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 aboveaverage 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(Pleuronectes veLulusandPlatichthys 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 human- mediated biological invasions should be avoided. Benthic Biological Invasions in Two Temperate Estuaries and Their Effects on Trophic Relations of Native Fish and Community Stability
by
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 presentedon July 14, 2000
APPROVED:
Redacted for Privacy
Major Professor, representing Fisheries Science
Redacted for Privacy
Chair of Department ofisheries and Wildlife
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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 toany reader upon request.
Redacted for Privacy
hnzalo C. Castillo, Author ACKNOWLEDGMENT
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 1
Significance of Biological Invasions 1
Mechanisms of Biological Invasions 2
Community Susceptibility to Biological Invasions 3
Extent of Estuarine and Marine Invasions 4
Potential Impacts of Nonindigenous Species 5
Invasions in U.S. West Coast Estuaries 5
Focus of the Present Thesis 7
Study Areas 8
Chapter Outline 9
References 10
Chapter 2: Distribution and Habitat Use by Noncoevolved Assemblages of Macroinvertebrates and Fishes in Two Temperate Estuaries 17
Abstract 18
Introduction 19
Methods 23
Results 28
Discussion 51
References 58
Chapter 3: Trophic Contribution and Selection of Native and Nonindigenous Prey by Native Fishes in Estuarine Rearing-Habitats 64
Abstract 65
Introduction 65
Methods 70
Results 75 TABLE OF CONTENTS (Continued)
Page
Discussion 101
References 108
Chapter 4: Predation on Native and Nonindigenous Aiuphipod Crustaceans by a Native Estuarine-Dependent Fish 113
Abstract 114
Introduction 114
Methods 116
Results 119
Discussion 131
References 135
Chapter 5: Absence of Overall Feedback in a Benthic Estuarine Community: A System Potentially Buffered from Impacts of Biological Invasions 138
Abstract 139
Introduction 140
Methods 143
Results 155
Discussion 166
References 172
Chapter 6: Conclusions 180
Summary 180
Recommendations for Future Research 185
Bibliography 186
Appendices 208
Appendix A: Complement of Chapter 2 209
Appendix B: Complement of Chapter 3 222
Appendix C: Complement of Chapter 4 243
Appendix D: Complement of Chapter 5 248 LIST OF FIGURES
Figure Page
2.1 Alsea Bay and Yaquina Bay estuaries 21
2.2 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) 32
2.3 Summer mean CPUS 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) 35
2.4 Clusters by taxa and sites based on invertebrate densities in sediment samples 38
2.5 Clusters by taxa and sites based on CPUS of fishes and decapods in seine samples 40
2.6 Mean summer densities of assemblages of native and nonindigenous invertebrates in sediment samples under various temperature-salinity combinations 43
2.7 Mean percent densities for summer assemblages of native and nonindigenous invertebrates under various temperature-salinity combinations 44
2.8 Mean percent richness for summer assemblages of native and nonindigenous invertebrates under various temperature-salinity combinations 45
2.9 Ordination of 35 benthic invertebrates from sediment samples and 12 high-tide intertidal sites along environmental gradients 46
2.10 Ordination of 12 fishes, three decapods (code in parenthesis) and 12 high-tide intertidal sites along environmental gradients 48
3.1 Fish and invertebrate collection sites in Alsea Bay and Yaquina Bay 68
3.2 Mean number and volume of prey species in the diets of English sole and starry flounder 78
3.3 Percent volume of major prey by origin in English sole and starry flounder diets 80
3.4 Percentages of prey frequency of occurrence, number and volume of main dietary items of English sole by estuary section 83 LIST OF FIGURES (Continued)
Figure Page
3.5 Percentages of prey frequency of occurrence, number and volume of main dietary items of starry flounder by estuary section 85
3.6 Total prey volume of native and nonindigenous species and all taxa combined as a function of fish weight in intertidal areas 90
3.7 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 92
3.8 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) 94
3.9 Johnson's selection index (line) and ranks of prey usage and availability (bars) for flatfish in Alsea Bay 97
3.10 Johnson's selection index (line) and ranks of prey usage and availability (bars) for flatfish in Yaquina Bay 99
4.1 Amphipod body length (from telson to eye, Y) with length of 4th articLe 2nd antenna (X) by species and sex 120
4.2 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 123
4.3 Mean number of Corophium consumed by Pieuronectes vetulus, in single-species experiments 125
4.4 Strauss' selection index by prey size (4th article 2nd antenna) and Corophium species consumed by Pleuronectes vetuius 126
4.5 Percent of eaten and urieaten Corophium by size (4th article 2nd antenna) in 10 tanks with sand substratum 127
4.6 Mean number of Corophium consumed by Pieuronectes vetulus in mixed-species experiments 129
4.7 Percent of uneaten Corophiumby size (4th article 2nd antenna) in sand and mud substrata 130 LIST OF FIGURES (Continued)
Figure Page
5.1 Ecological interactions between guilds 1, 2 and 3 in numbered circles and attendant community matrices 147
5.2 Basic guild structure of activity models for the benthic community of the Yaquina Bay estuary 156
5.3 Basic guild structure for trophic models of the benthic community in the Yaquina Bay estuary 158
5.4 Distribution of feedback in models for the pre-invaded and the invaded benthic community of Yaquina Bay 164
5.5 Percent of models with near-zero feedback (-2 F 2) 165 LIST OF TABLES
Table Page
2.1 Substrata, vegetation types and macrophyte density of intertidal sites in the Alsea Bay and Yaquina Bay during summer 1993 27
2.2 Summer density and occurrence (OC) of intertidal benthic invertebrates in core samples from Alsea Bay and Yaquina Bay 29
2.3 Summer catch per unit effort (CPUE) and occurrence (OC) of fishes and decapods in the Alsea Bay and Yaquina Bay 34
2.4 Percent of community variance explained, eigenvalues and correlations for the three main axes of CCA ordinations 50
3.1 Total number, mean total length and total weight of juvenile English sole and starry flounder 74
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 74
3.3 Frequency of species occurrence in the diet of juvenile English sole (E) and starry flounder (S) in Alsea Bay and Yaquina Bay 76
3.4 Percent of number of prey by taxa and species origin for two size classes of juvenile English sole in Alsea Bay and Yaquina Bay 87
3.5 Percent of number of prey by taxa and species origin for three size classes of juvenile starry flounder in Alsea Bay and Yaquina Bay 88
4.1 Walking (WA), swimming (SW) and partially visible (PV) Corophium spp. in sand substrate 122
4.2 Walking (WA), swimming (SW) and partially visible (PV) Corophium spp. in sand and mud substrata 128
5.1 Activity and trophic invertebrate guilds assigned to qualitative models of Yaquina Bay 145 LIST OF TABLES (Continued)
Table Page
5.2 Guild structure of activity models and assumptions; number of guilds; number of alternative models and invasion status of the community for each community structure 150
5.3 Guild structure of trophic guild models and assumptions; number of guilds and invasion status of each community structure 151
5.4 Alternative activity guild models for the intertidal benthic community of Yaquina Bay 160
5.5 Alternative trophic guild models for the intertidal benthic community of Yaquina Bay 161 LIST OF APPENDIX FIGURES
Figure Page
B.1 Total weight (W) and total length (L) of English sole and starry flounder 224
B.2 Fulton's condition factor of English sole and starry flounder 228
B.3 Percent of dietary overlap (DO±) and trophic breadth (B1 )for flatfish 231
B.4 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 234
0.1 mphipod dry weight (Y) with length of 4th article 2nd antenna (X) by species and sex 244
0.2 Mean number of survivingCorophiumin single-species predation treatments and in controls without fish .. .. 245
C.3 Mean number of survivingCorophiurnin mixed-species predation experiments in sand and mud treatments and in controls without fish 246 LIST OF APPENDIX TABLES
Table Page
A.1 Summer mean density and overall occurrence of intertidal invertebrates in sediment samples from Alsea Bay 210
A.2 Summer mean density and overall occurrence of intertidal invertebrates in sediment samples from Yaquina Bay 214
A.3 Summer mean catch per unit effort (CPUE) and occurrence of fishes and decapods in Alsea Bay as determined from seine sampling 219
A.4 Summer mean catch per unit effort (CPUE) and occurrence of fishes and decapods in Yaquina Bay as determined from seine sampling 220
A.5 Life-mode and functional-groups of nonindigenous invertebrates found in intertidal and subtidal areas of Alsea and Yaquina Bay 221
B.1 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 226
B.2 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 .. . 227
B.3 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 236
B.4 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 240
0.1 Density and number ofCorophium salmonisin the benthos and the diet of juvenile English sole collected in intertidal areas at high tide 247
D.1 Number of prey and their percent frequency of occurrence in stomachs of juvenile staghorn sculpin(Leptocottus armatus) 249
D.2 Activity models derived from models in Figure 5.2 .. . 250
D.3 Trophic models derived from models in Figure 5.3 . . . . 251 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 2
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 human- mediated 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) 3
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) .Both 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) . Many 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 4 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); 5
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 6 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 sixsian 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 spatio- temporal use of the Squamish estuary, British Columbia, to the production of Eogammarus confervicolus (Levings 1980) .Chum 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 7 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:
Chapter 2: 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.
Chapter 3: 1) 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 8 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.
Chapter 5: 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 functional- group interactions models in the benthic community of Yaquina Bay.
Study Areas
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.9 kmn 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 9 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 (Canton 1979)
Chapter Outline
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(Platichthysstellatus) are considered in field analyses (Chapter 3) .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) 10
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Chapter 2
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. 18
Abstract
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, ship- traffic, 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. 19
Introduction
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) . The 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) . The 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), non- competitive 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) . However, the impacts of NIS invasions on these rearing areas is virtually 20
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) . One approach 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 temperature- salinity combinations?; and 5) Are native species and NIS similarly distributed acrossenvironmentalgradients from 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 21
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. Figure 2.1 23 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) . Yaquina Bay 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)
Methods
Sampling
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) . In each 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 24
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 25 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, y,z axes) against
representative midpoints of salinities (5: 1-9 °/, , 15: 10-19 25: 20-290/ < 35: 30-34.9 0/) and temperatures (15: 13-
16 °C,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. 26
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) .Further variables 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.
Estuary Upstream Substrata Aquatic vegetation Macrophyte Density Site Km Algae Eelgrass
Alsea Bay
Al 4.5 Sand/Polychaete tubes C C 3
A2 4.9 Sand/Mud/Clay C A 4
A3 5.3 Sand/Mud A C 4
A4 7.6 Sand/Mud/Clay C C 3
A5 8.6 Sand/Mud R R 2
A6 10.6 Mud/Clay R N 1
Yaquina Bay
Yl 4.3 Sand A C 4
Y2 9.2 Sand/Mud A A 5
Y3 12.2 Sand/Mud/Cobble A A 5
Y4 14.9 Sand/Cobble/Clay C A 4
Y5 18.6 Mud C C 3
Y6 23.3 Mud/Woody debris R N I 28 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)
Results
Intertidal Habitats
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 < 0.05, n = 27) .Water transparency and salinity are positively associated in both estuaries (Alsea Bay: r = 0.51, P < 0.01, n = 27; Yaquina Bay: r = 0.87, P < 0.001, n = 27). Water transparency at high-tide was higher in Alsea Bay (mean = 1.75 m) than in Yaquina Bay (mean = 1.23 m, P < 0.001, n = 18), and such difference may be due to greater tidal exchange in Alsea Bay.
Species Densities and Richness
Core Sediment Samples
More native species and NIS are found in Yaquina Bay (47 native, 11 NI,8 cryptogenic) than in Alsea Bay (33 native,8 NI, 5 cryptogenic, Table 2.2), and the total number of species and supraspecific taxa was 102 in Yaquina Bay and 66 in Alsea Bay (Tables A.1 and A.2) .The NIS found only in Yaquina Bay are the polychaetes Streblospio benedicti andPseudopolydora paucibranchiataand the amphipod Corophium acherusicum. All NIS in Alsea Bay also occurred in Yaquina Bay (Table 2.2) .The 29
Table 2.2. SurmlIer density and occurrence (DC) of intertidal benthic invertebrates in core samples from Alsea Bay and Yaquina Bay. Possible species origin: native (N), nonindigenous (A = Atlantic, J = Japan,I = Asia) and cryptogenic (C) .Likely vector of introduction: ballast water (B), ship fouling (F), oyster culture (0) .Densities are the mean of six intertidal sites and three subtidal sites (Figure 2.1).
T axa Origin/VectorAlsea Bay Yaquina Bay No.m2 (OC) No.m2 (OC)
Annelida Polychaeta: Abarenicolapacifica N 8 (33) 1 (11) Arnaenaoccidentalis N - - 19 (11) Armandiabrevis N 2 (11) 3 (22) Boccardia proboscidea N 3 (11) 404 (33) Dorvillea annulata N - - 4 (22) Eteonecalifornica C - - 1 (11) Eteonecolumbiensis N 2 (11) Eteonespilotus N 77 (78) 84 (89) Eupolymnia heterobranchia N - 3 (11) Exogonelourei C - 2 (11) Glycera americana N - - 2 (11) Glycindearmigera N 1 (11) - Glycindepolygnatha N 29 (44) 18 (44) Heteromastus filiformis (A/B4O)"3 2 (22) 55 (67) Hobsoniaflorida (A/O?)2 758 (100) 1684 (67) Leithoscoloplos pugettensis N 1 (11) 12 (44) Lumbrineriszonata N - - 6 (11) Magelona hobsonae N - - 3 (11) Manayunkiaaestuarina N 4 (22) 724 (56) Mediomastuscaliforniensis N 137 (56) 426 (89) Nephtyscaeca N 2 (11) 3 (11) Nereislimnicola N 401 (78) 151 (67) Owenia fusiformis C - - 7 (11) Paraonellaplatybranchia N 1 (11) 26 (33) Phyllodocehartmanae N - - 5 (11) Platynereis bicanaliculata N - - 3 (22) Polydora cornuta (A/O,B,F)3 8 (22) 15 (44) Pseudopolydorakempi (I/O,B,F)3 254 (67) 388 (78) Pseudopolydora paucibranchia La (J?/O,B,F)3 - - 2 (11) Pygospio californica N - - 2 (11) Pygospio elegans C 3050 (67) 315 (78) Sphaerosyllis californiensis N - - 3 (11) Streblospiobenedicti (A/B, F, O)" - 987 (89) Crustacea Arnphipoda: Allorchestes angusta N 2 (11) 99 (11) Arnpithoelacertosa N 6 (22) 15 (22) Ampithoe valida (A/B, F, 0) 1,3 5 (22) 66 (33) Corophium acherusicum (A/F, 0)13 - - 49 (44) Corophium brevis N - - 288 (11) Corophium salmonis N 5917 (100) 4256 (100) Corophium spinicorne N 99 (67) 490 (100) Eobrolgusspines us (A/O)2 3 (11) 129 (33) Eogammarus con fervicolus N 52 (89) 68 (67) Eohaustoriusestuarius N 7 (33) 81 (33) Traskorchestia traskiana N - - 3 (11) 30
Table 2.2 Continued.
Taxa Origin/Cause Alsea BayYaquina Bay No.m2 (OC) No.m2 (OC)
Bra chyura: Cancer magister N - - 2 (11) Hemigrapsus oregonensis N 6 (44) 4 (22) Copepoda: Hemicyclops subadhaerens C 26 (44) 23 (67) Cumacea: Nippoleucon hinumensis (J/B)4 57 (78) 335 (100) Cumella vulgaris N 58 (67) 31 (56) Isopoda: Gnorisrnosphaeroma insulare N 3 (22) 5 (33) Gnorismosphaeroma oregonensis N - - 3 (33) Lironeca californica N - - 1 (11) Macrura: Neotrypaea californiensis N 35 (44) 34 (67) Upogebia pugettensis N 2 (22) 1 (11) Tanaidacea: Leptochelia dubia C 105 (22) 1507 (44) Sinelobus stanfordi C 1 (11) 15 (33) Mollusca Bivalvia: Clinocardium nuttalli N 2 (11) 15 (33) Cryptomya californica N 55 (44) 61 (33) Macoma baithica N 185 (100) 268(100) Macorna inquinata N 6 (11) - - Mya arenaria (A/O)1 28 (67) 78 (67) Myseila tumida N 4 (11) 27 (11) Mytilus californianus N 2 (11) - - Mytilus edulis C 3 (11) 36 (33) Transenelia tantilla N - - 28 (11) Gastropoda: Alderia modesta N 5 (11) 18 (33) Apiysiopsis enteromorphae N 6 (22) 8 (33) Littorina sitkana N - - 2 (11) Melanochiamys diomedea N - - 7 (22) Phoronida Phoronis pailida N 269 (56) 140 (44)
Sources:1 Canton(1979); 2 J.T. Canton (personal communication 1996); Cohen and Canton(1995); Canton and Geller(1993) 31 decreasing order of the three dominant species in Alsea Bay is: the native amphipod C. salmonis, the cryptogenic polychaete Pygospio elegans and the NI polychaeteHobsoniaflorida and in
Yaquina Bay: C. salrnonis, H. florida and the cryptogenic tanaid Leptocheliadubia(Table 2.2) Both the density and the percent density of NI to native species were higher in Yaquina Bay (Figure 2.2, p < 0.01, ANOVA) However, densities of both native species and all taxa are similar between estuaries (Figure 2.2,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)
Seine Samples
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) . The three 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 32
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 60 25 ALSEA BAY (Low Tide) ST ___ CR - 50 20 - r NA 20 - NI U) Ui -40 40 15 15 N Ui 30 30 6, 1, 7, 10 - 20 20 Cl) 10, 1, 17, 4 U) - 10 100 0 Ui I- 0 0 oS2 I 2 3 4 5 6 I 2 3 4 5 6 >- 40 5040 50 YAQUINA BAY (High Tide) 2 17,5, 26,8 ST 0 40 CR 30 NA 40 z NI >- z 30 30! 9, 2, 21, 6 U) 20 20 15, 4,18, 8 z 20 ')flUi 15 4, 15, 8 8,1,7,2 10 10 - 10 10
0 0 1 2 3 4 5 6 1 2 3 4 5 6 < Downstream SITES Upstream >- zZ Downstream SITES Upstream >
Figure 2.2 34
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.
