The Baltic and the Black Sea seriously contaminated by nonindigenous species? In Nonindigenous Estuarine & Marine Organisms (NEMO), 37-44. Proceedings of the Conference & Workshop. April 1993. U.S. Department of Commerce. Seattle, Washington. Levings, C.D. 1980. The biology and energetics of Eogammarus confervicolus (Stimpson) (Amphipoda, Anisogammaridae) at the Squamish River Estuary, B.C. Canadian Journal of Zoology 58:1652-1663. Li, H. W. and P. B. Moyle. 1981. Ecological analysis of species introductions into aquatic ecosystems. Transactions of the American Fisheries Society 110:772-782. Li, H.W. and P.B. Moyle. 1993. Management of introduced fishes. In Inland Fisheries Management in North America, eds. Kohier, C.C. and W.A. Hubert, 287-307. American Fisheries Society. Li, H.W., P.A. Rossignol and G. Castillo. 1999. Risk analysis of species introductions: insights from qualitative modeling. In Nonindigenous freshwater organisms, vectors, biology and impacts, eds. Claudi, R. and J.H. Leach, 431-447. CRC Press. Boca Raton, Florida. Livingston, R.J. 1980. Ontogenetic trophic relationships and stress in a coastal seagrass system in Florida. In Estuarine Perspectives, ed. V.5. Kennedy, 423-435. Academic Press. N.Y. USA.
May, R.M. 1979. Production and respiration in animal communities. Nature 282:443-444.
Meng, L. and J.J. Orsi. 1991. Selective predation by larval striped bass on native and introduced copepods. Transactions of the American Fisheries Society 120(2):187-192. Miller, L.D. 1990. Personal communication. 1990. Department of Fish and Game. Bay Delta Project. 4000 N. Wilson Way. Stockton, CA 95205. Miller, T.W. 1996. First record of the green crab, Carcinus maenas, in Humboldt Bay, California. California Fish and Game 82: 93-96.
Moyle, P.3. 1986. Fish introductions into North America. In Ecology of biological invasions of North America and Hawaii eds. H.A. Mooney and J.A. Drake. Ecological Studies 58:27-43. Springer-Verlag. Moyle, P.B. 1999. Effects of invading species on freshwater and estuarine ecosystems. In Invasive Species and Biodiversity Management, eds. Sandlund; O.T., P.J. Schei and A. Viken, 177191. Kiuwer Academic Publishers, Dordrecht, Netherlands.
Noyle, P.B. and Light. 1996. Fish invasions in California: Do abiotic factors determine success? Ecology 77:1666-1670. National Ocean Pollution Program. 1991. Understanding the sources, fates, and effects on aquatic organisms of pathogens and nuisance species that are introduced or influenced by human activities. In Chapter IV. Federal plan for ocean pollution, research, development, and monitoring. Fiscal Years 1992-1996. Pages 38-142. Prepared by the National Ocean Pollution Program Office for the National Ocean Pollution Policy Board. September 1991. U.S. Department of Commerce.
Nichols, F.H., J.K. Thompson, and L.E. Schemel. 1990. Remarkable invasion of San Francisco Bay (California, USA) By the Asian clam Potamocorbula amurensis. II. Displacement of a former community. Marine Ecology Progress Series 66: 95-101. .
Nilsson, N.A. 1985. The niche concept and the introduction of exotics. National Swedish Board of Fisheries. Institute of Freshwater Research. Drottningholm, Lund. Report No. 62:128135.
Oduor, G.I. 1999. Biological pest control for alien invasive species. In Invasive Species and Biodiversity Management, eds. Sandlund, O.T., P.J. Schei and A. Viken, 305-321. Kluwer Academic Publishers, Dordrecht, Netherlands. OTA. 1993. Harmful non-indigenous species in the United States. Office of Technology Assessment. U.S. Congress. U.S. Government Printing Office, OTA-F-565, Washington, D.C.
Pandian, T.J. 1970. Intake and conversion of food in the fish Limanda limanda exposed to different temperatures. Marine Biology 5:1-17. Percy, K., C. Sutterlin, D.A. Bella and P.C. Klingeman. 1974. Description and information sources for Oregon estuaries. Sea Grant College Program. Oregon State University, Corvallis, Oregon, 294 pp.
Pillay,T.V.R. 1992. Introduction of exotics and escape of farmed fish. In Aquaculture and the environment, 78-88. University Press, Cambridge, Great Britain. Pimm, S.L. 1982. Food webs. Chapman & Hall, J. W. Arrowsmith Ltd., Bristol, Great Britain, 219 pp. Pimm. S.L. 1989. Theories of predicting success and impact of introduced species. In Biological Invasions. A global Perspective, eds. Drake, J.A., H.A. Mooney, F.di Castri, R.H. Groves, F.J. Kruger, M. Rejmanek and M. Williamson, 351-367. Scope 37. John Wiley & Sons. Chichester.
Posey, M.H. 1988. Community changes associated with the spread of an introduced seagrass, Zostera japonica. Ecology 69:974-983. Ruiz, G.M., J.T. Carlton, S.D. Grosholz and A.H. Hines. 1997. Global invasions of marine and estuarine habitats by nonindigenous species: mechanisms, extent and consequences. American Zoologist 37:621-632. Sandlund, O.T., P.J. Schei and A. Viken. 1999. Introduction: the many aspects of the invasive alien species. In Invasive Species and Biodiversity Management, eds. Sandlund, O.T., P.J. Schei and A. Viken, 1-7. Kluwer Academic Publishers, Dordrecht, Netherlands. Sogard, S.M. 1994. Use of suboptimal foraging habitats by fishes: consequences to growth and survival. In Theory and application in fish feeding ecology, eds. D.J. Stouder, K.L. Fresh, and R. J. Feller, 103-131. University of South Carolina Press. Columbia, South Carolina. Steirer, F.S., Jr. 1992. Historical perspective on exotic species. In Introductions and transfers of marine species, ed. M.R. De Voe,1-4. South Carolina Sea Grant Consortium, Charleston. South Carolina. Stewart, J.E. 1991. Introductions as factors in diseases of fish and aquatic invertebrates. Canadian Journal of Fisheries and Aquatic Sciences 48: 110-117. Strong, D. 1997. Spartina in the San Francisco Bay region. In American Fisheries Society 127th Annual Meeting. Fisheries at Interfaces: Habitats, Disciplines, Cultures. 24-28 August 1997. Monterey, California. Abstracts L-Z:84. Suter, G. 1993. Exotic organisms. In Ecological risk assessment, ed. G.W. Suter 11,3 91-401. Lewis Publishers, Boca Raton, Florida. Thayer, G.W., W.E. Schaaf, J.W. Angelovic and M.W. LaCroix. 1973. Caloric measurements of some estuarine organisms. Fishery Bulletin, U.S. 71:289-296.
Welcomme, R.L. 1986. International measures for the control of introductions of aquatic organisms. Fisheries 11:4-9. Williamson, M. and A. Fitter. 1996. The varying success of invaders. Ecology 77:1661-1666.
Distribution and Habitat Use by Noricoevolved Macroinvertebrates and Fishes in Two Temperate Estuaries
G.C. Castillo"2, H.W. Li2, J.W. Chapman3, T.W. Miller3
Present address: Hatfield Marine Science Center. Oregon State University, Newport, OR 97365. 2
Oregon Cooperative Fish and Wildlife Research Unit, Department of Fisheries and Wildlife. Oregon State University Corvallis, OR 97331.