Alsea Bay Yaquina Bay Taxa OPUE (00) CPUS (00) (No103 m2) (No103 m2)
Fish Order Species Atheriniformes Atherinopsaffinis 0.2 (11) 15.2 (78) Lucaniaparva* 1 - 0.2 (11) Clupeiformes Alosasapidissima* 2 - 1.6 (44) Ciupeapaiiasii 0.1 (11) 16.9 (33) Engraulismordax 4016.7 (67) 3008.3 (78) Gasterosteiformes Auiorhynchusfiavidus - - 0.2 (11) Gasterosteus aculeatus 102.4 (67) 18.0 (78) Syngnathus ieptorhynchus - - 1.1 (22) Perciformes Cieveiandia ios 0.5 (11) Cymatogaster aggregata 2817.1 (100) 1272.0 (100) Hyperprosopon argenteum - - 1.5 (22) Lepidogobius iepidus - - 1.7 (22) Lumpenussagitta - - 0.2 (11) Phanerodonfurcatus - - 0.7 (22) Pholisornata 3.6 (67) 3.6 (44) Pholisschuitzi - - 0.3 (22) Pleuronectiformes Platichthysstellatus 9.5 (67) 14.7 (89) Pleuronectes vetulus 26.8 (56) 12.2 (67) Salmoniformes Hypomesus pretiosus 2.5 (56) 0.6 (33) Oncorhynchus kisutch 0.2 (11) - - Oncorhynchus tshawytscha 65.6 (100) 12.3 (89) Scorpaeniformes Cottus asper 0.6 (22) - - Leptocottus armatus 400.1 (100) 636.0 (100) Oligocottus macuiosus 0.3 (22) Crustacea Order Species Decapoda Cancer magister 4.1 (11) 20.8 (78) Cancer productus - - 0.5 (11) Crangonfranciscorum 55.6 (44) 54.4 (89) Hemigrapsusoregonensis 2.2 (67) 1.7 (56) Heptacarpus paiudicola 0.1 (11) 0.2 (11) Puggetia producta - - 0.2 (11)
1 Possibly introduced with oysters or ballast water (Hubbs and Miller 1965);2 Intended introduction(Craig and Hacker 1940) 35
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. 0.50 ALSEA BAY (Low Tide) YAQUINA BAY (Low Tide) NA FISH NA DECAPODS NI FISH 0 -0.40 1'- NA FISH 8 NA DECAPODS 11 -0.30
- 0.20 W0
C/) - 0.10 C/) I0.00 0 I 2 3 4 5 6 I 2 3 z 4 5 6 w ALSEA BAY (HighTide) 0.10 0 NA FISH a NADECAPODS 0.08
7 z 6 9 z
0.04 a- 0.02
IiIA'pIi 0.00 I 2 3 4 5 6 1 2 3 4 5 6 .< Downstream SITES Upstream < Downstream SITES Upstream>
Figure 2.3 37 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 low- tide only is suggested in Alsea Bay (P<0.01, ANOVA).
Taxa Assemblages
Sediment Samples
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) . Other groups 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)
Seine Samples
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 38
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 upper- and 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. 40 w z0 20
Cd) DISTANCE
0E - 40 20 0 ,-, I I I Taxa I Sites Al A2al A3 a3 A4 Y5 Y4 y2 Yl Y2y4 Y3A5y6Y6 A6 a6 BivaJvia Crustacea* y r y y Polychaeta**yyyYVvv yyyyyy Crustacea PoJychaeta V YVY VVVyyyyyyyyy Vv Polychaeta* "V'V VVyyyvyyyyyyyyy Bivajvia* y V V V V V V V V Gastropoda V V V V V V V V Bivalvia** V V Crustacea** V V V V V V v V Phoronida v V V V V 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 40
Figure 2.5. Clusters by taxa and sites based on CPUE offishes and decapods in seine samples. Species origin within taxa: nonindigenous (*), native (no asterisk) .High- and low-tide sites are represented respectively by upper- andlower-case letters (IUsea: A;a, Yaquina: Y; y) followed by sitenumber. The CPUE for species within each taxa are indicated in Tables 2.3;A.3; and A.4. w30 zC-) 15 r C,) DISTANCE 0 30 15 0 nl1L1I I Taxa I Sites Yl y6 y2 y4 A6 a3 a6 al Al A2 A3 A5 Y2 Y5 Y6 Y3 Y4 A4
Atheriniformes VV ' V v V Atheriniformes* 7 Clupeiformes* 7 7 7 Decapoda VVVYv V V VVV YVY Pleuronectiformes VV V V V V V V V V V V V V Salmoniformes VVVYVVVVVVVYVYV VY GasterosteiformesV VVV V V VV VV Clupeiformes V V V VV V V Perciformes yyVVyyVVVyyyyyyyy Scorpaeniformes V V V V V V V V V V V V V V V V
CPUE (No! 1000 m2):7:>0-10; Y: 11-100; V:ioi-i3Ooo;V:i,00i-io,000; Y':10,OOl-37,OOO
Figure 2.5 42 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) 43
A ALSEA BAY B ALSEA BAY
25
2O 2O
15 15 . . I -
Z . . 0 10 0 . / -.-.------24 u 5 21
18 <35 25 l5 15 5 4LINfl_r 124(' (°'oo)
C YAQUINA BAY D YAQUINA BAY
Figure 2.6. Mean summer densities of assemblages of native and nonindigenous invertebrates in sediment samples under various temperature-salinity combinations. Low- and high-tide sediment samples are included in intertidal areas of Alsea Bay and Yaquina Bay. 44
A ALSEA BAY B ALSEA BAY
100f 100
U) 80 d. 80
60 60 I 40./Lj 40 12 20i 15 X5, 0 S(,N/ 12' /Ty(%:24k
C YAQUINA BAY D YAQUINA BAY
100
cL
U, 80 z0 LU (0 60 z z 0 Z 40
(U 20 0
25-
12
Figure 2.7. Mean percent densities forsummer assemblages of native and nonindigenous invertebrates undervarious temperature- salinity combinations. Low- and high-tide sedimentsamples are included in intertidal areas of Alsea Bay and YaguinaBay. 45
A ALSEA BAY B ALSEA BAY
100
U)
z 6 z z0 U) U) 24 20 21
18
<35 - 15
12 A'
C YAQUINA BAY D YAQUINA BAY
Figure 2.8. Mean percent richness for summer assemblages of native and nonindigenous invertebrates under various temperature-salinity combinations. Low- and high-tide sediment samples are included in intertidal areas of Alsea Bay and Yaquina Bay. 46
Figure 2.9. Ordination of 35 benthic invertebrates from sediment samples and 12 high-tide intertidal sites along environmental gradients. Indicated are vectors of temperature, salinity and macrophyte density for the two main CCA axes and attendant intraset correlations. Sites correspond to those of Figure 2.1 and Table 2.1. Species origin after species code: nonindigenous (*); cryptogenic (**); native (no asterisk). Species code: Abre = Armandiabrevis; Aent = Aplysiopsis enteromorphae; Alac = Ampithoe lacertosa; Amod = Alderia modesta; Aval* = Ampithoe valida; Bpro = Boccardia proboscidea; Ncal = Neotrypaea californiensis;Cryc = Cryptomya californica; Csal = Corophium saimonis; Cspi = Corophiumspinicorne;Cvul = Cumella vulgaris; Econ =Eogammarus con fervicolus; Eest = Eohaustorius estuarius; Espi = Eteone spilotus; Gins = Gnorismosphaeroma insulare; Gpol = Glycinde polygnatha;Hfil* = Heteromastus filiformis; Hflo* = Hobsoniaflorida; Hore = Hemigrapsus oregonensis; Hsub** = Hemicyclopssubadhaerens; Ldub** = Leptochelia dubia;Lpug = Leitoscoloplos pugettensis; Maest = Manayunkia aestuarina; Mare* = Mya arenaria; Mbal = Macomabalthica;Mcal = Mediomastus californiensis; Nhin* =Nippoleucon hinumensis; Nlim = Nereis limnicola; Pcor* = Polydora cornuta; Pele** = Pygospio elegans; Pkem* = Pseudopolydora kernpi; Ppal = Phoronis pallida; Ppla = Paraonellaplatybranchia;Sben* = Streblospio benedicti; Stan** = Sinelobus stanfordi. 47
N Correlation with Axes Low-Mid Estuary ...... 1 2 Al Salinity 0.98 -0.14 Temp -0.81-0.51 Heter0.79-0.60 A3.
Y6 Ldub Yl A
Eest / / A PeIe Abre Upper Estuary Hflo* A2 Gpo!.... NIimAA ASfan Csal A ......
Aa ACVU!HoreA ANcaIBP0Axis I A4 Eco A PpIa A Mba! sPiMare1A PpaI Cspi Salinity AentPkem A A PCOT* A Cryc Nhin* Maes A L.pug A Gins ArnodA Temperature HfiI* Hsub1 A Macrophyte D. AvaI* Y5 Y4 Sben Y3 .....Low-Mid Estuary I...... Yaquina Bay)
Figure 2.9 48
Figure 2.10. Ordination of 12 fishes, three decapods (code in parenthesis) and 12 high-tide intertidal sites along environmental gradients. Indicated are vectors of temperature, salinity and macrophyte density for the two main CCA axes and attendant intraset correlations. Sites correspond to those of Figure 2.1 and Table 2.1. Species origin denoted after species code: * (nonindigenous); no asterisk (native). Species codes: Aaff = Atherinops affinis; Asap* Alosa sapidissima; Cagg = Cymatogaster aggregata;(Cfran) = Crangon franciscorum;(Cmag) = Cancer magister; Cpail = Clupea pailasii; Emor = Engraulis mordax; Gacu = Gasterosteus aculeatus;(Hore) =Hemigrapsus oregonensis;Hpret =Hypomesus pretiosus; Larm = Leptocottus armatus; Otsh = Oncorhynchus tshawystscha; Phor = Pholis ornata; Pste = Piatichthys steliatus; Pvet = Pleuronectes vetulus. 49
c1
s...... Correlation with Axes A3, 1 2 Salinity -0.720.67 Al Temp-0.01-0.95 Heter-0.84-0.35
Low-Mid Estuary A2 \.. Hpret Pveta Salinity (Hore)
Macrophyte D. Pho Otsh Axis I Cagg Larm4 A4 (Cfran) S (çmag) Gacu Emor Pste £ £ £ Y4 AaffAsap* Y A5 A6
Low-Mid Estuary (Yaquina Bay) Temperature Upper Estuary Y5 S
Figure 2.10 Table 2.4. Percent of community variance explained, eigenvalues and correlations for the three main axes of CCA ordinations. Ordinations for invertebrates in core-sediment samples and fishes and decapods in seine samples are based on 12 high-tide intertidalareas sampled in Alsea Bay and Yaquina Bay during summer 1993.
Invertebrates Fishes and Decapods Statistics Core samples Seine samples Axes Axes 2 3 1 2 3
% variance explained 28.2 13.2 8.7 24.9 20.4 5.8
Cumulative variance 28.2 41.4 50.1 24.9 45.3 51.1
Eigenvalue 1 0.27** 0.13** 0.08** 0.16** 0.13** QQ4-H
2 Q93** Q9Q** LC Correlation 0.93* 0.93* 0.96** 0.72k
Monte Carlo test P-values: (<0.l2) , (< 0.10), *(< 0.05), **(< 0.01) 1 Variance extracted by main axes from the total variance in species data for core samples (0.971) and seine samples (0.648) 2 Pearson correlations derived from species data and samplescores that are linear combinations of the environmental variables. 51
With the exception of the NI polychaete Hobsonia florida, which was among the most dominant species at low salinities in both estuaries, the distributional centers of most NIS in sediments coincided with above average water temperature, salinity and macrophyte density, and high densities of NIS were more associated to Yaquina Bay sites 3 and 4(Figure 2.9) .Most fishes and decapods, including the NI fish Alosa sapidissima are associated with above average salinity and macrophyte density and with mid estuarine sites, particularly in Yaquina Bay (Figure 2.10) The distributional centers for dominant and ubiquitous species in core samples (e.g.,C. salmonis and Eogammarus confervicolus, Figure 2.9) and seine samples (e.g.,C. aggregata and L. armatus, Figure 2.10) are often near the center of environmental gradients.
Discussion
For each of our six stated questions we conclude that: The habitats available to NIS in Alsea Bay and Yaquina Bay are alike in temperature, salinity and macrophyte density. Such habitat similarities and the proximity between these estuaries may have contributed to the high co-occurrence of established NIS. Densities, CPUE and richness of NIS in core- and seine-samples were greater in Yaquina Bay. Thus, our findings support the increased risk of species introductions in Yaquina Bay by ships' ballast water, sediments, and fouling organisms. Nevertheless, native invertebrates and fishes dominated in total density, CPUE and richness in each estuary. Except for native and NI polychaetes and atheriniformes (Tables 2.2 and 2.3), cluster analyses revealed that the distributions and densities of taxonomically similar native and NI invertebrates are not more related in comparison to distant taxa. Hence, noncoevolved interactions between similar taxa may not be more likely in comparison to distantly related taxa. 52
Higher intertidal densities of NIS over low-high tides occurred at:1) high-mid temperatures in both estuaries, 2) mid salinities in Alsea Bay and 3) mid-low salinities in Yaquina Bay. However, NIS only dominated in density in Yaquina Bay at high temperatures and mid salinities. Species richness of invertebrates under each temperature-salinity condition in both estuaries was dominated by native species. Most NI invertebrates at high-tide were at maximum abundance in habitats of higher macrophyte density, water temperature and salinity when compared to native invertebrates. The latter habitat conditions also coincided with the distributional centers of most native fishes and decapods and the introduced american shad. Hence, biological invasions may have particularly affected the species interactions of native species associated with the previous habitat conditions, mainly in the mid section of Yaquina Bay.
Increased invasion risk due to ballast-water traffic is consistent with the higher establishment of NIS in Yaquina Bay. However, oyster culture may have been the main vector for the reported benthic invertebrate introductions in both estuaries. All eight NI invertebrate species found in Alsea Bay co-occurred in Yaquina Bay and in the latter estuary we only detected three additional NIS of invertebrates which may not be necessarily linked to ballast water discharge. The risk of species introductions by oyster importations is also higher in Yaquina Bay as oyster culture in Alsea Bay was discontinued since the 1930s (A. Robinson, personal communication 1995) .Thus, the three additional NI benthic invertebrate species in Yaquina Bay are not necessarily the result of ship-mediated invasions. Even in the latter case, the few NIS found exclusively in Yaquina Bay imply that species released from ships' ballast water and sediments and ships' hulls may only have been secondary vectors of NI macrobenthic invertebrates in Yaquina Bay. Some NIS with euryhaline meroplanktonic stages could have been transported between estuaries by coastal currents. However, a major redistribution of NIS species by advection could not 53 explain the 42% higher richness of native invertebrates detected in Yaquina Bay under the same sampling effort. Thus, currents may be a secondary source of species introductions. The lower densities of NI invertebrates in Alsea Bay may be due to: a shorter history of species introductions by oyster- culture, absence of ballast-water release, and less favorable environment for some NIS known to be indicators of habitat pollution and/or disturbance (e.g., Mya arenaria, Pearson and Rosenberg 1978; Strebiospio benedicti, Grassle and Grassle 1974; Heteromastus fiiiformis, Pearson and Rosenberg 1978)
Although the evidence for competition is greater among taxonomically distant groups in comparison to similar taxa (Woodin and Jackson 1979), co-occurrence among taxonomically similar noncoevolved groups in Alsea Bay and Yaquina Bay may have been historically higher than implied from cluster analyses. Most NIS found in our surveys are sedentary tube-builders (Table A.5), and are likely to compete with mobile species (Castillo et al. 2000) .Moreover, both negative and positive trophic interactions among NIS present in our surveys (Table A.5) and native species in Yaquina Bay have been implied (Castillo et al. 2000).
Clustering of sites from seine-samples suggested more faunal contrasts between estuaries than among downstream and upstream
sites (Figure 2.5) . However, sediment-inhabiting fauna form a
clear cluster in upstream sites (Figure 2.4, cluster 4) . The latter difference could be explained by the sedentary nature of invertebrates in sediments and/or their more physiologically restricted distributional ranges. Besides, substantially more fishes than decapods can migrate and use intertidal areas during flood tides (Figure 2.3)
The unexplained variance in species-densities in both sediment and seine samples in CCA ordinations may be accounted by variations in individual-species responses to the environmental factors considered, as well as to variations in other factors 54 within and between estuaries (e.g., tidal exchange, food availability, species interactions) .Further CCA analyses (G.G. Castillo, personal observation), indicated that ordination of fishes and decapods based on densities of benthic species from core sediments (native, NI and all species) were non-significant and the three main axes only accounted for 27.4% of the variance. Moreover, inclusion of invertebrate densities from core samples along with macrophyte density, water temperature and salinity only caused a non-significant increase over the total variance (<
5%) .Hence, OPUE of fishes and decapods along both estuaries are not related to invertebrate densities in sediments, regardless of the species' origin. Although prey depression by predators could make it difficult to detect a predator-prey relation, evidence from soft sediment communities clearly suggests reduced prey response following predator removal, even after extended periods (Peterson 1979) Hence, potential numerical responses of predators may be overridden by other factors determining habitat selection by predators. Additional CCA analyses also support a lack of association between invertebrate densities in sediments and the six major benthic species in seine samples (fishes: Leptocottus armatus, Pleuronectes vetulus, Platichthys stellatus, decapods: Cancer rnagister, Hemigrapsus oregonesis and Crangon franciscorum), (G.C. Castillo, personal observation)
The higher abundances of NI invertebrates in Yaquina Bay imply a greater trophic effect of biological invasion in that estuary when compared to Alsea Bay. Such hypothesis is supported by a significantly higher contribution of NIS to the prey-base of native juvenile flatfish in Yaquina Bay (Castillo 2000) Considering the similar patterns of prey selection between native and NI prey by juvenile flatfish in the field (Castillo 2000) and in laboratory experiments (Castillo et al. In press), no direct adverse trophic impacts of NIS for flatfishes in estuarine rearing areas are implied. Yet, biological invasions may have 55
reduced the stability of the benthic community in Yaquina Bay (Castillo et al. 2000)
Despite the importance of bottom-up effects by NIS of eelgrass and cordgrass in Northeast Pacific estuaries (e.g., Posey 1988; Daehler and Strong 1996) and the top-down effects of conspicuous NI invertebrates such as the clam Potamocorbula amurensis and the crabs Eriocheir sinensis (Cohen and Carlton 1995) and Carcinus maenas (Grosholz and Ruiz 1996), the bottom-up impacts of most estuarine NI invertebrates remain largely unknown.
Common or abundant taxa not identified to species level are: oligochaetes, nematods, chironomids, nemerteans and the Capitella species complex. The presence of the latter taxa and cryptogenic species in sediments imply that the extent of benthic invertebrate invasions in Alsea Bay and Yaquina Bay may have been underestimated. Five of the eight cryptogenic species co-occurred in both estuaries and the remaining cryptogenic species were rare
and only appeared in Yaquina Bay (Table 2.2) . Major difficulties in resolving the native geographic range of cryptogenic species often include both cosmopolitanism and unresolved taxonomy.
Besides of Crassostrea virginica and C. gigas, other NIS not detected in our surveys included: the Atlantic mud crab (Rhithropanopeus harrisii) .This species was only found in Yaquina Bay when trawling the channel area adjacent to site Y6 during summer 1993 (G.C. Castillo, unpublished data); the Asian amphipodGrandidierella japonica,observed in Yaquina Bay in 1994 and 1996 (J.W. Chapman, personal communication 1997; G.C. Castillo, personal observation) and in Alsea Bay in 1997 (J.W. Chapman, personal communication 1997); and the European green crab(Carcinus inaenas)detected in Yaquina Bay and Alsea Bay and other Northeast Pacific estuaries in 1998 (Behrens and Hunt, in press) Considering the high diversity and abundance of NI fishes in some California estuaries (e.g., Moyle 1985; Moyle 1986), the low 56 occurrence of NI fishes in our surveys does not imply low risk of fish introductions in Northeast Pacific estuaries. We found a single individual of the NI fish Lucania parva in Yaquina Bay. The NI American shad is reported in Alsea Bay by Monaco et al. (1990) but it did not appear in our samples. Two NI fishes reported only once in Yaquina Bay are the threadfin shad Dorosoma petense (Krygier et al. 1973) and the brown bullhead Ictalurus nebulosus (De Ben et al. 1990)
Several non-mutually exclusive causal mechanisms may play a role in producing species-area relations (Connor and McCoy 1979) Fish richness along U.S. west coast estuaries increases with estuary area (Bottom and Jones 1990; and mouth-depth (Monaco et
al. 1992) . Whether the latter factors or other processes could account for differences in richness of invertebrates among these estuaries remains to be tested. Species richness within estuaries may be enhanced by the presence of the native eelgrass Zostera marina in most sites of Alsea Bay and Yaquina Bay. The presence of Z. japonica in Yaquina Bay further suggests increased richness and densities of benthic fauna similar to those reported by Posey (1988) in invaded microhabitats of the South Slough of Coos Bay, Oregon.
The 3D plots of densities or richness of invertebrate under different temperature-salinity combinations synthesized community patterns not apparent from ordination analyses of species at high-tide. Because salinities decreased in most sites at low- tide, total densities of NI invertebrates in 3D plots are more related to lower salinities than those inferred from CCA analyses at high-tide. In the latter case, the highest densities of most NIS in sediment samples coincided both with above-average temperatures (> 17.8 °C) and salinities (> 22.50/) 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 57 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. 58
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Chapter 3
Trophic Contribution and Selection ofNative and NonindigenouS Prey by Native Fishes instuarine Rearing-Habitats
G.C. Castillo1,2 H.W. Li2, J.W. Chapman3, T.W.Miller3
1 Present address: Hatfield Marine Science Center.Oregon State University, Newport, OR 97365.
2 Oregon Cooperative Fish and Wildlife ResearchUnit, Department of Fisheries and Wildlife. Oregon State University Corvallis, OR 97331. Hatfield Marine Science Center. Oregon State University, Newport, OR 97365. 65
Abstract
We determined the summer prey richness, preycomposition 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 commonin 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 seemshigher 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 ofnative 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.
Introduction
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 66
1998; Castillo 2000) .Major mechanisms ofNISintroductions in Northeast Pacific estuaries include propagation ofspecies associated with imported oysters, ballast water release and fouling on hull of ships (Carlton 1979; Cohen and Carlton1998)
Potential impacts of conspicuousNISin U.S. west coast estuaries since the late 1980s (e.g., the clamFotamocorbula amurensis; the crabsEriocheirsinensis 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 byNISand 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 (Platichthysstellatus); staghorn sculpin (Leptocottus armatus); and chinook salmon (Oncorhynchustshawytscha)preyed upon at least one macrobenthic NIS (Castillo et al. 1995) .If native 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. 67
Assuming that predator-prey coevolutionis not a critical determinant of prey usage, we would expectsimilar selection patterns between native and NI prey typesby generalist predators, but the evidence is mixed. Preyselection by larvae of the introduced striped bass(Morone saxatilis)seems higher for at least one coevolved copepod whencompared to two NI copepods (e.g., Meng and Orsi 1991). In contrast,juvenile English sole did not consistently select two native amphipods overtwo NI amphipods (Castillo et al. In press)
Given the limited understanding on theimplications of coevolved and non-coevolved predator-preyinteractions in invaded estuaries, we address the following questions: 1)Is the richness of native and NI invertebrates in the fish dietproportional to their attendant richness in the environment?; 2) What arethe contributions of native species, NIS and cryptogenicspecies to the food-base of native fish in terms of frequencyof occurrence, number and volume of prey?; 3) Does the numberand volume of prey vary with fish size and weight?; 4) Aredaily dietary patterns in number and volume of prey evident?; 5)Is the total number of prey in the fish diet correlatedwith benthic densities of invertebrates and with CPUE of fish in intertidal areas?;and 6) Is the overall prey selection by native fishsimilar on native and NI prey types? To address these questions we surveyed intertidal and adjacent subtidal rearing-habitats ofjuvenile 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 reachhighest densities during summer before migrating to deeper habitats (Orcutt 1950; Krygier and Pearcy 1986; Boehlert and Mundy 1987) .Juveniles of 68
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). 69
Figure 3.1 70 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) .Only 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)
Methods
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°/ closer to the ocean (lower estuary) to 5°/ 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 (0indepth) 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 71 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 (1-600mm3) . Due to 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. 72
Computations
The frequency of occurrence of prey species k in the fish diet
(Ok)is defined as:
= 100
where kis 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 (i.e., native, NI and cryptogenic) in the fish diet (MO) using the formula:
MO =
Where °kj is the corresponding°kvalue for prey species , of origin.3 and m is the total number of prey of origin. We computed the mean number of all prey species of origin (MN) in the fish diet using the formula:
Mn = n_li [Nklj]
Where NkI is the number of prey species k of origin in fish for a total prey species of origin and for fish analyzed. Likewise, the mean prey volume of all species of origin (MV) in the fish diet is computed by substituting Nkl byVkj!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)
OICk= [Ok+%Vk+ %Nk]3'
which we defined by the percentages of prey frequency of
occurrence (Ok), volume (%Vk) and number(%Nk)contributed by itemk. The OIC index can range between 0%(i.e., itemk is absent in all fish examined) to 100%(i.e., only itemk occurs in all 73 fish examined) .We used the OIC index to rank up to 15dominant items in the fish diet. Estuaries are divided intothree sections to determine the OIC values in fish from downstream toupstream sites: lower estuary (sites 1-2), mid estuary (sites3-4), 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 includedin the OIC index.
We estimate prey selection of macroinvertebratesin the fish diet with Johnson's (1980) selection index(E1):
= n1 [r s]
where rjj is the rank of usage of prey itemby fish (based on the abundance of prey1 in the fish diet), is the rank of availability of item to fish (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 value. Only 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 arecompared 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. 74
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 Number Length(cm) Weight (g) Estuary of fish Mean SE Mean SE
English sole: Alsea Bay 135 6.8 0.1 3.0 0.1 Yaquina Bay 113 7.1 0.1 3.4 0.2
Starry flounder: Alsea Bay 61 9.0 0.6 14.9 3.1 Yaquina Bay 61 13.3 0.6 34.4 3.9
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)
Estuary No. Species No. Species (Occurrence) Species Environment Fish Diet
Native NI Native NI
Alsea Bay
English sole 31 8 18 (13.3) 5 ( 9.0) Starry flounder 30 8 8 (22.1) 5 (12.8) Both species 34 8 20 (16.0) 5 (10.2)
Yaquina Bay 8 English sole 49 11 20 ( 7.8) (14.5) Starry flounder 50 11 16 (10.7) 8 (17.9)
Both species 50 11 25 ( 8.8) 9 (15.7) 75
Results
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) .In 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 <0.05, t-test) .The fish diet in Alsea Bay is dominated by native species both in prey number (Figure 3.2: A, C) and volume (Figure 3.2: E,G). In Yaquina Bay neither native or NIS dominated in prey number (Figure 3.2: B,D) and volume (Figure 3.2: F, H).
The volume of major taxa in the fish diet was not consistently higher for native species. In Alsea Bay, the diet of English sole is dominated by native bivalves and crustaceans (Figure 3.3: A) but the diet of starry flounder is dominated by the NI bivalve Mya arenaria and by native polychaetes (Figure 3.3: C). 76
Table 3.3. Frequency of species occurrence in the diet of juvenile English sole(E) and starry flounder (S) in Alsea Bay and Yaquina Bay. Species origins: native (N); nonindigenous - Northwest Atlantic (A); Japan (J); Western Pacific (P)- and
cryptogenic (C) . Potential vectors of introduction are: oyster (0); fouling of ship hulls (F), and ballast water (B)
Estuary
Species Alsea Yaquina Origin /Vector F S F S
B iva lvi a
Clinocardium nuttallii 10.4 0.9 3.3 N Cryptomyacalifornica 2.2 3.3 1.8 1.6 N Macoma baithica 3.0 18.0 0.9 13.1 N Mya arenaria 17.8 3.3 4.4 23.0 A/O' Myselia tumida 3.0 1.6 N Transenneiiaeantilla 0.9 N
Crustacea
- Araphipoda: Aliorchestes angusta 1.8 N Ampithoe lacertosa 1.6 N Ampithoe valida 0.9 A/(B;F;O)1'4 Corophium acherusicum 2.7 3.3 A/(F;O)"4 Corophium saimonis 45.2 87.0 37.2 70.5 N Corophium spinicorne 31.1 9.7 31.1 N Corophium brevis 7.1 N Eobroigus spinosus 2.2 1.6 1.8 3.3 A/O2 Eogammarus confervicolus 9.6 4.4 1.6 N - Cumacea: Cumella vulgaris 47.4 1.6 22.1 4.9 N Nippoleucon hinumensis 16.3 6.6 38.1 18.0 J/B3'4 - Decapoda: Neotrypaea californiensis 4.4 2.7 9.8 N Crangon franciscorum 1.6 0.9 4.9 N Hemigrapsus oregonensis 5.2 N Upogebia pugettensis 2.2 1.6 N - Tanaidacea: Leptochelia dubia 0.7 12.4 3.3 C Pancolus californiensis 2.2 0.9 C Sinolobus stanfordi 3.0 3.0 C 77
Table 3.3. Continued.
Estuary Species Aisea Yaquina Origin /Vector E S E S
Polychaeta
Amaeana occidentalis 4.4 N Armandia brevis 39.3 4.4 N Boccardia proboscidea 5.3 N Dorvillea annulata 0.9 C Eteone spilotus 4.4 3.3 16.8 4.9 N Eteone californica 0.9 C Glycinde polygnatha 4.4 0.9 N Glycinde armigera 0.9 N Heteromastus filiformis 1.6 A/(B;O)1' Hobsonia florida 9.6 57.4 2.7 49.2 A/ O?)2) Manayunkia aestuarina 4.4 16.4 C Mediomastus californiensis 1.5 2.7 3.3 N Nephtys caeca 0.7 N Nereis limnicola 1.5 67.2 11.5 N Owenia fusiformis 1.5 0.9 C Paraonella platybranchia 1.6 N Phyllodoce hartmana 0.7 0.9 N Pseudopolydora kempi 6.7 3.3 25.7 36.1 P/(B;F';O)"4 Pygospio californica 1.5 N Pygospio elegans 17.8 1.6 10.6 1.6 C Streblospio benedicti 39.8 24.6
Teleost
CymatogasLer aggregata 1.6 1.6 N Engraulis mordax 1.6 N
Sources: 1Canton (1979); 2 J.T. Canton (Williams College, CT, personal communication 1996); 3Carlton and Geller (1993); Cohen and Carlton (1995) 78
Figure 3.2. Mean number and volume of prey species in the diets of English sole and starry flounder. Prey origin: native (NA); nonindigenous (NI) and cryptogenic (CR) .Different letters over the bars of each graph indicate significant differences between means (t-test, P < 0.05), with number of fish used in each estuary indicated in graphs A to D. 79
A English sole (Alsea) B English sole (Yaquina) No. Fish = 135 No. Fish = 113 a 40 40
30 30 >- w 20 20
10 10 u-.= b C 0 NA NI CR
C Starry Flounder (Alsea) D Starry Flounder (Yaquina) 50 No. Fish = 61 No. Fish = 61 a z 40 40 a 30 30 w 20 a 20
10 10 b C 0 0 NA NI CR NA NI CR
E English sole (Alsea) F English sole (Yaquina) a
30
20
0
G Starry flounder (Alsea) H Starry flounder (Yaquina)
250 250
200 200
150 150
100 100
50 50
0 0
PREY ORIGIN
Figure 3.2 80
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. 81
A English Sole (Alsea Bay) B English Sole (Yaquina Bay)
ST (1.1) CR 6.6) NI (5.1) NI (39.1) ... / NA (22.7) NA (16.7) NA (12.4)
-ST (0.5) ST (1.1) CR )0.9k, NI NI (1.1) ST (0.2) CR (0.4) NI (5.9) NA (4.2)
ST (7. ST (2.6)
CR (5.0)
NA (26.0) NI (8.0) NA (31.6)
C Starry Flounder (Alsea Bay) D Starry Flounder (Yaquina Bay)
1CR+ST (0.1)
NI (24.8) NA (29.0)
CR (<0.1) ST (1.0)- NA (1.1)
NA (6.6) NI (0.7). NA (9.6) CR + ST (0.3). NA (14.4)
NA (17.1)
NI (12.5) NI (33.5)
Figure 3.3 82
In Yaquina Bay, NI polychaetes and native crustaceans were major prey of English sole (Figure 3.3: B) and starry flounder (Figure 3.3: D). Only three of the 122 starry flounder examined had consumed fish, all native: one northern anchovy (Engraulis mordax) in Yaquina Bay and one shiner surfperch (Cymatogaster aggregata) in each estuary. Fish comprised the entire volume of native prey classified in the category of "other taxa" (Figure 3.3: C and D)
Habitats with the highest contribution of NIS to the fish diet are the mid section of Yaquina Bay in case of both fish species and in the lower section of Alsea Bay in case of starry flounder (Figures 3.4 and 3.5) .At least two NI prey per estuary section occurred among the 15 most important items of English sole
(Figure 3.4) . The NI polychaetes Strebiospio benedicti and Pseudopolydora kempiand the NI cumacean Nippoieucon hinumensis were prevalent in the diet of English sole from the mid section
of Yaquina Bay (Figure 3.4: D) . The NI polychaetesPseudopolydora kempi and Strebiospio benedicti were major prey of starry
flounder in the mid section of Yaquina Bay (Figure 3.5: D) . The NI bivalve Mya arenaria dominated the diet of starry flounder in the lower section of Alsea Bay where four other NI prey occurred (Figure 3.5: A).