Hatfield Marine Science Center. Oregon State University, Newport, OR 97365.
We determined the species richness, densities of benthic macroinvertebrates, cath per unit effort (CPUE) of fishes and decapod crustaceans and environmental relations during summer in intertidal areas of two intermediately invaded estuaries, the Alsea Bay and Yaquina Bay (Oregon, USA) .
We find higher densities
and richness of nonindigenous species (NIS) of invertebrates in the deeper estuary exposed to ballast-water traffic (Yaquina Bay)
All eight introduced invertebrates in Alsea Bay co-occur in
Yaquina Bay. In the latter estuary only the polychaete Streblospio benedicti is common among the three NIS of
invertebrates not detected in Alsea Bay. The only NIS of fish are Alosa sapidissima and Lucania parva, both species are uncommon and occur only in Yaquina Bay. We attribute the high co-
occurrence of NIS between estuaries primarily to oyster-mediated invasions and secondarily to potential dispersal of NIS by currents. The higher densities and richness of NIS in Yaquina Bay could be due to: longer history of oyster reintroductions, shiptraffic, and/or better conditions for NIS in the more disturbed and polluted habitats of Yaquina Bay. Noncoevolved interactions among similar taxa may not be more likely when compared to distantly related taxa. Highest mean densities of NIS of invertebrates at low- and high-tide coincided with: 1) high-mid
temperatures in both estuaries, 2) mid salinities in Alsea Bay and 3) mid-low salinities in Yaquina Bay.
Most of the population
variations of invertebrates and fishes in intertidal areas at high-tide are accounted for by macrophyte density, water temperature and salinity. High values for the latter three environmental factors are associated with greater NIS densities in most invertebrates. The CPUE of native fishes and decapod crustaceans do not vary with invertebrate densities in sediment samples. Further species introductions should be prevented if dominance of native species and their potential ecological functions are to be maintained.
The human-mediated dispersal of nonindigenous species (NIS)
around the world has produced severe effects on aquatic communities (Elton 1958; Baltz 1991; Li and Moyle 1993; Cohen and Canton 1998) . Many aquatic organisms have been introduced
through aquaculture and fisheries activities (Canton 1992)
transport of ballast water from ships is recognized as the major recent human-mediated vector for the movement of aquatic organisms within and between oceans (e.g., Williams et al. 1988;
Jones 1991; Canton and Geller 1993; Smith et al. 1996) Sediments carried in ballast tanks and fouling organisms
externally attached to ship's hulls are also potential vectors of
species introductions (Canton 1996a) Over 234 NIS are established in Pacific coast estuaries of North America, where they often are the dominant macrofauna
(Canton 1979; Cohen and Canton 1995)
With few exceptions,
nonindigenous (NI) coastal invertebrates are restricted to calm-
water embayments, estuaries and harbors (Canton 1979)
establishment of NI invertebrates may be related to: absence of competition with native species, creation of novel habitats by humans to which only certain NIS are adapted, competitive
displacement of native species by NIS (Canton 1979), noncompetitive species interactions (e.g., Cohen and Carlton 1995), and reduced community response to invasions (Castillo et al. 2000)
Considering that the type and degree of species
interactions in benthic communities is influenced by the ecological similarity among species (e.g., Whitlatch 1980; Woodin 1983), the need for comparing the distribution of noncoevolved taxa seems critical to infer which groups of organisms are more likely to interact in estuaries.
Northeast Pacific estuaries are nursery grounds for many native species of fishes and invertebrates (e.g., Haertel and Osterberg 1967; Pearcy and Myers 1974; Bayer 1981; De Ben et al. 1990; Bottom and Jones 1990; Jones et al. 1990)
impacts of NIS invasions on these rearing areas is virtually
unknown. Ballast water sampling from 159 cargo ships arriving from Japanese ports to Coos Bay, Oregon, revealed 367 taxa, many not identified to species level, including all major and most minor phyla (Carlton and Geller 1993)
Only four of the 60 known
established NIS in Coos Bay have been ascribed primarily to
ballast water discharge (J.T. Canton, unpublished data) Nevertheless, ballast water in that estuary may have reintroduced many NIS established by earlier vectors such as oyster culture and external fouling. Alternatively, many established NIS introduced by ballast-water release may remain unrecognized.
Estimates of the risk of species invasions in estuaries have been prevented by the effort required to monitor vectors of species
introductions and species invasions (Canton l996a)
to estimate the extent of invasions by different vectors is to
compare estuaries that have historically differed in vectors of species invasions.
We surveyed intertidal areas in two invaded estuaries that differ in their risk of ballast-water mediated species introductions. We ask five questions: 1) Are the environmental
characteristics of intertidal habitats available to NIS greatly different between estuaries?; 2) Are total densities and richness of NIS and native species greatly different between estuaries?;
3) Are taxonomically close native species and NIS distributed in common assemblages?; 4) How do the total abundances and richness of native and NI invertebrates vary under various temperaturesalinity combinations?; and 5) Are native species and NIS similarly distributed across downstream to upstream areas?
Our objectives are to provide
answers to these questions based on surveys conducted in the Alsea Bay and Yaquina Bay estuaries on the central Oregon coast, USA (Figure 2.1)
Unlike Yaquina Bay, Alsea Bay is a not a port
for cargo vessels or commercial fishing. Between 1960 and 1969
Yaquina Bay received 848 thousand metric tons of shipping traffic (Percy et al. 1974)
Since the l870s both estuaries were used for
culturing introduced Atlantic oyster (Crassostrea virginica) and subsequently Pacific oysters from Japan (C. gigas), two
Figure 2.1. Alsea Bay and Yaquina Bay estuaries. Indicated are the intertidal sites where benthic invertebrates and fishes were sampled and the means and ranges of salinity, water temperature and transparency at high-tide (S) and low-tide (0) during summer 1993. Species were also sampled at low-tide in sites marked with an asterisk.
potentially important vectors for additional inadvertent species introductions to Northeast Pacific estuaries (Carlton 1979)
Although pre-invasion data on species composition and densities are not available for Alsea Bay and Yaquina Bay, our study provides a baseline for evaluating future community changes. Alsea Bay and Yaquina Bay are drowned river estuarine einbayments (Bottom et al. 1979)
The classes of intertidal and
adjacent subtidal habitats in these two estuaries are characterized by unconsolidated shores and aquatic beds (e.g., Cowardin et al. 1979)
Yaquina Bay is surrounded by more
development than Alsea Bay and only Yaquina Bay is dredged annually. Nearly 54% of the 16 km2 of the surface of Yaquina Bay is intertidal (Hamilton 1973; Cortright et al. 1987)
is well-mixed from summer to winter and partly-mixed in spring (Burt and McAllister 1959)
Average depth is about 6 m and tidal
effects extend 42 km upstream (Percy et al. 1974)
Alsea Bay is
25 km south of Yaquina Bay. Alsea Bay averages less than 2 m in depth and tidal effects extend 26 km upstream (Percy et al. 1974)
Nearly 71% of the 9 km2 of the surface of Alsea Bay is
intertidal (Hamilton 1973; Cortright et al. 1987) and it is well mixed during summer (Burt and McAllister 1959)
We conducted four intertidal surveys of benthic invertebrates, fishes and large epibenthic invertebrates during summer 1993. Summer coincides with the highest use of Oregon estuaries by fishes (Bayer 1981; De Ben et al. 1990) and with highest population densities of macrobenthos (Walker 1973)
estuary we sampled six high-tide sites during afternoon hours (Alsea Bay: Al-A6; Yaquina Bay: Y1-Y6)
Three of the latter sites
were sampled at low-tide during daylight morning hours (Alsea
Bay: al, a3, a6; Yaquina Bay: y2, y4, y6) . Figure 2.1. Each
estuary was surveyed within the following periods: July 5-7, July 18-19, August 2-3 and September 16-17, 1993.