3. Does the number and volume of prey vary with fish size and weight?
Yes, increasing fish size is associated with fewer prey in English sole (Table 3.4) and starry flounder (Table 3.5). Harpacticoids dominated the diet of the smaller English sole and starry flounder in Alsea Bay (Tables 3.4 and 3.5) .The importance of harpacticoids in Yaquina Bay was only apparent in the diet of
English sole (Table 3.4) . The diets of the smaller starry flounder in Yaquina Bay is dominated by NI polychaetes (Table 3.5), mainly Streblospio benedicti andHobsoniaflorida. Unlike fish size, weight of starry flounder is related to prey volume in both estuaries. However, only the volume of native prey 83
Figure 3.4. Percentages of prey frequency of occurrence, number and volume of main dietary items of English sole by estuary section. Frequency of prey occurrence is indicated under volume bars. Rank 1 indicates the prey with highest overall item contribution. Origin code for species-level taxa: nonindigenous (*); cryptogenic (**) native (no code) t8
brev7s ChironoiTidae -I Ohgochaeta Oligochaeta -i * * * hinumensis Nippoleucon elegans Pygospio I R -n ta/is 0cc/den Amaena parts Rlychaeta Streblospio * benedict! rmtter Rant I garis rrtter Rant vu! Cumella I vulgar/s Cumella spinicorne Corophium salmonis Comphium spilotus Eteone brevis Corophium peces Wood
pieces Wood I * kempi Pseudopolydora (4 oscidea prob Boccardia (s) balfhica Macoma 0) parts Fklychaeta J salmon,s Corophium (s) ba//h/ca Macoma * h,numensis Nippoleucon ** dubia Leptochelia * benedicti Streblosp!o -4 0) Harpacticoida C) Harpacticoida 0)
I I 0 (7, 0 01 0 (71 0 0 (7, 0 (71 0 (71 0 0 .4 (71 NJ N.) (71 4 (7' 1)J (7) fervicolus con Eogammarus Iimnicola Nereis pieces Wood flu/ta/li! Clinocardium californiensis Neotrypaea rrtter Rant I * kempi Pseudopolydora * hinumensis Nippoleucon * flonda sofia Hob * * elegans Pygospio sp Capitella parts CapeIlid parts FIychaeta 5) Otgochaeta 5) C.) * arenaria Mya spilotus Efeone (7 5) -4 parts Fklychaeta * hinumensis Nippoleucon (7 5) vulgaris Cumella salmonis Corophium 0 ** elegans Pygospio 5) (71 vulgaris Cumella Ostracoda brevis IArmandia I salmonis Corophium (7 01 pieces Wood -4 (Cl Harpacticoida Co (s) balthica Macoma -4 Harpacticoida (7 I I I I I I 0 (71 0 (71 0 (71 0 0 (71 0 (7) 0 (7 0 C) -J (Il I'.) p) (71 0 -J (71 N.) N.) (71
NUMBER % VOLUME % NUMBER % VOLUME % 85
Figure 3.5. Percentages of prey frequency of occurrence, number and volume of main dietary items of starry flounder by estuary section. Frequency of prey occurrence is indicated under volume bars. Rank 1 indicates the prey with highest overall item contribution. Origin code for species-level taxa: nonindigenous (*); cryptogenic (**) native (no code) 1% VOLUMEU PS PS% NUMBER O 4 % VOLUME Ce P-i % NUMBERP-i Ce -1 4% VOLUMECe P-i % NUMBERPS Ce 0 -4 0 0 0 Nereis limo/cola Ce 0 Ce 0 Ce 0 I CeI 0 L (I Ce 0 t Ce 0 C) Ce C-, 0 ¶ ...t Ce 0 Ce I 0 Ce 4 0 'CPS ¶ g Core ph/urn salmon/s F-rpaclicoida (SPS 0 tiers/s limnicolaMacoma balm/ca (s) PS [ I Wood p/eces Myaarenaria(s) *Corophium salmon/s (.1. a,CeA g(j Hobson,af florida * I Dptera pupae Cttironorridae -vC a,CeA p I Wood pieces t-lobsonia florida * CeA IIMacoma ianl patter barth/ca (S) Ce-4 I Wood piecesCorophiurn spinicorne (I,m Ce--4 z HarpacticoidaCorophiumttani nailer spin/come Ce0 Ce-S I OsiracodaCyrrialogaster aggregate Harpacticoida 0Ce I OstracodaFant PaperMacoma ba/thica m 0 vi Ce I Cryptomya cal/fornicaNip F'seudopolymiorapoleucon hinurnens/s kempi * * 'SPS ,-aMyaarenor/a * Macoma baithica (s) C.SP-S C-,P-ia . jjI Macoma ba/thicalHobsoniaflorida * CeA al Cyclopoida Cam siphon CeA CeA ° toB J Mia arenaria * Eobro/gus spinosus * 0Ce CePe 0 CePS Coroph/um salmonis 0Ce Ce S 0 m 0Ce (0 CeP4 0 CePS Pseudopol1eiora kempi * 0Ce Ce 1 0 CJ 0Ce C) - C-' CePS a 0 CePS 0Ce Ce -S 0 vi 0 aC-, Corophium sp/nicorneHobsonia florida * 'SP-i b2 Corophium salmon,s C-iPS I Macoma ballhica5nt (s) riot/er I WodI Karpacticoida pieces Macoma baithica (s) a,CeA I Wood pieces NippoleuconM'a arenaria hinumensis * * CeA MacmyraEngraulis ba/thica mordas Leptochelia dub/a .. j j Nereis limnicola I Manayunkia aestuarina Ce-4 Macomaklychaeta balihica parts (s) Ce -1 Ffe/ychaetal-rpacticoida parts cS Macomapantciiir000rridae trailerbait h/ca 0 CorophiumNeotrypaeaMacoma spin/comeCaliforniensis barth/ca 0Ce I NeotiypaeaEob re/gus sp/trosuscal/fern/ens/s * CrangonOstracoda franciscorum CS10 I Cap/telIldI-'arpacticolda parts CSPS I Pygospio elegans..Chnocardium flute/I/i C"A Pa 0 I Rtychaeta partsMya arenana * CeA I Eteone spllo/tusHobsonia florida * CeA at ., fI LlpogebiaNippoleucon Pseudopolysiorapugettensis hinumensis * kempi * 87
Table 3.4. Percent of number of prey by taxa and species origin for two size classes of juvenile English sole in Alsea Bay and Yaquina Bay. NI = nonindigenous prey. The ratio of fish containing prey items to the total number of fish analyzed is indicated under each size range. The mean prey number per size- range is also indicated (SE = standard error)
Taxa Alsea Bay Yaquina Bay Origin Total size range (cm) Total size range (cm) (3.0-7.0)(7.1-11.0) (3.0-7.0)(7.1-11.0) 66/68 60/67 40/49 50/64
Bivalvia Native 48.8 25.3 5.2 13.0 NI 0.1 2.2 0.2 0.2 All 48.9 27.5 5.4 13.2
Harpacticoida All 38.9 44.1 68.9 15.1
Macro-Crustacea Native 3.9 11.3 2.5 17.3 NI 0.6 0.4 8.5 15.7 Cryptogenic 0.0 0.7 0.6 9.0 Supra-specific 1.4 0.4 0.1 0.3 All 5.9 12.7 11.7 42.2
Polychaeta Native 1.0 11.9 2.9 7.5 NI 0.6 0.2 10.0 19.8 Cryptogenic 4.5 0.8 0.5 1.1 Supra-specific 0.1 1.0 0.1 0.3 All 6.2 13.9 13.5 28.8
Miscellaneous All 0.1 1.8 0.5 0.7
All taxa 100.0 100.0 100.0 100.0
5 No. prey /fish 108.7 24.4 64.9 57.4 SE 13.3 3.3 11.8 9.7 Table 3.5. Percent of number of prey by taxa and species origin for three size classes of juvenile starry flounder in Alsea Bay and Yaquina Bay. NI = nonindigenous prey. The ratio of fish containing prey items to the total number of fish analyzed is indicated under each size range. The mean number of prey per size range is also indicated (SE = standard error)
Alsea Bay Yaquina Bay Taxa Size interval (cm) Size interval (cm) Origin (3.0-7.0)(7.1-11.0)(11.1-24.0) (3.0-7.0) (7.1-11.0)(11.1-24.0) 31/31 13/13 14/17 14/15 9/9 34/37
Bivalvia Native 0.9 1.4 8.7 4.5 37.2 6.2 NI 0.0 0.0 15.4 2.1 2.0 2.1 All 0.9 1.4 24.1 6.7 39.2 8.3 Harpacticoida All 71.4 28.2 12.9 2.8 4.0 0.1 Macro-Crustacea Native 22.4 61.7 44.1 17.5 25.1 60.5 NI 0.3 0.0 1.0 4.5 1.0 0.9 Crypt o geni c 0.0 0.0 0.0 4.4 0.0 1.5 Supra-specific 0.0 0.0 0.0 1.0 0.0 0.8 All 22.7 61.7 45.0 27.4 26.1 63.8 Polychaeta Native 2.3 4.9 7.7 3.8 4.9 1.0 NI 2.1 3.5 5.1 57.0 23.2 26.1 Crypt o geni c 0.0 0.0 0.3 1.0 0.0 0.0 Supra-specific 0.0 0.0 0.0 0.0 0.7 0.2 All 4.4 8.5 13.2 61.8 28.9 27.3 Miscellaneous All 0.7 0.2 4.8 1.4 1.7 0.5
All taxa 100.0 100.0 100.0 100.0 100.0 100.0
No. prey/ fish 67.3 32.8 22.2 123.5 77.0 51.2 SE 19.0 8.1 3.6 24.1 21.7 9.8 89 and all prey combined are significantly related to starry flounder weight (Figure 3.6: B and D)
4. Are daily patterns of prey number and volume evident in the fish diet:(A) for all prey combined? and (B) between native and NI prey?
Yes, unlike starry flounder, English sole collected in the morning at low-tide contained fewer prey (Figure 3.7: A to D) and lower prey volume in comparison to high-tide afternoon hours
(Figure 3.7: E,F, P<0.05). Yes, proportionally higher prey numbers and volumes of native prey than NI prey occurred in Alsea Bay for each fish species during high-tide afternoon hours (Figure 3.7: A,C,G,E;
x2 test; p<0.001) . Except for prey numbers of English sole in Yaquina Bay, proportionally higher numbers and volumes of native prey than NI prey occurred in the fish diet from Yaquina Bay during low-tide morning hours (Figure 3.7: D, F, H; X2 test, P< 0.001)
5. Are total prey numbers of native, NI and all invertebrates in the fish diet correlated with intertidal benthic densities of: A) all invertebrates? and B) each fish species?
No, a relation between number of prey in the fish diet and invertebrate density among all intertidal sites (Figure 3.8: A to F) was only apparent for NI prey of starry flounder (r=0.78, p
<0.05) and possibly for NI prey of English sole (r=0.71, p< 0.10) No, the number of prey in the fish diet is not consistently correlated with intertidal CPUE for each fish species (r 0.55; p>0.10; Figure 3.8: A to D and G-H) .In addition, CPUE of each fish species were not clearly correlated with benthic invertebrate densities (r 0.53, p>0.10) 90
Figure 3.6. Total prey volume of native and nonindigenous species and all taxa combined as a function of fish weight in intertidal areas. An asterisk after R2 values indicates a significantlinear relation (P < 0.05) A English sole (Alsea Bay) B Starry flounder (Alsea Bay) All taxa, R2 = 0.43*) 600 (1:AlI taxa, R2 <0.00) 3000 Native, R2 = 0.44*) (2:Native, R2 = 0.09) A NI, R2 = 0.03) ' (3:NI,R20.01) n28 400 n29 2000 ,
C., E 200 1000 I uJ 2 3 3 0 123 45 67 8 0 20 40 60 80 100 120 C English sole (Yaquina Bay) D Starry flounder (Yaquina Bay) >- (1:All Taxa, R2 = 0.16*) Ui 600 -(I:A!I Taxa, R' <0.00) 3000 '(2:Native, R2=0.15*) (2:Native, R2 = 0.03) 0 A (3:NI, R2 = 0.03) (3:NI, R2 = 0.02) n41 400 2000 'I
200 U U 1000 A . U 2 2 '3 3 20 40 60 80 100 120
TOTAL FISH WEIGHT (g)
Figure 3.6 92
Figure 3.7. Mean number (A-D) and volume (E-H) of prey in the diet of juvenile English sole and starry flounder collected at low- andhigh-tide. Prey origin: native (NA); nonindigenous (NI); cryptogenic (CR) and supraspecific taxa (ST). Letters over bars indicate significant differences between mean numbers or volumes of all prey combined between low and high tide (t-test; P < 0.05) and numbers over bars indicate the ratios of NA to NI prey in terms of prey numbers (A-C) and volumes CE-H), with number of fish used in each estuary in parenthesis (A-D) 93
A English Sole (Alsea Bay) B English Sole (Yaquina Bay)
200 Ii.'NA b 200 LUJJNA NI 85.7 (n 29 E.iiii NI
I ICR 150 CR 150 ST ST >- b LU 100 100 0.9 n 60 a: 0 a IIIIIIIIIIIIIIIII 50 11.9n 106) 50 1.2 n=53 LL h.iiIjiIhii(I: '-'Cl) LOW HIGH LOW HIGH mw C Starry Flounder (Alsea Bay) D Starry Flounder (Yaquina Bay)
I NA 80 a 5.7 n= 33 NI zo EJ CR 6O ST Z60 ST a a 40 10.6 (n= 28 40 n 18 LU 20 20 IIIIIIIIIIIIIIIII
0 0 LOW HIGH LOW HIGH
E English Sole (Alsea Bay) F English Sole (Yaquina Bay) 0 NA DIIJNA CR 100 100 Nl ST JCR 75 ST LU b 50 9.5 a 25 4.4 IIIIIIUIIIIIIIII >Lio. JJIIIIIIIJHIIUI LOW HIGH
G Starry Flounder (Alsea Bay) H Starry Flounder (Yaquina Bay) NA iINI ICR ST I I NA ICR a NI ST 1.7 E 400 400 0.9 <300 300 Ui 200 200
100 100
LOW HIGH LOW HIGH
TIDE
Figure3.7 94
Figure 3.8. Mean number of invertebrates in the flatfish diet (A- D), total densities for all invertebrates in the benthos CE-F) and CPUE of flatfish (G-H) .Included are all high-tide sites in the Alsea Bay and Yaquina Bay. T-bars denote standard error. 95
250A EnglishSole Diet 250: B English Sole Diet
All Taxa All Taxa 200 NatIve 200 --0-Native -- Non-IndIgenous - - - Nonlndigenous 150 / / 100. 100 >- W - == -_50 - 50 -___._&------
0 0 I 3 4 5 6 23456 Q200: 200- C Starry Flounder Diet D Starry Flounder Diet .--AllTaxa All Taxa ---c-- NatIve - O-NatIve 150 - .- Non-Indigenous Non-Indigenous
100
...... 50 50 \ n 0 -.- -- 0 '0 3 4 6 1 2 3 4 5 6
50 E Intertidal Benthic Invertebrates F Intertidal Benthic Invertebrates All Taxa All Taxa N40 0 - Native - a - NatIve - Non-indigenous -. Non-indIgenous >- 30
Z 20-
.0 10 I- 0 0 3 5 56
60- 60-G Intertidal Fish H Intertidal Fish N 50- 50- bigllsh sole E -.-- gIlsh sole Starry flounder WeC40 Starry flounder 40 0. 30= z020- 20- 10- 0- 0- I 4 5 6 2 5 6
ALSEA BAY SITES YAQUINA BAY SITES UPSTREAM -* UPSTREAM -*
Figure 3.8 96
6. (A) Are individual prey selected differently? If so,(B) are native and NI prey types selected differently?
Yes, prey selection varied among individual preyin the diet of each fish species in Alsea Bay (Figure 3.9) andYaquina Bay (Figure 3.10). No, similar selection between native and NI prey types was apparent for both fish species in Alsea Bay(Kruskal-Wallis test,
P>0.54) and Yaquina Bay (Kruskal-Wallis test, P>0.83).
Highly selected prey between estuaries included the NI polychaete Pseudopolydora kempi in the diet of starry flounder and Cumella vulgaris and Macoma balthica siphons in the dietof English sole (Figures 3.9 and 3.10). Highly selected prey in Alsea Bay by both fish species included the NIS Nippoleucon hinurnensis and Pseudopolydora kempi(Figure 3.9: A and B). In Yaquina Bay only the native cumacean Cumella vulgaris was among the five most selected prey by both fish (Figure 3.10: A and B) .Only starry flounder included NIS among the five most selected prey in Yaquina Bay (Figure 3.10: B).
Unlike prey usage, prey availability was significantly correlated with prey selection by each fish species (r 0.58, P <0.002) .Availability of native species, NIS and all items in the fish diet are significantly correlated with their corresponding usage by each fish species (r 0.59, P<0.05) 97
Figure 3.9. Johnson's selection index (line) and ranks of prey usage and availability (bars) for flatfishin Alsea Bay. Selection rank 1 indicates the item most selected. Prey availability and usage increase toward rank 1. Differences in individual prey selection were significant among English sole items (F9,21 = 56.1, P < 0.001) and starry flounder items (F10,16= 8.48, P <0.01). 98
A JOHNSONS SELECTION INDEX (ENGLISH SOLE ALSEA BAY)
-3 -2 -1 0 1 2 3 0 oleucon hinurnensis*
2 Cumella vu! z 4- Pseudo dora kern
Eteone spilotus
Pygospio e!egans**
Capite!!a sp. 8 Corophium sa/rnonis
10 10 8 6 4 2 0 2 4 6 8 10 RANK AVAILABILITY RANK USAGE
B JOHNSONS SELECTION INDEX (STARRY FLOUNDER ALSEA BAY)
-3 -2 -1 0 1 2 3
Di .tera a californica
Pseudo0! dora kem.i z .r. phium salmonis N...l hi Coro hium s inicorne 0 '0Nereis /irnnico/a Macoma baithica Chironomidae
10 Macorna ba/thica S Hobsonia florida*
12 10 8 6 4 2 0 2 4 6 8 10 RANK AVAILABILITY RANK USAGE
Figure3.9 99
Figure 3.10. Johnson's selection index (line) and ranks of prey usage and availability (bars) for flatfish in Yaquina Bay. Selection rank 1 indicates the item most selected. Prey availability and usage increase toward rank 1. Differences in selection among items were significant for English sole (F20,31= 58.9, P < 0.001) and starry flounder (F1625 = 63.1, P <0.001) 100
A JOHNSON'S SELECTION INDEX (ENGLISH SOLE YAQUINA BAY) -8 -6 -4 -2 0 2 4 6 8 0 Cumella vulgaris 2 Chironomjdae 4 Eteo e Il. a balthica (S) 6 S Ni000levcon hinumensis Hnhcnni flnriria* 8 Fngammanic rnnfprvirnlus Corophium brevis 10 Efnne piIrttc 12 IpfncheIidi,hi Boccardia Droboscideá 14 Streblospip benedicti* Corophium sninico Pseudopolydora kern pi* 16 Pyqospio eIejans1 18 Mediomatu.s ca1ifnmincis Capitella sp. 20 Corophium salmonis OIigQc haeta 22 15 10 5 0 5 10 15 RANK AVAILABILITY RANK USAGE
B JOHNSON'S SELECTIONINDEX (STARRY FLOUNDER YAQUINA BAY) -6 -4 -2 0 2 4 6 0 1 NentrvneR c.Iifnrniencjs 2- Mviaranari* Cumella vulaaris 4- rakern pi* 6- CoroDhium sninicorne Nereis limnicala Eteone snilatus 8- Streblospio benedicti* 10 Macoçna haIfhiri 1S CaDitellaSD. 12 - Corophium sa NippoIiirrr1 hinnmncLc * 14 - Hobs Macama baithica 16 Manay
18 15 10 5 0 5 10 15 RANK AVAILABILITY RANK USAGE
Figure 3.10 101
Discussion
For the six stated questions we conclude that: Richness of native species and NIS in the diets of native fishes are proportional to their corresponding richness in the benthic habitat. Then, fish do not distinguish between native species and NIS. Neither native species or NIS dominated the diet of English sole and starry flounder in Yaquina Bay in terms of mean frequency of prey occurrence and prey number and volume. However, native prey dominated the diet of the latter two species in Alsea Bay, both in number and volume of prey. The mean frequency of prey occurrences, numbers and volumes of native and NI prey were higher in comparison to cryptogenic species in each estuary. The diet of larger English sole and starry flounder included fewer prey, particularly, fewer harpacticoids. Unlike NI prey, the total volume of native prey increased with the weight of starry flounder in Alsea Bay and Yaquina Bay. Thus, despite the lack of fish bias in terms of richness of native and NI invertebrate prey types, larger starry flounder may rely more on native prey in these estuaries. Prey number and prey volume of English sole were lower during low-tide morning hours in comparison to high-tide afternoon hours. Reliance of English sole and starry flounder on native prey seems higher in intertidal areas of Alsea Bay and in subtidal areas of Yaquina Bay adjacent to the intertidal zone. A relation between the combined number of prey in the fish diet and benthic invertebrate densities along intertidal areas was only apparent for NI prey. However, total numbers of NI and native prey in the fish diet are not consistently related to intertidal CPUE of fish. Native and NI prey types are similarly selected by juvenile English sole and starry flounder in Alsea Bay and Yaquina Bay. Thus, individual prey characteristics are more critical than predator-prey coevolution in determining prey selection. 102
Introduced species are a major part of the food-base for juvenile English sole and starry flounder in Yaquina Bay and in mid Yaquina Bay areas these NIS have greatly altered the original estuarine food webs supporting demersal fish. However, richness of NIS in the fish diet was similar between Alsea Bay and Yaquina Bay (Tables 3.2, B.1 and B.2). Despite the difference in contribution of NI prey to the fish diet between Alsea Bay and Yaquina Bay, weight-length relations and condition factors for each fish species between these two estuaries were similar (Appendix B, Figures B.1 and B.2). Although such well-being indices do not necessarily imply similar food quality and growth rates of fish between the studied estuaries, similar growth rates for English sole are also apparent between juveniles residing in Yaquina Bay and offshore nursery grounds (Rosenberg 1982) and among Yaquina Bay, Monterey Bay, Puget Sound (Krygier and Pearcy 1986) and for combined growth estimates in two Washington estuaries and adjacent nearshore areas (Shi et al. 1997) .Then, the higher standing stock and productivity of estuaries in comparison to offshore areas (e.g., Krygier and Pearcy 1986; Gunderson et al. 1990) may not result in higher growth rates for juvenile English sole. However, considering the significantly higher densities of age-0 English sole in estuaries than in offshore areas (Krygier and Pearcy 1986), the growth potential of age-0 English sole could be higher in estuaries than in offshore areas if the reported density-dependent growth at age-1 (Peterman and Bradford 1987) also occurs at age-O.
The broad prey-base of juvenile flatfish between Yaquina Bay and Alsea Bay and the similar prey selection between native and NI prey types supports the generalist feeding strategy suggested for estuarine fish in the east U.S. coast (Miller and Dunn 1980) but contrasts with the specialist feeding strategy of juvenile salmonids in some Northeast Pacific estuaries (e.g., Healey 1979;
Levings 1980) . A generalist strategy for juvenile English sole and starry flounder is further implied by their non-significant difference between interspecific and intraspecific diet overlaps 103
(Appendix B, Figure P.3) .Such feeding pattern may be due at least to two non-exclusive factors: First, a flexible foraging strategy (e.g., broad range of prey-images and foraging tactics), as implied by the comparable proportions of species richness in the fish diet and in the environment between native species and NIS. Second, many native and NI prey in Alsea Bay and Yaquina Bay have similar morphologies and lifestyles that may facilitate a broader predator niche (e.g., Sih 1987).
Although we did not investigate nocturnal feeding activity of fishes, the observed diurnal feeding pattern of English sole is not an artifact due to potential differences in fish size. Stomach fullness of fish collected in the morning was lower than in the afternoon for English sole (Table B.l) but not for starry flounder (Table B.2) .Increasing prey volume throughout the day in the absence of intertidal fish migration is consistent with visually oriented predation. Such predation mode is suggested for age-0 English sole both from offshore nursery areas (Hogue and Carey 1982), and from laboratory experiments under different visibility conditions (Castillo et al. In press) .Nevertheless, English sole larger than 24 cm TL may prey more heavily at night (Becker and Chew 1987) .By contrast, larger starry flounder (30- 56 cm TL) may only feed during daylight hours (Miller 1967) Although the mixed semidiurnal tides in Northeast Pacific estuaries may confound the extent of tidally-driven feeding patterns detected for other fish species exposed to regular tidal cycles (e.g., Thijssen et al. 1974; Summers 1980), fish migrations to intertidal areas may enhance prey consumption.
We ascribe the apparent lack of association between the Johnson's prey selection index and prey usage to variations in usage of particular prey among fish. In contrast, prey selection in both estuaries seems related to prey availability. Since prey selection depends on prey preference, detectability and ease of capture (Paloheimo 1979), higher prey availability may favor their detectability and ease of capture. However, differences in preference among prey cannot be inferred from our field selection 104 analyses as resource preference is often considered a choice made at equal availabilities (e.g., Ellis et al. 1976).
Prey availabilities may be influenced by factors not evaluated in our study (e.g., prey distribution, size, behavior, sediment type, turbidity) .Nevertheless, the implied similar selection between native and NI prey types is supported by the fact that many native and NI prey in Yaquina Bay and Alsea Bay arealike in terms of:1) taxonomic groups,2) individual prey volume (Figure B.4), 3) life modes, 4) vertical distribution and functional classifications (Castillo et al. 2000) .Moreover, a lack of selectivity by fish for native or NI prey types is also supported by most laboratory feeding experiments in which juvenile English were exposed to equal availabilities of two native and two NI amphipods (Castillo et al. In press)
The studied areas may include NI harpacticoids introduced through mechanisms available to NI macrofauna. Yet, the taxonomy, ecology and evolution of harpacticoids and other meiobenthic taxa are too poorly documented to determine their origins. Although the percent volume of harpacticoids(supraspecific taxa in crustacea, Figure 3.3) was minor in comparison to their numbers in the fish diet, harpacticoids are among the most important prey for the early juvenile stages of starry flounder (McCall 1992) and English sole (Toole 1980)
Despite the similar dietary overlap between fish species, the polychaetes Hobsonia florida and Nereis limnicola, and diptera (including chironomids) are more common in upstream areas where salinities are less than l5°/ and where starry flounder are more abundant than English sole. In contrast, the cryptogenic cosmopolitan polychaete Pygospio elegans are more common in the diet of English sole collected at downstream intertidal areas with salinities over 3Q0/. The NI polychaete Streblospio benedicti is a major prey for fish in Yaquina Bay. However, it was absent in all fish and benthic core samples from Alsea Bay. 105
More species could be added to the list of native or NI prey if cryptogenic species are resolved. The Capitella species complex, a group of opportunistic polychaete species usually erroneously referred to as "Capitella capitata" (Grassle and Grassle 1976), was a minor cryptogenic component in the diet of both fish species. However, it is a major subtidal prey for English sole and two other flatfish species in a developed Puget Sound area (Becker and Chew 1987)
Pleuronectids are not among the 44 fish families in which plants are important food items (Gerking 1994) .However, plant matter and woody debris were common items in the diet of English sole and starry flounder in our study (Figures 3.4 and 3.5) Although the latter items could have been incidentally ingested with other prey, their potential nutritional value deserves further study considering their presence elsewhere in the diets of both English sole (e.g., Toole 1980) and starry flounder (e.g., Orcutt 1950) along with potential digestive role of microflora in the digestive system of some fish (Gerking 1994)
Although the meaning of dietary indices that combine various prey attributes (e.g., frequency of occurrence, prey number, volume and weight) is less intuitive when compared to individual prey attributes (Bowen 1983), the dietary importance of a prey cannot be directly inferred from a single attribute (e.g., a prey with the highest volume may not be the most nutritional) Nevertheless, our index of overall item contribution is meaningful in identifying potentially relevant prey items because this index was significantly related to each of the prey attributes included (r 0.63, P<0.05) and because all such attributes are interrelated in the diet of each fish species in each estuary (r 0.52, p<0.001, Tables B.3 and B.4)
Based on the age-length relations for English sole (Rosenberg 1982) and starry flounder (Campana 1984) our field data suggest an estuarine residence periods of up to 3 years for starry 106 flounder and up to 1 year for English sole. Besides, some starry flounder of ages 0 to 2 can also rear in fresh water (Moyle 1976;
G.C. Castillo, personal observation) . Starry flounder is then potentially more susceptible to long-term shifts in estuarine food webs and other human impacts in comparison to English sole. A higher sensitivity of starry flounder to such impacts is consistent with its sharp long-term decline in commercial landings from Oregon (Berry et al. 1980; Lukas and Carter 1998)
Although our study does not suggest direct adverse trophic effects of biological invasions on juvenile flatfish, other short and long-term effects of noncoevolved species interactions are possible. Qualitative modeling on intertidal communities from Yaquina Bay implies that biological invasions have increased the risk of decline in community stability (Castillo et al. 2000) Estuarine nursery areas are usually assumed to confer protection to juvenile fish from predators when compared to offshore areas (e.g., Haedrich 1983). However, Gunderson et al. (1990) suggested higher predation on English sole and Dungeness crab in two Washington estuaries in comparison to offshore areas. The latter authors reported that most early age-0 juvenile English sole eventually migrate into estuaries where they reside before start migrating offshore at about 7.5 cm TL. Hence, the migration between alternative nursery areas should be considered when assessing the potential role of biological invasions and other impacts on growth, survival and recruitment.
The fact that summer CPUE of fish along our study sites were not clearly associated with total prey densities in the environment is consistent with the evidence that individual home ranges of many fish species are not optimal habitats for feeding and growth (Sogard 1994) . Alternatively, fish may not be limited by food (S.M. Sogard, personal communication, 2000), or CPUE of fish may be accounted for by other environmental factors
(Castillo 2000) . However, we would expect to observe a closer relation between densities of prey and CPUE of their fish 107 predators over an annual cycle, particularly in larger estuaries (e.g., Bottom and Jones 1990). Spatial variations for the combined density of many prey species should be reduced in comparison to individual prey species. Hence, the apparent lack of relation between the total number of native species in the fish diet and intertidal invertebrate densities along the estuaries does not apply to individual prey. Further biological invasions may not enhance the food availability of native benthic fish for several reasons, including:(1) few NIS account for a large portion of the fish food-base (Figures 3.4 and 3.5);2) selection varies greatly among individual NI prey (Figures 3.9 and 3.10); and 3) introduced predators (e.g., Kimmerer et al. 1994; Cohen et al. 1995; Grosholz and Ruiz 1995), cordgrass (e.g., Daehler and Strong 1996; Feist and Simenstad 2000) and other nuisance species could reduce the value of shallow rearing habitats for native estuarine-dependent species. 108
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Chapter 4
Predation on Native and Nonindigenous Pamphipod Crustaceans by a Native Estuarine-Dependent Fish
G.C. Castillo 1,2 H.tL Li2, J.W. Chapman3, T.W. Miller3
1 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.
Submitted to First National Conference on Marine Bioinvasions, J. Pederson (ed.), MIT Sea Grant College Program, January 1999. Cambridge, MA. In Press. 114
Abstract
The importance of nonindigenous species (NIS) within guilds supporting native species in higher trophic levels is a critical concern in the biology of invasions. We find that predator-prey coevolution may not allow predicting the order of consumption and selection for similar prey types. We conducted laboratory experiments to test for prey selection by juvenile English sole (Pleuronectes vetulus - native to the west coast of North America), using native amphipods (Corophium salmonis and C.
spinicorne) and northwest Atlantic arnphipods(C. acherusicumand C. insidiosum) .Single-species prey consumption in sand substratum was greater on C.spinicorneand C. acherusicum than on C. insidiosum and C. salmoni.s. Prey selection on both NIS was significantly greater than on native species over mud substratum but not over sand substratum. Predation of all Corophium species was greater over sand substratum than over mud substratum. No sex-selective predation occurred on any species in either substratum type, and prey size-selection was only suggested for C. acherusicum in both substrata types. Interspecific prey selection may vary with visibility, substratum type and prey behavior. Both NIS of amphipods are potentially capable of supporting higher trophic levels of native species.
Introduction
Estuaries in the Northeast Pacific are among the most invaded aquatic habitats in the world (Carlton 1979; Chapman 1988; Cohen and Carlton 1995) but almost nothing is known of the ecological effects of these invasions. Consumption of nonindigenous species (NIS) by native fish in estuaries can be substantial (Castillo et al. 1996) .This study is the first comparison of selection for native and introduced species in estuaries by native fish (i.e., the proportion of a given prey in the diet relative to its proportion in the environment) .The study of Meng and Orsi (1991) 115
suggests that the larvae of the introduced striped bass (Morone saxatilis) select one coevolved copepod species over two noncoevolved introduced copepod species. In the Yaquina Bay and Alsea Bay estuaries in Oregon, predator-prey coevolution may not affect the order of prey selection by two species of juvenile flatfish (Castillo 2000) Prey selection by fish may depend on: visibility and exposure (McCall 1992; Schiacher and Wooldridge 1996); activity (Ware 1973; Magnhagen 1986); evasion (Fulton 1982); absolute and relative density of prey (Magnhagen 1985); size (Ringler 1979); social facilitation (Brawn 1969; 011a and Samet 1974) and water temperature (Moore and Moore 1976) .We measured differences in consumption and selection of two native species and two NIS of amphipods (Corophium spp.) by juvenile English sole (Pleuronectes vetulus) .If juvenile P. vetulus tends to prey opportunistically at equal prey densities, we would expect to find higher consumption and selection toward the most vulnerable prey species, regardless of a potential coevolved predator-prey relation. We tested whether the number of prey consumed by P. vetulus varies with species, size, or sex of amphipods, and whether prey selection varies in the presence of alternative prey or substratum type. We also tested whether the visibility and activity of prey differ among the four amphipod species.
Juvenile P. vetulus use Northeast Pacific estuaries as rearing areas where amphipods are an important part of their diet
(Haertel and Osterberg 1967; Toole 1980) . Corophium salmonis and C. spinicorne are both temperate species native to the west coast of North America. Corophium insidiosum and C. acherusicum are both semitropical species inadvertently introduced into U.S. west coast estuaries from the east coast of North America (Carlton 1979) .All four species are abundant in Northeast Pacific estuaries.
The increased invasion rate in San Francisco Bay (Cohen and Carlton 1998) suggests that the abundances of introduced species 116 have also increased in other Northeast Pacific estuaries. Whether native amphipods are being replaced by NI (nonindigenous) amphipods in Northeast Pacific estuaries, as reported for freshwater amphipods in Ireland (Costello 1993), is uncertain but the native amphipod,Corophiurnbrevis, may be extinct in San Francisco Bay and populations may have declined in Humboldt Bay, California, following the introduction of at least one arnphipod (J.W. Chapman and T.W. Miller, unpublished data).
Methods
Juvenile P. vetulus were collected during summer 1996 from intertidal flats of Yaquina Bay by seining and were then transported to the Hatfield Marine Science Center (HMSC) on the same day. The fish were treated with a 1:4,000 formalin solution for 1 h to kill parasites and reduce fish mortality (Kamiso and Olson 1986) .The fish were sorted by size and maintained in continuous water flow and natural photoperiod. Fish used in experiments ranged from 5.1 cm to 6.6 cm total length (mean = 5.7 cm, sd = 0.35, n = 90) .Juvenile fish were fed live Tubifex, defrosted Artemia sauna, and 1-mm food pellets (Bioproducts) The varied diet conditioned the fish to multiple food types.
Native amphipods were collected from marina floats and mudflats of Yaquina Bay and used directly in experiments. Populations of the NI C. acherusicum and C. insidiousum were collected from floats and boats in Humboldt Bay and Yaquina Bay and cultured in the laboratory. Corophium acherusicum were also obtained from additional cultures (John Sewall, U.S. EPA,
Hatfield Marine Science Center, Newport, OR 97365) . Cultured amphipods were held in aerated rectangular 8.7-1 dish pans at 30 ppt and 25°C. Cultured diatoms (Chaetoceros calcitrans) were provided twice weekly and a mixture of powdered dry food (parts per ingredient: 1.3 Neonovum, 10 alfalfa, 20 Tetramin, and 10 wheat grass leaves) was provided every other day. The four amphipod species were treated with antibiotics (after Pelletier 117 and Chapman 1996) for 3 d prior to the experiments to increase survival. Older juveniles and adults retained on a 351-im sieve (Tyler Standard) were used in the experiments.
All amphipod prey populations were maintained at ca. 20°C for a minimum of 4 d prior to the experiments in the fall of 1996. Twenty-four acclimated amphipods were introduced into 5.8-1 glass aquaria (23 cm L x 15 cm W x 17 cm D) containing aerated brackish water (30 ppt salinity, 14°C) and a 0.5-cm layer of benthic sediment. Amphipods were left undisturbed for 24 h to allow tube building in the sediment. Then one juvenile P. vetulus was introduced into each tank. The fish were left undisturbed in the tanks for 48 h (from 6 p.m. to 6 p.m.) and then removed. The water and sediment of each tank were sieved to recover amphipods. All amphipods remaining on the sieve were then counted. The number of prey consumed was estimated from the initial number of prey minus the number found on the sieve.
Single fish were placed in tanks containing 24 defrosted Artemia sauna during each of the four species treatments to control for variations among predators. A given treatment was considered invalid and repeated when Artemia were not consumed in the control. Twenty-four amphipods were added to each of three tanks and maintained without predators to estimate losses other than fish predation. Both controls were used in single- and mixed-species experiments.
Single-factor ANOVA was used to test the significance of the difference between treatments. The Tukey multiple range test (hereafter, HSD, Sokal and Rohlf 1995) was used to test the significance of pairwise differences between treatment means. The chi-squared statistic (X2) was used to test the significance of the difference between treatment and control means. 118
Single-Species Experiments
Prey Visibility: Predator-free experiments were used to test whether amphipod visibility in tanks with sandy sediment varies among species at 14°C. Amphipod visibility was assessed from 2- mm observations for individuals swimming, walking, or partially visible above the sediment of 10 tanks. Twelve males and 12 females per species were introduced in each tank 1 h before the first observation. These observations were repeated after 1 h, after 3 h and then every 3 h up to 24 h.
Prey Activity: Interspecific differences in amphipod activity was tested in predator-free experiments. Activity was estimated from the average distance traveled per 5-sec in plastic containers (8 cm L x 8 cm W x 10 cm D) under the following conditions: no sediment, water at 14°C and at 24°C and 30 ppt salinity. One amphipod per species and sex was introduced in each of 10 replicated containers and left undisturbed for 24 h before the tests. Amphipod activity was measured with and without disturbance caused by suctioning from a small pipette as a proxy predation attempt (after Neng and Orsi 1991)
Prey Consumption: All single-species predation experiments were performed on sandy substrates (98% sand and 2% mud) since P. vetulus occurs predominantly in sandy substrates in Yaquina and Alsea Bay estuaries. Single-species treatments consisted of 24 amphipods (12 randomly selected individuals of each sex) in a tank with sandy substratum. Treatments were replicated 10 times for each amphipod species. Fish were deprived of food for 72 h prior to the feeding trials.
The Strauss' (1979) index of food selection (L)is adapted to estimate size-dependent selection by fish: = r1 -p1
where Lis the selection for prey of size ; r1 and p are the proportions of prey of size consumed and available prior to the 119 feeding trials, respectively. The value of L ranges from -1 to +1. Thus, the more selected sizes will have the largest L value. Alternatively, size selection was estimated from the cumulative difference in size distributions between eaten and uneaten individuals (Kolmogorov-Smirnov test, Tate and Clelland 1957) Precise determination of prey size and sex were obtained from the length and morphology of the fourth article of the second antenna. Measurements were made using a stereomicroscope equipped with an ocular micrometer. The size distribution of consumed amphipods in each tank was estimated from the size distribution of prey measured before the experiments minus the size distribution after the experiments.
Mixed-Species Experiments
The mixed-species treatments consisted of 24 amphipods (3 females and 3 males per species) in sandy substrate (98% sand and
2% mud) and muddy substrate (2% sand and 98% mud) . Each treatment was replicated 25 times in the same aquaria used in single- species experiments. Variation in amphipod visibility with substratum type was assessed for all species combined in the 25 tanks prior to the feeding experiments. Fish were starved for the 48 h preceding these experiments. The proportion of the tank bottom free of sand upon completion of the experiments was used as a proxy measure of fish disturbance. The high turbidity in mud treatments only allowed determinations of only the proportions of tanks in which the bottom was visible.
Results
Prey Sizes and Weights
The body lengths and weights of the two native species are greater than those of NIS (Figures 4.1 and C.l) and antenna 120
MALES FEMALES
C. salmonis C. salmon is ii 2 Y = 2.269X°4122 + 2.269X°4116 V 0.758 + 7.386X (R2 0.80,38 d.f.) (R20.81, 37 cli.) 0 00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0
C. spinicorne C. spinicorne
2 Y= 0.579 + 4.891X - 1.31 1X2 y .0.336+ 6.916X- 1.451X2 (R2= 0.86, 39 d.f.) (R20.87, 39 d.f.)
00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0
C. acherusicum C. acherusicum
Y= 1.396 + 1.797X V = 1.785 e14 (R2= 0.80,39 d.f.) (R2 0.40, 39 d.f.)
0 0 00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0
C. insidiosum C. insidiosum 6 6Y= 0.429 + 6.113X- 3.848X2 Y.2.05 + 25.03X -31.45X2 (R2= 0.69, 38 d.f.) (R2= 0.52, 39 d.f.)