For each site we collected invertebrates in core sediment samples along three transects parallel to the shore. Each 30 in
transect was sampled using 10 equally spaced sediment cores (core diameter: 3.2 cm, depth: 13 cm) . The transects were at depths of 0,
40, and 80 cm high-tide level. Core samples from each transect
and tide are composited and washed on a 500 pm sieve, fixed in 5% buffered formalin and stained with Rose Bengal. Species identification was possible for 67% of the taxa and identified species comprised over 92% of the most abundant taxa.
Fishes and large epibenthic invertebrates were collected with a beach seine (32 m L x 1.8 m H and 0.8 cm stretched mesh size) Seining area encompassed a 163 m2 semicircle between the water line (0 m depth) and deeper areas at high- and low-tide. All species collected in beach seine were identified.
Water temperature, salinity (refractometer based) and water transparency (Secchi disk diameter: 20 cm) were determined at each site immediately prior to seining and sediment sampling.
All intertidal sites are qualitatively classified by macrophyte density. The latter is ranked from lowest (rank 1) in predominantly muddy substratum to highest (rank 5) in predominantly sandy substratum based on the presence and amount of aquatic vegetation as follows: 1) algae are rare and no
eelgrass is present; 2) both algae and eelgrass are present and rare; 3) either algae or eelgrass are common or both are common but not abundant; 4) algae are common and eelgrass are abundant or viceversa; 5) both algae and eelgrass are abundant.
Species origins are assessed from criteria for introduced species (Carlton 1979; Chapman 1988; Chapman and Carlton 1991; Chapman and Carlton 1994) and from reported species introductions in 13.5. west coast estuaries (Carlton 1979; Lee et al. 1980;
Canton and Geller 1993; Cohen and Carlton 1995) . Taxa identified
to species-level but of unknown origin are referred as
cryptogenic (Canton 1996b) .
Invertebrates not identified to
species level are classified as supra-specific taxa and may include native and/or NIS.
Data Analyses Faunal densities in core sediments and catch per unit effort in beach seine samples (hereafter CPUE) are converted, respectively, to numbers of individuals per m2 and 1000 m2
Differences in mean faunal density, CPUE and richness within each estuary are evaluated by single-factor ANOVA. Two-factor ANOVA is used to evaluate faunal differences in density (or CPUE) and richness: 1) between estuaries and among months, and 2) among sites and core sample transects. Pair-wise associations among faunal densities, CPUE, richness and environmental factors are
evaluated through Spearman's rank correlations (Devore and Peck 1986)
Hierarchical clusters of taxa and sites are derived from Statgraphics Plus 2.1. using Euclidean distance (Ludwig and Reynolds 1988) and Ward's linkage method (Ward 1963). Species
densities or CPUE in each site are grouped by origin in major taxa (e.g., native polychaeta; NI polychaeta). Mean summer densities (D) or CPUE are log-transformed, log10(D+1) or log10
(CPUE +1), and similarities among taxa and sites are inferred from the respective clusters.
Mean total densities of native and NI invertebrates in core sediment samples were plotted in 3D graphs (x,
z axes) against
representative midpoints of salinities (5: 1-9 °/, 15: 10-19 25: 20-29 0/ < 35: 30-34.9 0/) and temperatures (15: 13,
18: 17-19 °C, 21: 20-23 °C)
Such range of temperature-
salinity combinations were derived from all high- and low-tide sites sampled in the four surveys in each estuary. Percent of mean densities and richness of NIS relative to native species (i.e., NIS + native species = 100%), are used to compare
dominance patterns under each temperature-salinity combination.
We used canonical correspondence analysis (hereafter CCA, Ter Braak 1986) to describe how intertidal species densities respond to environmental gradients. CCA was selected for synthesizing species patterns at high-tide as this ordination method excels at representing data sets where species responses to important environmental variables are unimodal (e.g., hump-shaped response surfaces, McCune 1997) . Species and sites are indicated by points
representing dominant patterns in community composition as explained by environmental factors. The latter are represented by vectors whose directions indicate increasing value of each environmental factor. The vector's origin corresponds to the mean of each environmental factor. The center of distribution for a given species along each environmental factor is inferred by plotting a perpendicular line from the corresponding vector to the species point.
The influence of rare species in the CCA analyses was reduced by including only those species found in three or more intertidal sites. Environmental factors included in CCA are alternative
combinations of temperature, salinity, water transparency and macrophyte density (Figure 2.1; Table 2.1) .
considered in the ordination of species from beach seine collections are CPUE of native species, NIS and all species in core samples. Mean summer densities of invertebrates in core
samples (D) and CFUE of species in seine samples are respectively transformed as D'13 and log13(CPUE+1) to reduce the influence of
dominant species. We used different transformations as CPUE differences among species were more extreme in comparison to species densities. Ordination scores for species and sites are standardized to mean zero and variance one and species scores are treated as weighted mean site scores to allow direct spatial interpretation of the relations between species points and environmental factors (McCune and Mefford 1997) . To compare the
importance of each environmental factor in structuring the ordinations, intraset correlations between environmental factors and the two main CCA axes are indicated in ordination plots. To evaluate the probability of spurious community-environmental
Table 2.1. Substrata, vegetation types and macrophyte density of intertidal sites in the Alsea Bay and Yaquina Bay during summer 1993. Upstream km indicates the kilometers from the river mouth to each site. Amount of aquatic vegetation: A = abundant, C = common, R = rare, not present = N. Macrophyte density increases from 1 to 5.
Aquatic vegetation Algae
Alsea Bay Al
Sand/Polychaete tubes Sand/Mud/Clay Sand/Mud Sand/Mud/Clay
Yaquina Bay Yl
Sand/Mud/Cobble Sand/Cobble/Clay Mud
A A A
relations, Monte Carlo analyses are used to test the null
hypothesis of no relation between the community matrix and the environmental matrix. The latter analyses are based on 1,000 runs of randomized data and were computed along with CCA analyses in PC-ORD 3.0 (McCune and Mefford 1997)
Macrophyte density was lowest in upstream sites and varied similarly between estuaries (Table 2.1; r = 0.98, p < 0.001, n = 6), and it increased with salinity (r = 0.78, p < 0.01, n = 12). Salinity and temperature are inversely related in both estuaries (Alsea Bay: r = - 0.71,
p < 0.001; Yaquina Bay: r = -0.43, p
0.20, ANOVA). Monthly
differences in richness and total density are not apparent for native species and NIS. Taxa richness increased along with their total density (r = 0.57, p < 0.01, n = 36), with macrophyte density (r = 0.73, p < 0.01, n = 12) and with salinity (r = 0.57, P < 0.01, n = 36) . Total densities of both native species and NIS
were similar between tides and among transects (ANOVA, P > 0.05)
The invertebrates collected in beach seine samples consisted of decapod crustaceans, all of which are native species (Table 2.3) .