0 0 00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0
ANTENNA LENGTH (mm)
Figure 4.1. Arriphipod body length (from telson to eye, Y) with length of 4th article 2nd antenna (X) by species and sex. Species origin: native (C. salmonis and C. spinicorne); nonindigenous (C. acherusicum and C. insidiosum) .All correlations are significant used in (P < 0.05) .Initial mean prey size and range ( I ) single-species experiments is shown in each case. 121
lengths of males are greater for C. acherusicum than for native species (Figure 4.1) .Sexual dimorphism based on antenna length is apparent in all species except C. spinicorne (Figure 4.1).
Prey Behavior
Visibility
The visibility of all amphipod species in tanks with sand (i.e., walking, swimming, or partially buried individuals) differed among species over the 24-h period (P < 0.05, ANOVA) and over the last 12-h period (P < 0.01, ANOVA). Corophiurn saimonis and C. insidiosum largely settled into the sediment during the first 6 h of the test (Table 4.1). The decreasing order of prey visibility over the entire daily period was: C. acherusicum > C. spinicorne > C. insidiosum > C. saimonis. Most visible individuals were walking (Table 4.1).
Activity
Undisturbed amphipod activity did not vary by sex or species at 14°C or 24°C (P > 0.10, ANOVA; Figure 4.2: A and B). At 14°C, activity following simulated predation attempts did not vary among females (P > 0.30, ANOVA; Figure 4.2: C) and varied only marginally among males (P = 0.05, ANOVA; Figure 4.2: C) with C. sairnonis most active and C. acherusicuni least active (94% HSD
test) . At 24°C, disturbance increased only the activity of female C. spinicorne (P < 0.01, ANOVA, 95% HSD test) .Activity of disturbed amphipods for all four species was greater at 24°C than at 14°C (males: P < 0.05, females: P = 0.01, ANOVA; Figure 4.2: C and D). Table 4.1. Walking (WA), swimming (SW) and partially visible (PV) Corophium spp. in sand substrate. Based on three observations per time interval in 10 tanks (sd = ± sample standard deviation). Rank 1 = Highest visibility.
Individuals (No./tankl2 minute observation)
Species (7:00-12:00) p.m. (3:00-9:00) a.m. (12:00-6:00) p.m.
WA SW PV Rank WA SW PV Rank WA SW PV Rank C. salrnonis mean 0.53 0.03 0.30 4 0.17 0.00 0.17 4 0.37 0.00 0.40 2 sd (±) 0.21 0.06 0.10 0.15 0.00 0.15 0.06 0.00 0.10
C. spinicorne mean 2.70 0.37 1.63 2 0.57 0.07 0.53 2 0.23 0.00 0.37 3 sd (±) 2.25 0.21 0.35 0.31 0.11 0.42 0.11 0.00 0.15
C. acherusicum mean 2.40 0.40 3.37 1 1.43 0.17 1.17 1 1.00 0.03 1.23 sd (±) 0.95 0.26 2.06 0.40 0.12 0.38 0.35 0.06 0.55
C. insidiosum mean 0.57 0.87 0.10 3 0.23 0.23 0.00 3 0.13 0.03 0.00 4 sd (±) 0.06 0.31 0.10 0.15 0.15 0.00 0.15 0.06 0.00 123
A UNDISTURBED 14°C B UNDISTURBED 24°C
12 12- MALES FEMALES
U) L() U E SAL SPI ACH INS SAL SPI ACH INS C) Ii'
C DISTURBED 14°C D DISTURBED 24°C 0 12 12- b
8 8-
4 oiLI iiI.i I SAL SPI ACH INS SAL iiSPI ACH INS
SPECIES
Figure 4.2. Mean activity (distance traveled in 5s) by 10 males and 10 females of each Corophium species held at 14°C and at 24°C (SAL = C. salmonis, SPI = C. spinicorne, ACH = C. acherusicum, INS = C. insidiosum) .with a standard error scale over each bar and significant differences denoted by letters a and b. 124
Single-Species Consumption
Pleuronectes vetulus feeding varied among species (Figure4.3: A) irrespectively of sex (P < 0.05, ANOVA; Figures4.3: B and C), with predation on C. spinicorne and C. acherusicum nearlytwice as high as on C. salmonis and C.insidiosum (95% HSD test) Significant differences in prey survival between treatmentsand controls occurred in all species except C.sairnonis (1-tailedX2 test, P < 0.05, Figure C.2) . Size of P. vetulus did not affect prey consumption (r = 0.03, p > 0.05, n = 40) Sex-selective predation was not apparent (P > 0.10, 2-tailedX2 test; Figure 4.3: B and C) .Although size-selective predation was not suggested from the Strauss' index (Figure 4.4), size selection based on size distribution of prey may occur for C. acherusicum (P < 0.01, Kolmogorov-Smirnov test; Figure 4.5).
Mixed-Species Experiments
Corophium visibility was greater in sand than in mud both 1 h and 24 h following amphipod introduction (P < 0.05, ANOVA;Table
4.2) .Except for C. insidiosum, predation was higher in sand than in mud (P < 0.001, ANOVA; Figure 4.6: A and B) .Species-selective predation occurred in both substrata (P < 0.05; ANOVA; Figure 4.6: A and B) .In sand,C. acherusicum was selected more frequently than both native species (95% HSD test) .In mud, selection was higher on both NIS (95% HSD test; Figure 4.6: B). Differences in prey survival between treatments and controls were suggested for all species except C. saimonis in sand (P < 0.05, 1-tailed X2 test) and for no species in mud (P > 0.05, 1-tailedX2 test, Figure C.3) .Sex-dependent predation was not apparent on either substratum for any species (P > 0.05, 2-tailed X2 test; sand: Figure 4.6: C and E; mud: Figure 4.6: D and F).
Substratum type had no significant effect on the size distribution of uneaten prey (Figure 4.7; p > 0.05, Kolmogorov- 125
A (MALES + FEMALES)
24 20- 16- a 12- T 8-
0 SAL SPI ACH INS
B MALES 24 20 16 12 b b 8 a
1 4 0 SAL SPI ACHINS
FEMALES
24- 20 16- 12- b 1 b 8- a 4- 1 0 SAL SPI ACH INS
SPECIES
Figure 4.3. Mean number of Corophium consumed by Pleuronectes vetulus, in single-species experiments (SAL = C. salmonis, SPI = C. spinicorne, ACH = C. acherusicum, INS = C. insidiosum), with standard error scale indicated over each bar and different letters above bars showing significant differences among species. 126
C. salmonis C.spinicome 0.10- r-0.02 P>0.90 0.1 0 r = 0.04 P >0.80 0.05- . 0.05- S . I 0.00- I S 0.00- .5 S -0.05- I -0.05-
-0.10 .0.10 02000.81012 14 0.20.40.60.81.01.21.4
C.acherusicum C. insidiosum 0.10- 0.10- r=0.37 P=0.10 r0.26 P>0.40 . 0.05- 0.05-.
0.00- 0.00-__Is
-0.05- -0.05- S
.0.10 r -0.10 02 040.60.81.012 14 0.20.40.60.81.01.21.4
ANTENNA LENGTH (mm)
Figure 4.4. Strauss' selection index by preysize (4th article 2nd antenna) and by Corophiumspecies consumed by Pleuronectes vetulus. 127
4O Corophium salmonis
Not Eaten (x = 0.42, se 3O 0.01, n = 179) Eaten (x = 0.42, so = 0.01, n = 61) 20-
10 I IFii Inn - 0.2 0304 0.5 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3
40 Corophium spinicorne Not Eaten (k = 0.64, se = 0.02, n = 57) 30 Eaten (x = 0.65, so = 0.02 n = 183)
20
10
., . InIIII1HIH LI fl Ill H 0.2 0.3 0.4 0 5 06 0.7 0.8 0.9 1.0 1.1 1.2 1.3
40 Corophium acherusicum
30 Not Eaten (* = 0.40, so = 0.03, n =57) Eaten (* = 0.60, se = 0.02 n = 183) 20
10
I. i . . . I H In 0.2 03 0.4 0.5 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3
40- Corophium insidiosum Not Eaten (= 0.34, se= 0.01, n = 151) 30- Eaten (x = 0.35, se = 0.01, n = 89) 20-
ID-
0 1-IFF I ri - 0203 0.4 05 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3
ANTENNA LENGTH (mm)
Figure 4.5. Percent of eaten and uneaten Corophium by size (4th article 2nd antenna) in 10 tanks with sand substratum.
Mean antenna length (mm) , standard error and total number of prey shown in parentheses. Table 4.2. Walking (WA), swimming (SW) and partially visible (PV) Corophium spp. in sand and mud substrata. Based on three observations per time interval in 25 tanks (sd = ± sample standard deviation).
Individuals (No./tank/2 minute observation)
Substrate (7:00-12:00) p.m. (3:00-9:00) a.m. (12:00-6:00) p.m.
WA SW PV WA SW PV WA SW PV
Sand mean 1.24 0.57 0.60 0.61 0.17 0.24 0.68 0.27 0.33 sd (±) 0.62 0.09 0.18 0.16 0.17 0.07 0.21 0.12 0.17
Mud mean 0.55 0.44 0.31 0.49 0.11 0.36 0.44 0.01 0.35 sd (±) 0.49 0.11 0.36 0.14 0.10 0.08 0.07 0.02 0.06 129
SAND MUD
A(MALES + FEMALES) B (MALES + FEMALES) b 5
I 4 3- 3- b b 2 2 .c 1-
SAL SPI ACH INS SALSPI ACH INS
C MALES D MALES 5-
4- 4- 3- 3 2 2 a b I
SAL SPI ACH INS SAL SPI ACH INS
E FEMALES F FEMALES 5- 5-
4- 4-
3- 2 a a 2 C -r bc I i- ab a rrI
SALSPIACH INS SAL SPI ACH INS
SPECIES
Figure 4.6. Mean number of Corophium consumed by Pleuronectes vetulus in mixed-species experiments (SAL = C. saimonis) , SPI = C. spinicorne, ACH = C. acherusicum, INS = C.insidiosum), with standard error scale over the bars and different letters above bars indicating significant differences among species. 130
Corophium salmonis Sand(x=0.41,se=0.01, n = 86) __ Mud = 0.43, se= 0.01, n 130) 20-
10-
I.11 . .n 0.2 0 3 0.4 05 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3
Corophium spinicorne Sand(x0.70, se0.02, n 81) Mud (x = 0.72, so = 0.01, n = 127)
>- 10-
.n H .H II In 0 .III In.nln 0 0.2 0.3 0.4 05 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3 U- o 40 Corophium acherusicum Sand (x = 0.43, se =0.03, n = 50) I- 30 z Mud (x = 0.46, se =0.02, n 95) LU 20- C) 10 LU
I H IR.H.I 0.2 03 0.4 05 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3
Corophium insidiosum Sand ( 0.34, so = 0.01, n = 75) EJ Mud (*0.36, se = 0.02, n = 93)
10-
IIrin 0.2 03 0.4 0.5 0.6 0.7 0.8 0.9 1.0 1.1 1.2 1.3
ANTENNA LENGTH (mm)
Figure 4.7. Percent of uneaten Corophium by size (4th article 2nd antenna) in sand and mud substrata. Mean antenna length (mm), standard error and total number of uneaten prey in 25 tanks per substratum are shown in parentheses. 131
Smirnov test) .Size-selective predation was not suggested, except for male C. insidiosum in sand and C. acherusicumin mud (P < 0.05, ANOVA) .Fish size seemed independent of overall prey consumption in sand (r = 0.10, P > 0.05, n = 25) and mud(r = - 0.20, P > 0.05, n = 25)
Discussion
Juvenile P. vetulus preyed on all Corophium species usedin our study. The rank of predation insingle-species experiments was not consistently related to speciesorigin as seen in mixed- species experiments. In the latter case, predation was more intense on sand than on mud and the ranking of prey selection was higher for NIS in both substrata. Thus, consumption of each species depends on the substrata and the species compositionof the prey populations. The two most consumed species in single-species experiments (C. spinicorne and C. acherusicum) were the most visible overthe 24-h observation period. Therefore, P. vetulus seems to be an opportunistic predator and predation risk varies with prey behavior and sediment type. Every other factor (e.g., prey origin, size, sex, activity, water temperature) seemed less consequential in determining potential vulnerability to predators. Our finding of significantly higher selection forboth NIS in mud is tentative due to the reduced predation in this substrate. However, fine-medium sand is the major sediment type in Yaquina Bay (Kulm and Byrne 1967). Thus, our selection experiments in sand are more representative of natural conditions than are the experiments in mud.
Increase of activity with temperature were evident for both male and female Corophium and escape responses may be faster at higher temperatures for C. spinicorne than for the other species (Figure 4.2: D) .Yet, growth is reduced at temperatures over 25°C in C. salmonis and C. spinicorne but enhanced in C.acherusicurn 132
and C. insidiosum by even higher temperatures (J.W. Chapman, in progress) .For C. spinicorne, reduced growth may thus be a trade- off for decreased predation at higher temperatures.
Morphological and/or behavioral differences between male and female Corophium did not result in sex-selective predation by juvenile P. vetuius in our laboratory experiments. Moreover, field sex selection for C. saimonis by P. vetuius was suggested at only one of the three intertidal sites examined in Yaquina Bay and Alsea Bay (Table C.1) .In the latter case,C. saimonis in benthic core samples from three sites showed an increase in the average female/male ratio with water depth from about 1.0 (0-40 cm depth) to about 1.4(80 cm depth) .Thus, a spatially heterogeneous sex ratio of prey in the environment may confound the actual prey availability.
Reimers et al.(1978) found that juvenile chinook salmon (Oncorhynchus tshawytscha) in the Sixes River estuary, Oregon, consumed more male than female Corophium. They also reported that male Corophium occurs more often on the surface of the substrate than females, but females outnumber males. Higher selection of male over female amphipods (Grandidierella lignorum) was also reported for juveniles of another fish (the sparidLithognathus lithognathus),and selection resulted from increased exposure of males to predators (Schiacher and Wooldridge 1996) .The difference in sex-selective predation between our experiments and the previous two studies can be explained in part by differences in fish foraging. Unlike 0.tshawytschaand L.lithognathus,a significant proportion of the diet of juvenile P. vetulus
comprises infauna (Castillo et al. 1996) . Thus, sex-related differences in prey accessibility may be minimized by juvenile P. vet ulus. Observation of P. vetulus feeding behavior in the present study were limited by the reduced activity of fish and the secretive behavior of the prey. The fish settled on the substratum (usually within 1 h following their introduction into 133 the tanks) and most fish remainedin a fixed area on the bottom or partially buriedin the sediment for several hours. Preliminary observations in which prey wereadded from the top of the tank showed eye movements in thesettled fish. Once the prey reached the bottom of the tank, the fishused its dorsal and anal fins to support its body and "shuffled"forward toward the prey, after which it quickly consumed the prey.Several fish attacks followed by regurgitation were needed tosuccessfully ingest large C. .spinicorne of either sex. In one case alarge C. spinicorne male oriented its second antennaesideways when confronted by a fish, a defense mechanism alsoreported by Reimers et al. (1979)
Foraging juvenile and adult P. vetulus usetheir pointed snouts as a shovel to extract infauna(Ambrose 1976; Hulberg and Oliver 1978) .Despite such behavior, the percent ofthe tank bottom area free of sand at the end of theexperiment was not correlated to total prey consumption (r = 0.17,P > 0.05, n 25) suggesting that overall sediment disturbance byfish does not enhance prey consumption. Juvenile P. vetulus seem visually oriented predators.Stomach fullness of juvenile P. vetulus increases throughoutdaylight hours in nearshore areas (Rogue and Carey1982) and estuaries (Castillo 2000) .Fish-induced turbidity in the mud treatment reduced visibility and could explain therelatively low predation on Corophium in mud.
Due to turbidity, the tank bottoms of mostmud replicates were not visible from the water surface at theend of the experiments. However, predation on amphipods was notsignificantly higher in mud replicates where the bottom was visible oncompletion of the experiment (mean consumption = 7.1 prey, n = 7fish) than in tanks in which the bottom was not visible (meanconsumption 5.8 prey, n = 18 fish, P > 0.50,ANOVA). Thus, differences in 134 amphipod behavior between sand and mud aswell as reduced prey visibility in mud could account for thegreater consumption by fish in the sand replicates.
Although the Strauss' prey sizeselection index was not significantly correlated to prey sizefor any species, C. acherusicum showed the highest correlation(Figure 4.4) .Thus, this index is partially consistentwith the significantly greater size distribution of consumed prey forthe latter species (Figure 4.5). Our experiments support fieldobservations in the sense that some NI invertebrate prey canbe highly selected by juvenileP. vetulus (Castillo 2000) .With the exception of C. insidiosum, field data for Yaquina Bay and Alsea Bayestuaries have shown that P. vetulus and starry flounder(Platichthys stellatus) prey on all Corophium speciesconsidered here (Castillo et al.1996) The low abundance of NI Corophium in thefish diet in that study may be due to the high intertidaldistribution of introduced Corophium. Although direct negativetrophic effects of NI Corophium spp. on P. vetulus are not impliedfrom our study, we can not predict the effect of anincreasing number of NIS on the abundance of P. vetulus. Yet, additionalguilds in the food-base of native predators could lead to adecline in community stability (Castillo et al. 2000) 135
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Chapter 5
Pbsence of Overall Feedback in a Benthic Estuarine Community: A System Potentially Buffered from Impacts of Biological Invasions
G.C. Castillo 1,2 H.W. Li 2 P.A. Rossignol
1 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. Department of Entomology, Oregon State University Corvallis, OR 97331.