The only two NIS of fishes are the American shad (Alosa
sapidissima, native to the east coast of the U.S.) and the cyprinodont rainwater killifish (Lucania parva, native to the east coasts of the U.S. and Mexico) . Both NIS of fishes were from
Yaquina Bay and were at low densities (Table 2.3)
dominant fishes in both estuaries are the northern anchovy (Engraulis mordax), shiner surfperch (Cymatogaster aggregata) and staghorn sculpin (Leptocottus armatus, Tables 2.3, A.3 and A.4) The CPUE and richness of fishes were at least 10 and two times greater than the decapods at every site in both estuaries. The CPUE and richness between estuaries were similar both in the case of fishes and decapods (Figure 2.3, ANOVA, P > 0.20)
Mean summer CPUE of fishes and decapods did not vary with invertebrate densities in sediments (native, NI, or all taxa; r < 0.26, p > 0.20, n = 12)
Species richness of fishes and decapods
Figure 2.2. Surmner mean density of benthic invertebrates in sediment core samples from Alsea Bay and Yaquina Bay(bars) and percent density of nonindigenous to native species (circle) Taxa: native (NA); nonindigenous (NI); cryptogenic (CR) and supraspecific (ST) The number of taxa per site are indicated above bars (from left to right: ST; CR; NA; NI) .
25 ALSEA BAY (Low Tide) ST
- 50 20 -
30 6, 1, 7,
10, 1, 17, 4
2 0 40 z
YAQUINA BAY (High Tide)
9, 2, 21, 6
20 10 10
15, 4,18, 8
15 4, 15, 8
< Downstream SITES Upstream >Figure 2.2
Downstream SITES Upstream >
Table 2.3. Summer catch per unit effort (OPUS) and occurrence (00) of fishes and decapods in the Alsea Bay and Yaquina Bay. Based on beach seine collections at six intertidal sites and three subtidal sites per estuary during summer 1993. Nonindigenous species are indicated by an asterisk.
Fish Order Species Atheriniformes
affinis parva* Clupeiformes Alosa sapidissima* Ciupea paiiasii Engraulis mordax Gasterosteiformes Auiorhynchus fiavidus Gasterosteus aculeatus Syngnathus ieptorhynchus Perciformes Cieveiandia ios Cymatogaster aggregata Hyperprosopon argenteum Lepidogobius iepidus
furcatus Pholis ornata Pholis schuitzi Pleuronectiformes Platichthys stellatus Pleuronectes vetulus Salmoniformes Hypomesus pretiosus Oncorhynchus kisutch
Scorpaeniformes Cottus asper Leptocottus armatus Oligocottus macuiosus Crustacea Order Species Decapoda Cancer magister Cancer productus Crangon franciscorum Hemigrapsus oregonensis Heptacarpus paiudicola Puggetia producta 1
2817.1 (100) -
1.6 16.9 3008.3
(44) (33) (78)
(11) (78) (22)
1272.0 (100) 1.5 1.7 0.2 0.7 3.6 0.3
(22) (22) (11) (22) (44) (22)
0.6 (22) 400.1 (100) 0.3
55.6 2.2 0.1 -
(44) (67) (11) -
20.8 0.5 54.4 1.7 0.2 0.2
(78) (11) (89) (56) (11) (11)
Possibly introduced with oysters or ballast water (Hubbs and
Miller 1965); 2 Intended introduction (Craig and Hacker 1940)
Figure 2.3. Summer mean CPUE of fishes and decapods in seine samples from the Alsea Bay and Yaquina Bay estuaries (bars) and percent of CPUE of nonindigenous species relative to native species (circle) Taxa origin: native (NA); nonindigenous (NI) The number of species per site are indicated above bars. .
ALSEA BAY (Low Tide)
YAQUINA BAY (Low Tide)
NI FISH 1'-
- 0.20 W - 0.10
ALSEA BAY (HighTide)
0.00 0 z w 0.10 0
NADECAPODS 7 6 9
IA 'p 4
.< Downstream SITES Upstream
< Downstream SITES Upstream
in both estuaries and the percent of CPUE of NIS relative to native species in Yaquina Bay exhibited no clear spatial patterns (Figure 2.3) .
Species richness of fishes and decapods increased
with macrophyte density (r
0.51, P < 0.10, n
12) but it was
not correlated with salinity, temperature and transparency (r 0.35, P
0.16, n = 18)
Unlike decapods, both richness and total
CPUE of fishes differed among months (P < 0.05, ANOVA), possibly due to higher fish mobility. However, higher CPUE of fish at lowtide only is suggested in Alsea Bay (P < 0.01, ANOVA).
Most taxa of different origin occurred in different assemblages. Clustering of invertebrates into three groups revealed that only native and NI polychaetes shared an assemblage along with native crustaceans (group 2, Figure 2.4)
were less dominant. Group 1 consisted of an intermediate-density group composed of native bivalves, NI crustaceans and cryptogenic polychaetes. Group 3 included most taxa and had low occurrence and intermediate to low density (Figure 2.4)
Except for upstream
sites, a 4-group level of sites revealed clusters nearly exclusive of each estuary (Figure 2.4)
A 3-group level of similarity among taxa indicated that only native and NI atheriniformes shared a group of low CPUE and very low co-occurrence along with NI clupeiformes (Group 1, Figure 2.5, Table 2.3)
Hence, potential interaction between
noncoevolved fishes may be low. Group 2 included an intermediate CPUE group of native fishes and decapods with high occurrence and group 3 consisted of native fishes with intermediate to high CPUE and occurrence (Figure 2.5) . A 4-group level of sites revealed
Figure 2.4. Clusters by taxa and sites based on invertebrate densities in sediment samples. Species origin within taxa: nonindigenous (*), cryptogenic (**) and native (no asterisk) High- and low-tide sites are represented respectively by upperand lower-case letters (Alsea: A; a, Yaquina: Y; y) followed by site number. Species densities for each taxa are indicated in Tables 2.2; A.1; and A.2.
z 20 Cd)
Taxa I Sites Al A2 al A3 a3 A4 Y5 Y4 y2 Yl Y2 y4 Y3 A5 y6 Y6 A6 a6 BivaJvia r Crustacea* y
Polychaeta**yyyYVvv y Crustacea
PoJychaeta V YVY V Polychaeta* "V'V y Bivajvia*
Vyyyvyyyyyyyyy VVVVVVV V
Gastropoda V V V
Crustacea** V Phoronida
V V V
Density(No/ m2): v: >0-10; V: 11-100; Y:1O1-1,000; V:i3O0i-io,000; V:10,001-21,000 Figure 2.4
Figure 2.5. Clusters by taxa and sites based on CPUE of fishes and decapods in seine samples. Species origin within taxa: nonindigenous (*), native (no asterisk) . High- and low-tide sites are represented respectively by upper- and lower-case letters (IUsea: A; a, Yaquina: Y; y) followed by site number. The CPUE for species within each taxa are indicated in Tables 2.3; A.3; and A.4.
Taxa I Sites
Yl y6 y2 y4 A6 a3 a6 al Al A2 A3 A5 Y2 Y5 Y6 Y3 Y4 A4
Atheriniformes Atheriniformes* Clupeiformes* Decapoda
Gasterosteiformes Clupeiformes Perciformes
V VV V
CPUE (No! 1000 m2): 7: >0-10; Y: 11-100; V:ioi-i3Ooo;
VVVVVVVVVV V:i,00i-io,000; Y':10,OOl-37,OOO
that only group 4 is not exclusively composed by sites from each estuary (Figure 2.5) .
Such contrast between estuaries seem mostly
due to the absence of NI fishes in Alsea Bay and to the virtual absence of native atheriniformes in the latter estuary.
Invertebrate Density and Richness Vs. Temperature-Salinity Native and NI invertebrates in sediment samples occurred under the same temperature-salinity combinations in each estuary when combining data for low- and high-tides (Figure 2.6: A-D). Total Densities of NIS were more prevalent at: 1) high-mid temperatures in both estuaries, 2) mid salinities in Alsea Bay (Figure 2.6: A), and 3) mid-low salinities in Yacjuina Bay (Figure 2.6: C).
Yet, no consistent density patterns between estuaries are apparent for native invertebrates (Figure 2.6: B,D)
Native invertebrates dominated under each temperature-salinity condition in Alsea Bay, both in terms of mean percents of density (Figure 2.7: B) and richness (Figure 2.8: B) .
Despite the higher
percent of richness for native species in Yaquina Bay (Figure 2.8: D), the NIS reached maximum mean percent densities (46% to 68%) at high temperatures and mid-low salinities (Figure 2.7: C)
Species Along Environmental Gradients Environmental influences on densities of invertebrates in sediments and on CPUE of fishes and decapods in intertidal habitats at high-tide are best explained by gradients in salinity, temperature and macrophyte density. Sites with similar environmental characteristics were consistently grouped in both CCA ordinations (Figures 2.9 and 2.10) . The first three CCA
ordination axes for the latter community-environmental associations accounted for nearly half of the total variance for
bothdensities of invertebrates in sediments and for the CPUE of fishes and decapods from seine samples (Table 2.4)
B ALSEA BAY
A ALSEA BAY
17.8 °C) and salinities (> 22.5
The implied response of individual species to environmental factors considered in CCA analyses at high-tide were consistent with the original species densities and their distributions in relation to environmental factors. Dominant patterns for NIS at
high-tide coincided with high temperatures and salinities known to be optimum for the growth of NI amphipods in U.S. west coast estuaries in comparison to those for native amphipods (J.W. Chapman, in progress)
Environmental influences on native species and NIS remain to be investigated for seasons other than summer. Nevertheless,
invertebrate abundance at mid- and high-latitudes is mainly associated with the annual temperature cycle (Nichols and Pamatmat 1988; Day et al. 1989) and seasonal food availability (Day et al. 1989).
The status of invasion biology in estuaries is too incipient to estimate the percent of "harmful" species in terms of ecological or economic impacts (Ruiz et al. 1997) . Nevertheless,
Alsea Bay and Yaquina Bay are susceptible to NIS invasions from highly invaded regional areas (e.g., Coos Bay, San Francisco Bay)
and from more distant donor areas. Further inadvertent or intentional species introductions should be prevented if dominance of native species and their potential ecological functions are to be maintained.
Baltz, D.M. 1991. Introduced fishes in marine systems and inland seas. Biological Conservation 56:151-177. Bayer, R.D. 1981. Shallow-water intertidal ichthyofauna of the Yaquina Estuary, Oregon. Northwest Science 55:182-193. Behrens, S.Y., and C. Hunt. In press. The arrival of the European green crab Carcinus maenas in the Pacific Northwest. Dreissena 11(1).
Bottom, D.L. and K.K. Jones. 1990. Species composition, distribution, and invertebrate prey of fish assemblages in the Columbia River Estuary. Progress in Oceanography 25:243-270. B. Kreag, F. Ratti, C. Roye, and R. Starr. 1979. D.., Habitat classification and inventory methods for the management of Oregon estuaries. Estuary Inventory Report 1. Oregon Department of Fish and Wildlife, Portland, Oregon, USA,
Burt, W.V. and W.B. McAlister. 1959. Recent studies in the Hydrography of Oregon estuaries. Oregon Fish Commission Research Briefs 7:14-27.
Canton, J.T. 1979. History, biogeography and ecology of the introduced marine and estuarine invertebrates of the Pacific Coast of North America. Ph.D. dissertation. University of California, Davis, 904 pp. Carlton, J.T. 1992. Dispersal of living organisms into aquatic environments as mediated by aquaculture and fisheries activities. In: pages 13-46, A. Rosenfield and R. Mann (eds.) Dispersal of living organisms into aquatic ecosystems. A Maryland Sea Grant Publication, College Park Maryland. Carlton, J.T. 1996a. Pattern, process, and prediction in marine invasion ecology. Biological Conservation 78: 97-106. Carlton, J.T. 1996b. Biological invasions and cryptogenic species. Ecology 77:1653-1655. Carlton, J.T. Unpublished data. Maritime Studies Program. Williams College. Mystic Seaport, Mystic, Connecticut.
Canton, J.T. and J.B. Geller. 1993. Ecological roulette: The global transport of nonindigenous marine organisms. Science 261:78-82.
Castillo, G.C. 2000. Benthic biological invasions in two temperate estuaries and their effects on trophic relations of native fish and community stability. Ph.D. thesis. Oregon State University, Corvallis, Oregon.
Castillo, G.C., H.W. Li, J.W. Chapman and T.W. Miller. In press. Predation on native and nonindigenous amphipod crustaceans by a native estuarine-dependent fish. In: J. Pederson (ed.), National Conference on Marine Bioinvasions Proceeding. January 1999. Massachusetts Sea Grant, Cambridge, Massachusetts Institute of Technology, MA. Castillo, G.C., H.W. Li, and P.A. Rossignol. 2000. Absence of overall feedback in a benthic estuarine community: a system potentially buffered from impacts of biological invasions. Estuaries 23:275-291. Chapman, J.W. 1988. Invasions of the Northeast Pacific by Asian and Atlantic gamrnaridean amphipod crustaceans, including a new species of Corophium. Journal of Crustacean Biology 8:364-382. Chapman, J.W. 1997. Personal communication. Hatfield Marine Science Center, Oregon State University, Newport, Oregon 97365.
Chapman, J.W. In progress. Hatfield Marine Science Center. Oregon State University. Newport, Oregon 97365.
Chapman, J.W. and J.T. Canton. 1991. A test of criteria for introduced species: The global invasion by the isopod Synidotea iaevidorsaiis. Journal of Crustacean Biology 11:386400.
Chapman, J.W. and J.T. Canton. 1994. Predicted discoveries of the introduced isopod, Synidotea iaevidorsaiis (Miers, 1881). Journal of Crustacean Biology 14:700-714. Cohen, A.N. and J.T. Canton. 1995. Nonindigenous aquatic species in a United States estuary: A case of the biological invasions of the San Francisco Bay and Delta. Biological Study. A Report for the U.S. Fish and Wildlife Service, Washington, D.C. and the National Sea Grant College Program, Connecticut Sea Grant. Cohen, A.N. and J.T. Canton. 1998. Accelerating invasion rate in a highly invaded estuary. Science 279:555-558. Connor, S.F. and S.D. McCoy. 1979. The statistics and biology of the species-area relationship. The American Naturalist 113:791-833. Cortright, R., J. Weber and R. Bailey. 1987. The Oregon estuary plan book. Oregon Department of Land Conservation and Development, 126 pp.
Cowardin, L.M., V. Carter, F.C. Golet and E.T. LaRoe. 1979. Classification of wetlands and deepwater habitats of the United States. Fish and Wildlife Service. U.S. Department of the Interior. FWS/OBS-79/31. 131 pp.