Estuaries 23(2) :275-291 April 2000 139
Abstract
Species introductions are among the most dramatic human-induced impacts on aquatic and terrestrial ecosystems around the world. Stability patterns of an estuarine benthic community were investigated through guild interaction models representing the community before and after human-mediated species invasions. The study area was Yaquina Bay, a developed estuary on the central Oregon coast, USA, where at least 12 species of nonindigenous invertebrates have been inadvertently introduced. Three of the introduced species (the polychaetesHobsonia floridaand P.seudopolydorakempi and the cumaceanNippoleucon hinumensis)are probably among the 10 most abundant invertebrate species in the intertidal benthic community. To estimate effects and potential risks of species introductions on the native community we constructed 2 types of community models based on functional-group interactions, namely, activity guild models and trophic guild models. In both cases we observed that overall feedback has a strong tendency towards zero in pre-invasion and post-invasion models. We generated 12,000 random models of similar size and could not detect this tendency. We therefore suggest that the weak or absent overall feedback in this community may be an ecological property and not an intrinsic property of large systems as such. The reduced response to input from either invertebrate invasions or removal of native top predators, may to some extent buffer the community from such impacts. Alternative guild models suggested increased risk of stability decline in the invaded community even after accounting for potential complexity effects on stability. Thus, further species introductions in this intermediately invaded estuary should be avoided. 140
Introduction
The possible connection, whether positive or negative, between stability and complexity has profound theoretical and practical ecological implications, particularly in conservation (e.g., Goodman 1975; Pilette et al. 1990; Pimm 1991, Li et al. 1999). The generally accepted negative association implies that species invasions will destabilize native systems because negative feedback, that is, the self-regulatory capacity of a system in response to input, is reduced through increased complexity resulting from additional species interactions in the community. Levins (1975) suggested that complexity may lead to situations where input (change in birth or death rate) results in undesirable consequences, such as decline in population size, and that such systems may tend towards a value of zero in overall feedback (i.e., systems that seem to lack response to input). In contrast to negative feedback, those systems with positive feedback move away from equilibrium in response to input. The importance of the latter feedback type has being increasingly recognized within natural systems (e.g., DeAngelis et al. 1986; Stone and Weisburd 1992) .Hence, the need to consider whole system responses, and alternative feedback types seems critical for assessing the risk of human induced changes on community stability, particularly in increasingly disturbed systems such as estuaries. Estuarine and marine benthic systems are known for their complex biological interactions (Gray 1977; Miller et al. 1996) A broad range of taxonomic and trophic groups have been introduced in U.S. west coast estuaries where they are often numerically dominant (Canton 1979; Castillo 2000) .The highest number of species introductions (n = 234) has been reported for San Francisco Bay, California (Cohen and Canton 1998) .In Oregon, 60 nonindigenous species (NIS) have been reported in Coos Bay (Ruiz et al. 1997); 21 in Yaquina Bay and 13 in Alsea Bay (Castillo 2000) .Yet, no extinctions of native species have been attributed to biological invasions in estuarine or marine 141 habitats (Canton 1993) as opposed to invasions in freshwater and terrestrial systems (Office of Technology Assessment 1993) .It has not been determined whether the pervasive, human-mediated biological invasions found in these systems have reduced stability or whether communities have simply increased in complexity and maintained their original stability characteristics, whatever these may have been.
One notable aspect of marine and estuarine benthic systems seems to be their high degree of interactions other than predator-prey, particularly in sediment inhabiting organisms (e.g., Woodin 1981; Brenchley 1982; Lopez and Levinton 1987). Important types of interactions occurring in these communities include:(1) trophic group amensalism (non-competitive trophic interference of deposit feeders on suspension feeders, Rhoads and Young 1970; Wildish 1986);(2) interference competition and/or amensalism among bioturbators (i.e., burrowers and deposit feeders), tube builders and suspension feeders (Reish and Alosi 1968; Levinton 1977; Levin 1982; Wilson 1991);(3) size-dependent interference competition, arnensalism or commensalism between mobile and sedentary organisms (Brenchley 1981; Posey 1987); and (4) mutualism or commensalism through coprophagy of enriched sediments by deposit feeders (Frankenburg and Smith 1967;
Brinkhurst et al. 1972) . Possibly, these interspecific interactions, usually with positive feedback reduce the negative feedback loops associated with predator-prey interactions and altogether may cancel each other in a way that results in zero overall feedback.
We investigated the potential impact of introduced species on the overall feedback strength of the benthic community of the Yaquina Bay estuary, where at least 12 invasive species of invertebrates, including the highly abundant polychaetesHobsonia florida andPseudopolydorakempi and the cumacean Nippoleucon hinumensis have become resident. Yaquina Bay is used by many 142 juvenile fishes and crustaceans as nursery area (e.g., Pearcy and Myers 1974; De Ben et al. 1990) .Native juvenile English sole (Pleuronectes vetulus) and starry flounder (Platichthys stellatus) consume a high proportion of both native and nonindigenous (NI) invertebrates in intertidal areas (Castillo 2000), which is consistent with findings from laboratory feeding experiments on juvenile English sole (Castillo et al. In press).
Have biological invasions increased the risk of stability decline in estuaries? Could native juvenile fishes play a critical role in maintaining the overall feedback and stability characteristics of intertidal estuarine communities?Our goal was to provide answers to these questions using alternative qualitative models of trophic and activity guild interactions. We further compared the stability indices of these models with randomly generated models to determine whether the observed feedback patterns were ecological or inherent to large complex systems in general.
All models considered in this study represent the benthic intertidal community from the mid-region of Yaquina Bay, a partially mixed drowned river estuary on the central Oregon coast (Bottom et al. 1979) .Yaquina Bay has served as port since the late 19th century, having experienced both industrial and residential development. Nearly 54% of the 16 km2 surface area of the estuary is intertidal (Hamilton 1973) Documented species introductions in this estuary began in the 1870s with the importation of Atlantic oysters which also may have served as a vector for other species introductions (Carlton
1979) .We attribute further sources of species introductions in Yaquina Bay to fouling organisms on the hull of ships and to discharge of ballast water from ships, as reported for other developed estuaries (Canton and Geller 1993; Cohen and Canton 1995) 143
Methods
Data Sources
Field data from two intertidal soft-sediment flat areas were used to derive the benthic assemblage structure of the mid- section of the Yaquina Bay estuary (sites 3 and 4, located at river kilometer 12.2 and 14.9, respectively, Castillo 2000). We focused on the summer species assemblage of the benthic community as this season coincides with the highest use ofYaquina Bay by juvenile fishes (Bayer 1981; De Ben et al. 1990) .Sites 3 and 4 were selected for modeling as they included a rich invertebrate assemblage of native species and NIS which was the most representative assemblage along the estuary. The latter assemblage pattern was revealed from the species composition at six intertidal sites ranging in salinity and sediment type from
34 O and sand (site 1)to 2% and mud (site 6), river kilometer 4 to 23, respectively. Each site was surveyed four times between July and September 1993.
We derived the benthic invertebrate assemblage at the two selected sites from 390 intertidal core sediment samples collected at high and low tide along three transects (each core sample was 3.2 cm diameter and 13 cm long), each transect was parallel to the coast and about 30 m long and consisted of equally spaced sediment cores. Each of the three transects was located at an approximate high-tide water depth of 0; 40; and 80 cm in sediment mixture of sand and mud. We derived trophic relations from the diets of 127 concurrently collected juvenile fish which represented all the major benthic oriented predator species from the intertidal fish assemblage of Yaquina Bay: Pleuronectes vetulus; Platichthys stellatus and Leptocottus armatus (Castillo 2000; Table D.1) .Fish were collected in intertidal areas with a seine (32 m long x 1.8 m high and 0.8 cm stretched mesh size. 144
Model Construction
Invertebrate species were assigned to guilds, which were the model variables or vertices. We considered guilds as assemblages of species that act in a similar way in a community (e.g., Fauchald and Jumars 1979; Simberloff and Dayan 1991) .By grouping species into guilds we reduced the high number of redundant species-level interactions in the community without considerable loss of detail. Two types of models and attendant invertebrate guild classification systems were considered:(1) activity models grouping invertebrates by activity (e.g., mobility) guilds (Table 5.1; based on Posey 1987), and (2) trophic models grouping invertebrates by trophic guilds (Table 5.1; modified from Day et al. 1989) We constructed signed digraphs of alternative models based on observed predation of fishes on invertebrates. The latter were assigned to guilds based on a literature review for each of the taxa considered (Table 5.1) .Our alternative models also accounted for the previously referred ecological interactions reported in soft sediment estuarine and coastal communities
(Figure 5.1) . Besides the guilds included in our models, other estuarine components (e.g., seagrass, abiotic factors, other estuarine subsystems) were assumed to play a role in self- regulation of non-predatory guilds. The latter guilds were thus implicitly connected to other estuarine components through negative feedback (Figure 5.1: A) .Activity models included at least predation and interference competition (Figure 5.1: B and
C) . Trophic models included at least predation and exploitation competition (Figure 5.1: B and G) Our criteria for inclusion of taxa in guilds were:(1) All taxa with a mean occurrence equal or greater than 10% in both benthic samples and in the fish diet were considered representative of the community (Table 5.1); (2) A predatory benthic fish guild was present in all models;(3) Preinvaded communities only included guilds composed of native species and supraspecific taxa considered to be at least partially native; Table 5.1. Activity and trophic invertebrate guilds assigned to qualitative models of Yaquina Bay. Taxa origin: nonindigenous (t), cryptogenic (i.e., unknown origin, *), native species and supraspecific taxa (no symbol). Life Mode: F = free-living; T = tube; U = burrow. Depth range: a = surface; b = subsurface; ab = surface-subsurface. Basic activity guild codes are indicated by first letter: S (sedentary); M (mobile) .Relative sizes of activity guilds are indicated by second letter: s (small); i (intermediate); 1 (large) .The number following relative size of activity guild applies only to those guild structures in Table 5.2 accounting for differences in life mode, depth distribution and taxa origin. Trophic guild codes: Sr=r surface-deposit feeder; Sb = subsurface-deposit feeder; Su = suspension/filter feeder; Ip = predatory invertebrates. The main activity or trophic mode of each taxa is listed first when two modes are shown.
Taxa Life Vertical Guild Mode Range Activity Trophic
Bivalvia: Cryptomya californica U1 ab 1 Su 1,2 Macoma baithica U 1,33 ab Sr 3'4;Su 31 Mya arenaria 1- U 1,33 ab Su ';Sr 29
Crustacea: Neotrypaea californiensis u' ab Mll Sb 2, 6 Corophium salmonis (T;F)6 ab 6 Sil 6,7 Sr8;Su29 Corophium spinicorne (T;F)6 ab 6 Sil 6,7 Sr 6;Su 29 Eogammarus con fervicolus F 6 a6 Ml2 6 Sr' Eohaustorius estuarius u 6,35 ab Mu 9 Sr? 30 J-iarpacticoida F ab 36 Msl '° (Sb;Su) " Nippoleucon hinumensis t F 6 ab? Msl 6,12 Sr? 12,13
Folychaeta: Capitella sp. T 1,6 ab ' Mi5 14 Sb 1,15 Capitellidae (T;U) 17 ab ' Mi6 12 Sb 12 Eteone spilotus F6 ab? 17 Mi2 17 Ip? 16,17 Heteromastus filIformis t T 6 ab ' M14 19 Sb ' Hobsonia florida t T 6 a6 Si2 6,20 Sr 21 Table 5.1. Continued.
Taxa Life Vertical Guild Mode Range Activity Trophic
Manayunkia aestuarina T 6,37 a6 Ssl 6,20 Sr 22.23; Su? 17 Mediomastus californiensis T 6 b 24 Mi3 24 Sb 24 Nereis limnicola (U;F) ab1 M13 25 Sr' Paraonella platybranchia T ' a Mi4 17 Sr? 17 Pseudopolydora kempi t T 6,38 a26 Sl2 20 (Sr;Su) 26 Pygospio elegans * T 6.17 a17 Si2 20 Sr 3;Su 17 Streblospio benedicti t T 6,17 a24 Si2 6 (Sr;Su) 24 Tharyx sp. F 6,23 a17 (Mi7;Sil) 17 Sr? 17
Miscellaneous: Nemertea (F;U) ' (a;b) 32 M1327 Ip' Oligochaeta T 27;U 40;F a 39;b12 M18 28 Sb'2
Sources:1 Rudy and Hay (1991);2 Peterson (1977); Brey (1991); Tunnicliffe and Risk (1977); Posey (1987);6 Personal Observation; Higley et al.(1984); 8 Taghon (1982); DeWitt et al.(1989);'° Illg (1975); " Chandler and Fleeger (1987);12 Kozloff (1990);13 Les Watling (University of Maine, Personal communication, 1997);14 Brenchley (1982); Salen-Picard et al. (1994);26 Blake (1994); ' Fauchald and Jumars (1979);18 Strelzov (1973); ' Neira and Hopner (1993);20 Hobson and Banse (1981);21 Hentschel and Jumars (1994); 22 Lewis (1968) 23 Jumars and Fauchald (1977);24 Kalke and Montagna (1991);25 Banse and Hobson (1974);26 Taghon and Greene (1992);27 Barnes (1980);28 Cook and Brinkhurst (1975); 29 John Chapman (Hatfield Marine Science Center, OR, Personal communication, 1998);° Reichert et al. (1985);31 Brafield and Newell (1961); 32 Blake (1993); Zwarts and Wanink (1989); Hedgpeth (1975); Meador et al. (1993);36 Yingst (1978); Light (1969);38 Lindsay and Woodin (1996); Giere (1993);40 Diaz (1979) 147
Figure. 5.1. Ecological interactions between guilds 1,2 and 3 in numbered circles and attendant community matrices. F overall feedback strength. Pointed arrow and bubble arrow indicate respectively positive and negative effect to the adjacent guild. A: Guild 1 is self-regulated; B: Guild 2 preys on guild 1; C: interference competition between guilds 1 and 2 (guilds harm to each other); D: amensalism between guilds 1 and 2 (guild 1 is harmed by guild 2 but the latter is not affected by guild 1); E: mutualism between guilds 1 and 2 (guilds benefit from each other); F: commensalism between guilds 1 and 2 (guild 2 benefits from guild 1 without affecting the latter); G: exploitation competition between guilds 1 and 3 (guild 2 is a limiting resource) 148
Links = 1 Loops = 1 Fn = -1
Links =2 B Loops = 1 0 Fn = -1
C 0 -1 Links2 -1 0 Loops = 1 Fn = 1
0 -1 Links = 1 D 00 Loops =0 Fn = 0
Links =2 Loops = 1 EOJ Fn = 1
F Links = 1 Loops = 0 Fn =0
G 010 Links =5 -1-1-1 Loops =3 010 Fn = 0
Figure 5.1 149
(4) Invaded communities included native species, NIS and cryptogenic species and supraspecific taxa;(5) Guilds composed of predatory invertebrates were included in all trophic models as they are potentially important in the community (Table 5.1); (6) Each taxa was assigned to a trophic guild based on its major feeding strategy (Table 5.1); and (7) With the exception of Tharyx sp., only one mobility type was assigned to each taxa.
A total of 104 alternative community models are considered, consisting of 44 activity models and 60 trophic models. All invertebrate activity guilds were self-regulated and preyed upon by juvenile fishes. Activity models assumed that invertebrate mobility types and/or sizes and vertical distribution were important in determining invertebrate interactions. Trophic models assumed that feeding strategies determined most community interactions. Two general types of invertebrate activity guilds were defined: mobile and sedentary. Because of the importance of size in mobility-type interactions, most activity guilds were subdivided into small, intermediate and large relative sizes (after Posey 1987) .Eight types of community structure were considered in alternative activity guild models (Table 5.2) .The most complex activity guild models also accounted for differences in invertebrate life mode and their vertical distribution in the sediment (Tables 5.1 and 5.2) .Only activity guilds sharing a common depth-range (i.e., depth a and ab; or b and ab) were assumed to interact in model structures accounting for depth- range of species within guilds (Table 5.2) Trophic models included 10 types of community structures and account for the feeding mode of species within guilds (Table 5.3). Importantly, our hierarchical order of models (i.e., activity vs. trophic models, community structure and type of interactions) allows to infer the effect of presence and absence of interactions by comparing baseline models with more complex models of the same guild structure. Table 5.2. Guild structure of activity models and assumptions; number of guilds; number of alternative models and invasion status of the community for each community structure. As in Table 5.1 the basic activity guild codes are indicated by first letter: S (sedentary); M (mobile) .Relative sizes of activity guilds are indicated by second letter: s (small); i (intermediate); 1 (large) .The number following relative size of activity guild applies only to community structures AVI*; AVII and AVIII which account for additional differences in taxa origin, life mode and depth and depth-range a = surface, b = sub-surface; ab = surface- subsurface) . P represents fishpredators.Asterisks denote both non-invadedcommunities and native guild. Invaded communities and introduced guilds lack asterisks.
Community Guild No.Guilds Basic Guild Structure Alternative Structure Assumptions1 (Variables) Models
Al * A 3 *; M*; 5* 3 All A 5 *; M*; S*; M;S 5 Alli * B 6 *; Ml*; Mi*; Ms*; 5j*; Ss* 6 AIV B 10 *; Ml*; Mi*; Ms*; 5j*; Ss*; 6 Ml; Ms; Sl; Si AV B;C 11 *; Ml*; Mi*; Ms*; Si*; Ss*; 6 Ml; Ms; Sl; Si; Mi AVI* D 11 *; Mllab*; Ml2a*; Ml3ab*; Milab*; 6 Mi2ab*; Mi3b*; M14a*; Mslab*; Silab*; Ssla* AVII D 16 *; Mllab*; Ml2a*; Ml3ab*; Milab*; 6 Mi2ab*; Mi3b*; Mi4a*; Mslab*; Silab*; Ssla*; Ml4ab; Mslab; Sllab; S12a; Si2a AVIII D;E 22 *; Mllab*; M12a*; Ml3ab*; Milab*; 6 Mi2ab*; Mi3b*; Mi4a*; Mslab*; Silab*; Ssla*; Ml4ab; Mslab; Sliab; S12a; Si2a; Ml3ab; Mi5ab; Mi6ab; Mi7a; Mi8ab; Sila
'Guild assumptions: Mobility-dependent interactions. Mobility and size dependent interactions. Cryptogenic guild Mi is nonindigenous. Mobility, size and depth dependent interactions. Cryptogenic guilds: Ml3ab; Mi5ab; Mi6ab; Mi7a; Mi8ab and Sila are nonindigenous. Table 5.3. Guild structure of trophic guild models and assumptions; number of guilds and invasion status of each community structure. Trophic guild codes: Sr = surface-deposit feeder; Sb = subsurface-deposit feeder; Su = suspension/filter feeder; Ip = predatory invertebrates. Additional variables in trophic models are the food base of both surface-deposit feeders and suspension feeders (Fr) and sub-surface deposit feeders (Fb) .P represents fish predators. Asterisks denote both non-invaded community structures and native guilds as opposed to invaded community structures and introduced guilds which lack asterisks.
Community Guild No. Guilds Basic Guild Structure Structure1 Assumptions2 (Variables)
TI* A 5 5*; Ip*; Sr*; Sb*; Su* T II * B 6 5*; Ip*; Sr*; Sb*; Su*; Fr* T III C 8 5*; ip*; Sr*; Sb*; Su*; Sr; Sb; Su T IV B 9 5*; Ip*; Sr*; Sb*; Su*; Sr; Sb; Su; Fr * TV D 9 *; Ip*; Sr*; Sb*; Su*; Sr; Sb; Su; Sb * T VI B; D 10 5*; Ip*; Sr*; Sb*; Su*; Sr; Sb; Su; Fr*; Sb * T VII C; E 9 *; Ip*; Ip; Sr*; Sb*; Su*; Sr; Sb; Su T VIII B; E 10 5*; Ip*; Ip; Sr*; Sb*; Su*; Sr; Sb; Su; Fr * T IX D; E 10 5*; Ip*; Ip; Sr*; Sb*; Su*; Sr; Sb; Su; Sb * TX B; D; E 11 5*; Ip*; Ip; Sr*; Sb*; Su*; Sr; Sb; Su; Fr*; Fb*
1Each community structure represents 6 alternative models. 2Guild assumptions: Absence of exploitation competition between Sr* and Su*. Exploitation competition between Sr* and Su*. Absence of exploitation competition among invertebrate guilds. Exploitation competition between Sb and Sb*. Interference competition between Ip* and Ip. 152
Feedback Calculation
We used overall feedback strength (hereafter referred to as F0 or feedback) to compare the general tendency toward qualitative stability of the community models: F0 = [-1]' D0 where Dis the determinant of the community matrix of order n, and n is the total number of variables (i.e, guilds) in the model. The feedback in any system will range from level 0 (F0 =-1) to level n (F0) .The more negative the feedback from level 1 to n, the more stable the system. Thus, a negative or positive feedback value is a condition contributing to a system being stable or unstable, respectively (Edelstein-Keshet 1988)
IfF0= 0, the system does not respond to input because of the equal number of positive and negative feedback terms or because the system lacks loops of length n. In either case, such a system may neither return to equilibrium nor become increasingly unstable following a perturbation.