Craig, J.A. and R.L. Hacker. 1940. The history and development of the fisheries of the Columbia River. Fisheries Bulletin, U.S. 32:133-216. Daehler, C.C. and D.R. Strong. 1996. Status, prediction and prevention of introduced cordgrass Spartina spp. invasions in Pacific estuaries, USA. Biological Conservation 78:51-58. Day, J.W.Jr., C.A.S. Hall and A. Yanez-Arancibia. 1989. Estuarine Ecology. John Wiley & Sons. New York, 558 pp. De Ben, W.A., W.D. Clothier, G.R. Ditsworth and D.J. Baumgartner 1990. Spatio-temporal fluctuations in the distribution and abundance of demersal fish and epibenthic crustaceans in Yaquina Bay, Oregon. Estuaries 13:469-478. Devore, J. and R. Peck. 1986. Statistics. The exploration and analysis of data. West Publishing Company. St. Paul, Minnesota, 699 pp. Elton, C.S. 1958. The ecology of invasions by animals and plants. Reprint 1972. Chapman and Hall, London. 181 pp.
Fauchald, K. and P.A. Jumars. 1979. The diet of worms: a study of polychaeta feeding guilds. Oceanography and Marine Biology: An Annual Review 17:193-284. Grassle, J.F. and J.P. Grassle. 1974. Opportunistic life histories and genetic systems in marine benthic polychaetes. Journal of Marine Research 32:253-284. Grosholz, E.D. and G. Ruiz. 1996. Predicting the impact of introduced marine species: lessons from the multiple invasions of the European Green crab Carcinus maenas. Biological Conservation 78 : 59-66.
Haertel, L.S. and C.L. Osterberg. 1967. Ecology of zooplankton, benthos and fishes in the Columbia River Estuary. Ecology 48:459-472. Hamilton, S.F. 1973. Oregon estuaries. State of Oregon. State Land Board. Division of State Lands, 48 pp. Hubbs, C.L. and R.R. Miller. 1965. Studies of cyprinodont fishes. XXII. Variation in Lucania parva, its establishment in western United States and description of a new species from an interior basin in Coahuila, Mexico. University of Michigan. Miscellaneous Publications Museum of Zoology 127:1-104. Jones, K., C. Simenstad, D. structure, distribution, epibenthos, and plankton Progress in Oceanography
Higley and D. Bottom. 1990. Community and standing stock of benthos, in the Columbia River estuary. 25:211-242.
Jones, M.M. 1991. Marine organisms transported in ballast water. A review of the Australian Scientific Position. Bureau of Rural Resources. Australian Government Publishing Service. Canberra. Bulletin No. 11, 48 pp.
Krygier, E.E., W.C. Johnson and C.E. Bond. 1973. Records of the California tonguefish, threadfin shad and smooth alligatorfish from Yaquina Bay, Oregon. California Fish and Game 59:140-142. Lee, D.S., C.R. Gilbert, C.H. Hocutt, R.E. Jenkins, D.E. McAllister and J.R. Stauffer. 1980. Atlas of North American Freshwater fishes. North Carolina State. Museum of Natural History. Publication 1980-12 of the North Carolina Biological Survey.
Li, H.W. and P.B. Moyle. 1993. Management of introduced fishes. In: pages 287-307, Koheler, C.C. and W.A. Hubert (eds.) Inland fisheries management in North America. American Fisheries Society.
Ludwig, J.A. and J.F. Reynolds. 1988. Statistical ecology. John Wiley & Sons. New York, USA. 337 pp. McCune, B. 1997. Influence of noisy environmental data on canonical correspondence analysis. Ecology 78:2617-2623. McCune, B. and M.J. Mefford. 1997. Multivariate analysis of ecological data. MjM Software, Gleneden Beach, Oregon, USA. Monaco, M.E., T.A. Lowery and R.L. Emmett. 1992. Assemblages of U.S. west coast estuaries based on the distribution of fishes. Journal of Biogeography 19:251-267. Monaco, M.E., D.M. Nelson, R.L. Emmett, and S.A. Hinton. 1990. Distribution and abundance of fishes and invertebrates in West Coast Estuaries, Volume I: Data summaries. NOAA, Rockville, MA. 240 pp. Moyle, P.B. 1985. Patterns in distribution and abundance of a noncoevolved assemblage of estuarine fishes in California. Fishery Bulletin 84:105-117. Moyle, P.B. 1986. Fish introductions into North America: In: H.A. Mooney and J.A. Drake (eds.), Ecology of biological invasions of North America and Hawaii. Ecological Studies 58:27-43. Springer-Verlag. New York. Nichols, F.H. and M.M. Pamatmat. 1988. The ecology of the softbottom benthos of San Francisco Bay: A community profile. U.S. Dept. of the Interior. Fish and Wildlife Service National Wetlands Research Center. Washington, DC. Biological Report 85(7.19).
Pearcy, W.G. and S.S. Myers 1974. Larval fishes of Yaquina Bay, Oregon: a nursery ground for marine fishes? Fishery Bulletin U.S., 72:201-213. Percy, K., C. Sutterlin, D.A. Bella and P.C. Klingeman. 1974. Description and information sources for Oregon estuaries. Sea Grant College Program. Oregon State University, Corvallis, Oregon, 294 pp.
Pearson, T.H. and R. Rosenberg. 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: An Annual Review 16:229-311. Peterson, C.H. 1979. Predation, competitive exclusion, and diversity in the soft-sediment benthic communities of estuaries and lagoons. In: R.J. Livingston (ed.), p 233-264. Ecological Processes in Coastal and Marine Systems. Plenum Press, New York. Posey, M.H. 1988. Community changes associated with the spread of an introduced seagrass, Zostera japonica. Ecology 69:974-983. Robinson, A. 1995. personal communication. Hatfield Marine Science Center. Oregon State University, Newport, Oregon 97365.
Ruiz, G.M., J.T. Canton, E.D. Grosholz and A.H. Hines. 1997. Global invasions of marine and estuarine habitats by nonindigenous species: mechanisms, extent and consequences. American zoologist 37:621-632. Smith, L.D., N.J. Wonham, L.D. McCann, D.M. Reid, J.T. Carlton and G.M. Ruiz. 1996. Shipping Study II. Biological invasions by nonindigenous species in United States waters: Quantifying the role of ballast water and sediments. Parts I and II. Department of Transportation. U.S. Coast Guard. Marine Safety and Environmental Protection. Report No. CG-D-02-97. Washington D.C. Ter Braak, C.J.F. 1986. Canonical correspondence analysis: A new eigenvector technique for multivariate direct gradient analysis. Ecology 67:1167-1179. Walker, J.D. 1973. Effect of bark debris on benthic macrofauna of Yaquina Bay, Oregon. MS. Thesis. Oregon State University. Corvallis, Oregon, 94 pp. Ward, J.H. 1963. Hierarchical grouping to optimize an objective function. Journal of the American Statistical Association 58:236-244.
Williams, R.J., F.B. Griffiths, F.J. Van der Wal and J. Kelly. 1988. Cargo vessel ballast water as a vector for the transport of Non-indigenous marine species. Estuarine, Coastal and Shelf Science 26:409-420.
Whitlatch, R.B. 1980. Patterns of resource utilization and coexistence in marine intertidal deposit-feeding communities. Journal of Marine Research 38:743-765 Woodin, S.A. 1983. Biotic interactions in recent fossil benthic communities. In: pages 3-38, Tevesz, M.J. and P.L. McCall (eds.), Biotic interactions in recent and fossil benthic communities. Plenum Press, New York. Woodin, S.A. and J.B.C. Jackson. 1979. Interphylethic competition among marine benthos. American Zoologist 19:10291043.
Trophic Contribution and Selection of Native and NonindigenouS Prey by Native Fishes in stuarine Rearing-Habitats
H.W. Li2, J.W. Chapman3, T.W. Miller3
Present address: Hatfield Marine Science Center. Oregon State University, Newport, OR 97365.
Oregon Cooperative Fish and Wildlife Research Unit, Department
of Fisheries and Wildlife. Oregon State University Corvallis, OR 97331.
Hatfield Marine Science Center. Oregon State University, Newport, OR 97365.
We determined the summer prey richness, prey composition and selection for native and nonindigenous (NI) prey by native juvenile pleuronectids (English sole: Pleuronectes vetulus and starry flounder: Platichthys stellatus) . Study areas included
intertidal and adjacent subtidal fish rearing habitats in two Northeast Pacific estuaries, the Alsea Bay and Yaquina Bay (Oregon, USA) .
Major NI prey varied greatly among sites with the
polychaete Pseudopolydora kerripi, the clam Mya arenaria and the
cumacean Nippoleucon hinumensis being common in both estuaries. Native species dominated in Alsea Bay both in prey numbers and volumes for both fish species but in Yaquina Bay neither native or NI prey dominated. Fish reliance on native prey seems higher in intertidal areas of Alsea Bay and in subtidal areas of Yaquina Bay. Unlike NI prey, the total volume of consumed native prey increased with the weight of starry flounder in each estuary.
Intertidal CPUE OF fish were not correlated with total numbers of prey in the fish diet. The similar richness ratios of native to NI prey present in the fish diet and the benthos indicate that fish do not distinguish between native and NI prey species.
Predominant selection for native or NI prey types is not apparent by fish in either estuary. Thus, predator-prey coevolution is not a critical determinant of prey selection by these species.
An assumption that interspecific differences in prey selection and resource partitioning result from coevolution of species has
been increasingly challenged in studies of invaded communities (Moyle et al. 1982; Castillo et al. 1995; Castillo et al. In press) .
Nonindigenous species (NIS) are particularly important in
estuaries of the Pacific coast of North America, where at least 234 NIS of invertebrates, fishes, other vertebrates and vascular
plants have been introduced (Carlton 1979; Cohen and Canton
1998; Castillo 2000) . Major mechanisms of NIS introductions in
Northeast Pacific estuaries include propagation of species associated with imported oysters, ballast water release and fouling on hull of ships (Carlton 1979; Cohen and Carlton 1998) Potential impacts of conspicuous NIS in U.S. west coast
estuaries since the late 1980s (e.g., the clam amurensis; the crabs
sinensis and Carcinus maenas) have
caused great concern (e.g., Kimmerer et al. 1994; Grosholz and Ruiz 1995) . However, studies in virtually all these estuaries
lack the resolution to assess long-term changes in food web dynamics resulting from earlier biological invasions. Although the implications of such possible trophic changes are uncertain, changes in prey composition and prey density can affect the growth, survival and year-class strength of fish (Steele et al.
1970; Poxton et al. 1983). Changes in the latter population parameters could be enhanced by further human impacts such as pollution and habitat degradation (e.g., Cross et al. 1985; Sogard 1994)
Anecdotal information from northern San Francisco Bay suggests that native fish rely little on nonindigenous
(NI) prey (Carlton
1979). However, the food webs in the latter system are overwhelmingly dominated by NIS and cryptogenic species (i.e., species of unknown geographic origin, Carlton 1996b; Nichols et al. 1990; Cohen and Canton 1995) . In the Alsea Bay and Yaquina
Bay estuaries, Oregon, the juveniles of four species of native fish: English sole (Pleuronectes vetulus); starry flounder
stellatus); staghorn sculpin (Leptocottus armatus);
and chinook salmon (Oncorhynchus
preyed upon at
least one macrobenthic NIS (Castillo et al. 1995) .
fishes select more native prey in comparison to NI prey, and if the availability of native prey in the environment has declined as a result of biological invasions, then these introductions could be causing long-term declines of native estuarine-dependent fishes.
Assuming that predator-prey coevolution is not a critical determinant of prey usage, we would expect similar selection patterns between native and NI prey types by generalist predators, but the evidence is mixed. Prey selection by larvae of the introduced striped bass
seems higher for
at least one coevolved copepod when compared to two NI copepods (e.g., Meng and Orsi 1991). In contrast, juvenile English sole did not consistently select two native amphipods over two NI amphipods (Castillo et al. In press)
Given the limited understanding on the implications of coevolved and non-coevolved predator-prey interactions in invaded estuaries, we address the following questions: 1) Is the richness of native and NI invertebrates in the fish diet proportional to
their attendant richness in the environment?; 2) What are the contributions of native species, NIS and cryptogenic species to the food-base of native fish in terms of frequency of occurrence, number and volume of prey?; 3) Does the number and volume of prey vary with fish size and weight?; 4) Are daily dietary patterns in number and volume of prey evident?; 5) Is the total number of
prey in the fish diet correlated with benthic densities of invertebrates and with CPUE of fish in intertidal areas?; and 6) Is the overall prey selection by native fish similar on native and NI prey types? To address these questions we surveyed intertidal and adjacent subtidal rearing-habitats of juvenile English sole and starry flounder in the Alsea Bay and Yaquina Bay estuaries (Oregon, USA, Figure 3.1) .
These surveys revealed
higher densities of NI benthic macroinvertebrates in Yaquina Bay (Castillo 2000)
English sole and starry flounder are estuarine-dependent species (e.g., Pearcy and Myers 1974; Monaco et al. 1990). Many age-0 juveniles of these species recruit to shallow estuarine nursery areas where older individuals reach highest densities during summer before migrating to deeper habitats (Orcutt 1950; Krygier and Pearcy 1986; Boehlert and Mundy 1987) . Juveniles of
Figure 3.1. Fish and invertebrate collection sites in Alsea Bay and Yaquina Bay. Indicated are the means and ranges of water temperature (A) and salinity (S) observed at high tide during summer 1993 surveys. Alsea sites 2; 4 and 5 and Yaquina sites 1; 3 and 5 were sampled at low- and high-tide in the first survey and at high-tide thereafter. Sampled month/day: Alsea Bay (7/7; 7/19; 8/2 and 9/16); Yaquina Bay (7/5; 7/18; 8/3 and 9/17).
these two species prey on epifauna and infauna (Orcutt 1950; Haertel and Osterberg 1967; Collins 1978; Toole 1980) Yaquina Bay is 192 km south of the Columbia River estuary. Nearly 35% of its 15.8 km2 surface at mean high-water is intertidal (Hamilton 1973) . Alsea Bay is 25 kin south of Yaquina Bay (Figure 3.1) . About 46% of the 8.7 km2 surface of the Alsea
Bay at mean high-water is intertidal (Hamilton 1973) .
Yaquina Bay has received ballast water traffic. Both estuaries have been used for culture of introduced oysters (Carlton 1979) Unlike Yaquina Bay, oyster culture was discontinued in Alsea Bay in the 1930s (A. Robinson, personal communication 1995)
Field Sampling We conducted four surveys of six intertidal areas per estuary during daylight hours in summer 1993. Intertidal sites were in salinity areas ranging from about 34°/ (lower estuary) to 5°/
closer to the ocean
in upstream areas (upper estuary, Figure
3.1). A beach seine (32 mx 1.8 m and 0.8 cm stretched mesh size) was used to collect juvenile English sole and starry flounder. Seine was also used to collect fish from three to six subtidal areas adjacent to intertidal sites at low-tide (Figure 3.1) Seining area encompassed a semicircle of 163 m2 from the water line (0 in depth) to deeper areas. The catch per unit effort
(CPUE) of each flatfish species caught with seine in a given site was standardized to individuals per 1000 m2.
The 370 collected fish (Table 3.1) were immersed in a lethal doses of MS-222 (200 mg/l)
A 10% solution of buffered forinalin
was then injected into the coelomic cavity to fix prey items followed by fish preservation in 80% ethanol. Feeding habits of English sole are based on stomach contents. All prey items in the stomach and the anterior 1/3 of the intestine were analyzed in
the starry flounder since stomach fullness in this species was often low. We used Hogue and Carey's (1982) index of stomach fullness (0: < 5% full; 1: 5-25% full; 2: 25-50% full; 3: 50-75%
full; 4: 75-100% full) to account for differences in prey volume independent of fish size. Total volume of each prey item per fish was measured in a graduated centrifuge tube (1-600
the small size of most prey, this method was more practical than determining prey weight or displacement volume.
Availability of macroinvertebrates within each intertidal site is estimated from the mean number of prey collected over three parallel intertidal transects located at 0, 40 and 80 cm of water depths at the time of fish collections. Each transect is parallel to the water line, 30 m long and included 10 equally spaced core sediment samples. Each sediment core is 3.2 cm in diameter and 13 cm deep. We composited core samples from each transect and washed on a 0.5 mm mesh sieve. All invertebrates retained in the sieve were preserved in 10% buffered forinalin and later transferred to 70% ethanol.
Individual prey items were identified to species whenever
possible. Classification of species as native, NI, and cryptogenic is based either on reported species introductions
(Carlton 1979; Canton and Geller 1993; Cohen and Canton 1995) or by using criteria for detecting NIS (Carlton 1979; Chapman 1988). Taxa not resolved to species level are referred to as supraspecific taxa and they may include both native species and NIS.
Dietary analysis is restricted to fishes captured during two time periods: low-tide morning hours and high-tide afternoon hours to account for daily rhythms of feeding. Although prey items were found in the gut of 90.5% of all fish analyzed,
comparisons between total prey volume and fish size and weight are limited to fish collected during afternoon hours since prey volume in fish guts usually increased throughout the day. To estimate prey selection we determined prey availability from benthic cores collected at sites where fish were simultaneously sampled.
Computations The frequency of occurrence of prey species k in the fish diet (Ok)
is defined as:
is the number of fish containing prey k and n is the
total number of fish. We computed the mean prey frequency of occurrence for all species of a particular origin
native, NI and cryptogenic) in the fish diet (MO) using the formula:
MO = Where °kj is the corresponding °k value for prey species , of origin.3 and m is the total number of prey of origin. (MN)
We computed the mean number of all prey species of origin in the fish diet using the formula:
Mn = n_li
Where NkI is the number of prey species k of origin for a total
prey species of origin
Likewise, the mean prey volume of all species of origin
the fish diet is computed by substituting Nkl by Vkj! in the previous formula. We defined the overall relevance of prey item k to the fish diet by the index of overall item contribution (OIC)
which we defined by the percentages of prey frequency of occurrence (Ok), volume (%Vk) and number (%Nk)
itemk. The OIC index can range between 0% in all fish examined) to 100%
(i.e., itemk is absent
(i.e., only itemk occurs in all
fish examined) . We used the OIC index to rank up to 15 dominant
items in the fish diet. Estuaries are divided into three sections to determine the OIC values in fish from downstream to upstream sites: lower estuary (sites 1-2), mid estuary (sites 3-4), and upper estuary (sites 5-6)
Bivalves and their siphons are considered different items since siphon cropping by fish is more predominant than consumption of entire bivalves. Polychaete parts of unknown species, plant matter and pieces of woody debris are also included as prey items. Unidentifiable organic matter comprised less than 4% of the mean percent volume and is not included in the OIC index.
We estimate prey selection of macroinvertebrates in the fish diet with Johnson's (1980) selection index(E1): = n1
where rjj is the rank of usage of prey item
the abundance of prey1 in the fish diet), availability of item
is the rank of
(based on the benthic density of
item1) and n is the total number of fish. The more used and/or available an items is, the closer to one is its average rank. Hence, the most selected item has the lowest
macrobenthic prey occurring in at least 5% of the fish are included in the computations of prey selection. Unlike other selection indices, the exclusion of certain items from the analyses based on the Johnson's (1980) method do not alter the conclusions for the items considered. Selection estimates are based on the program PREFER v5.l, Pankratz, 1994) and computations of ranks for usage and availability are based on Quatro Pro 6.1.
Differences in mean prey occurrence, abundance and volume among native species, NIS and cryptogenic species are compared with t-tests for unequal variances using Statgraphics 2.1 Plus. Differences in the proportions of native and NI prey between low and high-tide were compared by X2 tests.
Table 3.1. Total number, mean total length and total weight of juvenile English sole and starry flounder. Fish were collected in Alsea Bay and Yaquina Bay during summer 1993 (SE = standard error)
Species Estuary English sole: Alsea Bay Yaquina Bay Starry flounder: Alsea Bay Yaquina Bay
Length (cm) Mean SE
Number of fish
Table 3.2. Species richness and frequency of occurrence of native and nonindigenous (NI) invertebrates in the environment and the diet of English sole and starry flounder. Mean frequency of prey occurrence in the fish diet is indicated in parenthesis. No significant differences are detected in the proportion of native to NI invertebrate species between the fish diet and the environment (P > 0.10; Fisher's exact test) and between the mean frequency occurrence of native and NI prey in the fish diet from each estuary (P > 0.05; t-test)
No. Species Environment
No. Species (Occurrence) Fish Diet
Alsea Bay English sole Starry flounder Both species
31 30 34
Yaquina Bay English sole Starry flounder Both species
49 50 50
11 11 11
9.0) (12.8) (10.2) (
Is the proportion of native and NI invertebrate species similar between the fish diet and the environment?
Yes, their proportional richness between the diets of each fish species and the benthic environment was similar in each estuary. Most prey species are native in both estuaries (Table 3.2) .
Except for fish consumed by starry flounder, all major prey
taxa included at least one NIS introduced from the Atlantic coast and/or the Western Pacific (Table 3.3) . Moreover, The proportion
of NIS to native species in the diet of both fish in Yaquina Bay (9/25) was similar to that of Alsea Bay (5/20; X2 = 0.15; P > 0.70)
What are the contributions of native species, NIS and cryptogenic species to the food-base of native fish in terms of frequency of prey occurrence and number and volume of prey? Similar mean occurrences of native and NI prey were evident in the diet of each fish species in each estuary (Table 3.2) .
no case did cryptogenic prey exceed mean occurrences of native and NI prey and occurrence of cryptogenic prey was significantly lower than native and NI prey in the diet of starry flounder from Yaquina Bay (P -
50 z w
Starry Flounder (Alsea)
Starry Flounder (Yaquina)
No. Fish = 61
No. Fish = 61
English sole (Alsea)
English sole (Yaquina)
Starry flounder (Alsea)
Starry flounder (Yaquina)
Figure 3.3. Percent volume of major prey by origin in English sole and starry flounder diets. Prey are classified as native (NA); nonindigenous (NI); cryptogenic (CR) and supraspecific taxa (ST) . Harpacticoids are included within crustaceans as ST.
A English Sole (Alsea Bay)
English Sole (Yaquina Bay)
NA (22.7) NA (16.7)
CR )0.9k, ST (0.2)
ST (2.6) CR (5.0) NA (26.0)
C Starry Flounder (Alsea Bay)
D Starry Flounder (Yaquina Bay) 1CR+ST (0.1)