Each element of the community matrix denotes the effect to variable from variable (for and from 1 to n) .For instance, a community matrix of three variables is represented as:
a12a13 11
a22a23 21
31 a32a33
where diagonal elementsa11,a22anda33represent the respective self-effects of variables 1,2 and 3. Additional elements in row represent the effect to variable from other variables. Three possible effects to variable from variable are assigned to each matrix element: positive(a1,= 1), negative = -1) and no effect (a = 0), which respectively denote positive, negative and no effect on the instantaneous growth of variable1 due to increase in the level of variable. Such effects within the community matrix can be illustrated through links in signed 153 diagraphs (Figure 5.1) .A loop is a series of links that returns to a given variable not crossing any intermediate variable twice. The theoretical absolute value of F loops is equal to the number of loops of length n, namely, (n-l)n.
We defined models with an F range from -2 to 2 as near-zero feedback to determine the extent and consistency of reduced feedback among community and random models. We selected the previous feedback range because it consistently described the great majority of the community models and facilitated comparison with random models. The biological meaning of such a narrow near zero feedback range is that community changes driven by external forcing or internal input will be likely dampened. We evaluated feedback from the value of the coefficients of the characteristic polynomial of the qualitative community matrix which indicate the net value of feedback ioops from a given feedback level (e.g., y) ranging from 1 to n. The feedback value at level y (F) is solved for all loops of length y and/or the product of loops that have no variables in common and whose summed length is y, that is:
y = (l)' L(x,y) x1 where x is the number of loops without variables in common and L(x,y) denotes the number of x loops with y elements whose total length is y (Levins 1974) Community models were classified according to:(1) ecological interactions,(2) number of trophic levels in the case of trophic models,(3) number of guilds,(4) number of links, and (5) connectance (i.e, number of interguild interactions/maximum
number of possible interguild interactions) . Potential relations between feedback and community complexity (number of guilds; number of links; number of links/number of guilds and connectance) were investigated through Spearman's rank correlations (Devore and Peck 1986) We calculated the so-called Routh-Hurwitz criteria (hereafter R-H criteria) for stability of the community models (Edeistein- 154
Keshet 1988) .The R-H criterion 1 is met when all levels of feedback are negative. The R-H criterion 2 requires that feedback at low levels be somewhat stronger than at the higher levels (as indicated by a positive Hurwitz determinant) .Models meeting both criteria, one criterion and neither R-H criteria were considered respectively as being stable, conditionally stable and not stable. We defined risk of decreased stability resulting from human-mediated biological invasions (RSD) as:
RSD=A - B (when A> B)and RSD=0 (when A B) where A is the percent of pre-invaded community models being stable or conditionally stable and B is the percent of stable or conditionally stable invaded community models. The value of RSD ranges between 0(no risk) and 100 (highest risk)
Model Computations
Mathcad 6.0 (MathSoft, Cambridge, MA) and Natlab 4.2 (Mathworks Inc., Saddle River, NJ) were used to compute the characteristic polynomial coefficients and I-Iurwitz determinants of the community matrices (Li et al. 1999). Theoretical models (i.e., random null models) were constructed with Quatro-Pro 6.0 (Corell, Ottawa, Canada) and analyzed as above to compare possible differences between theoretical and ecological patterns in feedback as a function of the number of guilds. Statistical analyses were performed with Statgraphics Plus 2.1 (Statistical Graphics Corporation, Rockville, ND) .Theoretical models were generated through random assignment of -1;0; or +1 interactions within the community matrix. Overall, we calculated the R-H criteria for each community model and the feedback of each community model and random (null) model. 155
Results
We calculated and compared the two R-H stability criteria for the alternative guild models (Figures 5.2 and 5.3). Nearly 52% of the activity models (Table 5.4) and 73% of trophic models (Table 5.5) achieved at least conditional stability. The previous difference in percentages could be due to factors other than shorter chain length of activity models, as the latter does not seem to contribute to overall stability implied by food web theory (Pimm 1991) .Three community complexity factors (number of guilds; links and links/guild) were correlated to feedback in all combined pre-invaded community models (r 0.40; p 0.01; n= 27) and all combined invaded community models (r 0.28; p<
0.05; n=77) . Thus, stability seems to decline with complexity irrespective of the invasion status of the community. Notably, both R-H stability criter:La were met to a greater extent by models of pre-invaded communities (activity models: 13%; trophic models: 25%) when compared to models of invaded communities (activity models: 3%; trophic models: 6%)
The percent of all stable and conditionally stable activity models was greater for the pre-invaded community (73%) relative to the invaded community (38%, Table 5.4) .Thus, activity models suggested increased risk of stability decline due to biological invasions (RSD= 35%) .The previous value of community stability decline was similar for trophic models (RSD=33%), In the latter case however, 67% of all the invaded community models were stable or conditionally stable while 100% of the pre-invaded community models achieved those same conditions (Table 5.5) .The attendant risk of stability decline for all activity and trophic models combined was 29%. To determine whether the increased risk of stability decline may be just a predictable consequence of increased complexity, we compared all pre-invaded and invaded community models within a common range of variables (5 to 11 guilds) .The risk of stability decline for all activity and trophic models combined was 19%. 156
Figure 5.2. Basic guild structure of activity models for the benthic community of the Yaquina Bay estuary. Each model represents one of the eight community structures shown in Table 5.2. Outlined guild interactions are common to all models w±th the same type and number of guilds. Guild codes: P = predatory fishes; S = sedentary invertebrates (Ss, Si and Si, respectively, denote sedentary small, intermediate and large size) .N = mobile invertebrate (Ns, Ni and Ml, respectively, denote mobile small, intermediate and large size) .Vertical distribution of guilds in structures AVI*, AVII and AVIII: a = surface, b = subsurface, ab = surface and subsurface. Only guilds includingnative species have an asterisk. Derivation of the remaining 36 activity models is shown in Table D.2. 157
AIV 0 A15
0G t1
Figure 5.2 158
Figure 5.3. Basic guild structure for trophic models of the benthic community in the Yaquina Bay estuary. Each model represents one of the 10 community structures shown in Table 5.3. Outlined guild interactions for each guild structure are common to all models sharing the same type of guilds. Guild codes: P = predatory fishes; invertebrates (Sr = surface- deposit feeder; Sb = subsurface-deposit feeder; Su = suspension/filter feeder, Ip = predatory invertebrates) Additional variables are the food-base of both surface-deposit feeders and suspension feeders (Fr) and subsurface deposit feeders (Fb) . Only guilds composed of native species include an asterisk. For derivation of the remaining 50 trophic models see Table D.3. 159
Figure 5.3 160
Table 5.4. Alternative activity guild models forthe intertidal benthic community of Yaquina Bay. Ecologicalinteractions considered are: PP = predation; IC = interferencecompetition and AN = amensalism. Conn = connectance, F =overall feedback strength. R-H criteria denotes which Routh-HurwitZcriteria are met. Models with and without an asterisk representpre-invaded and invaded communities, respectively. All modelsinclude two trophic levels (Figure 5.2)
CommunityEcological R-H Model Structure Interaction Guilds Links Conn F Criteria
A1* Al PP 3 6 0.66 -2 1;2 A2* Al PP; AM 3 7 0.83 -1 1;2 A3* Al PP; IC 3 8 1.00 0 2 A4 All PP 5 12 0.40 -4 1;2 A5 All PP; AM 5 16 0.60 0 2 A6 All PP; IC 5 20 0.80 4 2 A7 All PP; IC 5 22 0.90 0 2 A8 All PP; IC 5 24 1.00 0 -- A9* Alil PP; AM; IC 6 23 0.60 0 2 A10* Aill PP; AM; IC 6 30 0.83 2 2 All* Alil PP; AM; IC 6 27 0.73 0 1 Al2* Alil PP; AM; IC 6 26 0.70 6 2 A13* AIII PP; AM;IC 6 33 0.93 0 2 A14* Alil PP; AM; IC 6 30 0.83 0 2 A15 AIV PP;AM;IC 10 53 0.49 2 2 A16 AIV PP; AM; IC 10 80 0.79 0 -- A17 AIV PP; AM; IC 10 72 0.70 0 AiB AIV PP;AM;IC 10 63 0.60 19 A19 AIV PP;AM;IC 10 90 0.90 0 A20 AIV P9; AM; IC 10 83 0.82 0 -- A21 AV PP; AM; IC 11 66 0.51 0 2 A22 AV PP; AM; IC 11 99 0.81 0 -- A23 AV PP; AN; IC 11 87 0.70 0 A24 AV PP; AM; IC 11 74 0.58 16 A25 AV PP; AM; IC 11 111 0.92 0 A26 AV PP; AN; IC 11 100 0.82 0 -- A27* AvJ: PP; AM; IC 11 50 0.36 0 2 A28* AVI 99; AM; IC 11 101 0.83 0 -- A29* AVI P9; AM; IC 11 86 0.69 0 -- A30* AV]I PP; AN; IC 11 57 0.43 14 2 A31* AVI 99; AM; IC 11 108 0.88 0 -- A32* AVE 99;AM;IC 11 92 0.74 0 -- A33 AVII PP; AM; IC 16 109 0.39 45 2 A34 AVII 99;AM;IC 16 205 0.79 0 2 A35 AVII PP; AM; IC 16 177 0.67 0 -- A36 AVII P9; AM; IC 16 131 0.48 0 A37 AVII 99;AM;IC 16 227 0.88 0 A38 AVII PP; AM; IC 16 197 0.75 0 -- A39 AVIII PP; AM; IC 22 184 0.35 0 2 A40 AVIII PP; AN; IC 22 406 0.86 0 -- A41 AVIII P9; AM; IC 22 341 0.72 0 -- A42 AVIII P9; AM; IC 22 225 0.44 -23 1 A43 AVIII 99;AM;IC 22 448 0.92 0 -- A44 AVIII PP; AM; IC 22 383 0.80 0 161
Table 5.5. Alternative trophic guild models for the intertidal benthic community of Yaquina Bay. Ecological interactions considered are: PP = predation; EC = exploitation competition; IC = interference competition; AM = amensalism; CO = commensalism and MU = mutualism. F= overall feedback strength. R-H crit indicates which Routh-Hurwitz criteria are met. Models with and without an asterisk represent pre-invaded and invaded communities, respectively. Models include three or four trophic levels (Figure 5.3)
CommunityEcological R-H Model Structure Interaction Cuilds Links Connectance Fncrit
Ti * T I PP; SC 5 17 0.70 -1i;2 T2 * T II 2?; SC 6 20 0.60 0 2 T3 T III 22; EC 8 32 0.46 -11;2 T4 T VI 2?;EC 9 37 0.47 0 2 T5 T V PP; SC 9 35 0.42 0 2 T6 T VI PP;EC 10 40 0.42 0 2 T7 * T I PP; EC; AM 5 18 0.75 -i1;2 T8 * T II 22; SC; AM 6 21 0.63 1 2 T9 T III P2; SC; AM 8 36 0.55 -1 2 T10 T IV PP;EC;AN 9 41 0.53 0 2 Til T V PP; SC;AM 9 39 0.47 0 2 T12 T VI PP;EC;AM 10 44 0.47 0 -- T13 T VII 22;EC; IC 9 46 0.56 0 2 T14 T VIII 22; SC; IC 10 51 0.53 0 2 T15 T IX 22;EC; IC 10 49 0.49 0 2 T16 T X 22; SC; IC 11 54 0.47 0 2 T17 * T I PP;EC; MU 5 19 0.80 0 2 T18 * T II 22; SC;MU 6 22 0.67 1 2 T19 T III 22;EC; MU 8 40 0.59 3 -- T20 T IV P2; SC;MU 9 45 0.58 0 T21 T V 22; EC;MU 9 43 0.53 0 T22 T VI PP; EC; MU 10 48 0.51 0 -- T23 * T I 22; SC; CO 5 18 0.75 -1 2 T24 * T II 22; EC; CO 6 21 0.63 0 2 T25 T III P2;EC; CO 8 36 0.54 -11;2 T26 T IV P2;EC; CO 9 41 0.53 0 -- T27 T V P2;EC; CO 9 39 0.47 0 T28 T VI PP; EC; CO 10 44 0.47 0 T29 * T I PP; SC; AM;MU 5 20 0.85 0 2 T30 * T II PP; EC; AM; MU 6 23 0.70 2 2 T31 T III 22; SC;AM;MU 8 44 0.68 3 2 T32 T IV 22;EC;AM;MU 9 49 0.64 0 2 T33 T V PP;EC;AM;MU 9 47 0.58 0 -- T34 T VI PP; SC; AM; MU 10 52 0.56 0 2 T35 * T I 22; SC; AM; CO 5 16 0.80 -11;2 T36 * T II PP; SC; AM; CO 6 22 0.66 1 2 T37 T III P2; SC; AM; CO 8 40 0.61 -11;2 T38 T IV P2; SC; AM; CO 9 45 0.58 0 2 T39 T V 22;EC;AM; CO 9 43 0.53 0 2 T40 T VI P2; SC; AM; CO 10 48 0.51 0 2 162
Table 5.5. Continued.
CommunityEcological R-H Model Structure Interaction Guilds Links Connectance En crit
T41 T VII PP;EC;AM; IC 9 50 0.61 0 2 T42 T VIII PP;EC;AM; IC 10 55 0.58 0 -- T43 T IX PP;EC;AM; IC 10 53 0.53 0 2 T44 T X PP;EC;AN; IC 11 58 0.51 0 -- T45 T Vu PE;EC; IC;MU 9 54 0.66 0 2 T46 T VIII PP;EC; IC;MU 10 59 0.62 0 2 T47 T IX PP;EC; IC;MU 10 57 0.58 0 -- T48 T X PP;EC; IC;MU 11 62 0.54 0 2 T49 T VII PP;EC; IC; CO 9 50 0.61 0 2 T50 T VIII PP;EC; IC; CO 10 55 0.58 0 2 T51 T IX PP;EC; IC; CO 10 53 0.53 0 2 T52 T X PP;EC; IC; CO 11 58 0.51 0 -- T53 T VII PP;EC; IC;AN; MU 9 58 0.72 0 2 T54 T VIII PP; SC; IC; AM; MU 10 63 0.66 0 -- T55 T IX PP; SC; IC; AN; MU 10 61 0.62 0 2 T56 T X PP;EC; IC;AM; MU 11 66 0.57 0 2 T57 T VII PP;EC; IC;AM; CO 9 54 0.66 0 2 T58 T VIII PP; SC; IC; AM; CO 10 59 0.62 0 -- T59 T IX P9;EC; IC;AM; CO 10 57 0.58 0 2 T60 T X 99; SC; IC; AM; CO 11 62 0.54 0 -- 163
Moreover, in the latter case the connectance of invaded community models (mean= 0.60) was not greater than that of pre-invaded models (mean= 0.71) .Hence, such stability decline may be more related to destabilizing interactions in the invaded community than to increased complexity by itself. We observed that the majority of models had a low value of feedback. Typically the models exhibited no feedback at all or were close to zero. Remarkably, the percentage of models with zero feedback was nearly the same for activity models (75.0%) and trophic models (76.6%). The meaning of such a consistent result is that either most systems had no feedback at level n or the number of positive and negative loops with a combined length n were equal and canceled out. When the criterion of zero feedback was broadened slightly to include feedback values from -2 to 2 (i.e., near-zero F), trophic models were more prevalent (96.6%) than activity models (79.5%; Figure 5.4) .Hence, activity models imply a broader range of potential stability scenarios despite the strong tendency of both types of community models toward zero feedback.
Unexpectedly, removal of the fish guild from 12 randomly selected community models (AS; A9*; A22; A26; A34; T6; Til; T25; T34; T35*; T42 and T50) caused virtually no changes in the extent to which R-H stability criteria were met in the original models. Removal of the fish guild only changed the feedback in 2 models (T25 and T35*, both becoming more negative: -6 and -3,
respectively) . The only major changes observed were to models A34 and T50 which respectively failed and passed the R-H criterion 2 when the fish guild was removed. The tendency to zero feedback may simply be a intrinsic property of a large assemblage of variables. We therefore generated random models of size ranging from 2 to 22 variables. A thousand random models were generated for each size. Models of biological communities showed a nearly flat slope for the percent of near-zero feedback models with an increase in variables (Figure 5.5; non-significant regression slope,P > 0.10) By . 164
A Activity Models
Invaded 13 Pre-inv. 22
Trophic Models 12 46 Invaded Pre-inv.
-2to2
OVERALL FEEDBACK
Figure 5.4. Distribution of feedback in models for the pre- invaded and the invaded benthic community of Yaquina Bay. A: Activity models. B: Trophic models. The number of models with feedback less than -2, between -2 and 2(near-zero feedback) and greater than 2 are indicated above each bar. 165
3 18 100- 24 12 S C') I 4fl , '.0 -J I I %11I I w % 0 I S % I o 80- % S .. 'I''% S
% S '6: 6
cDZ60- -. COMMUNITY MODELS LU
LU LLW o__
LU
zLU --RANDOM MODELS
2 4 6 8 101214 16 18 20 22 NUMBER OF VARIABLES (GUILDS)
Figure 5.5. Percent of models with near-zero feedback (-2 F1, 2) . Included are all community models of Yaquina Bay (activity and trophic models combined) and 1,000 random models per variable. Numbers of community models per variable are indicated above the dashed line. 166 contrast, random models showed a sigmoidaldecline in near-zero feedback with increasing number of variables(Figure 5.5) .The proportion of near-zero feedback models with threevariables and greater differed significantly betweencommunity models and random models (X2 test; P < 0.001)
Discussion
We confirm that as complexity of a benthicestuarine community increases due to biological invasions, stability asmeasured through feedback tends to decrease. Although thisconclusion is fully consistent with modern theoretical discussions (e.g., Goodman 1975; Li and Moyle 1981; Pimm 1984; Haydon 1994), we notice that the risk of stability decline in the invaded community persists even after accounting for the effect of increased complexity in invaded community models. We reach a major novel conclusion, namely, that two largely independent functional-group interaction models of natural systems can maintain a large majority of near-zero feedback ascommunity complexity increases. In contrast, the proportion of randomly generated models with this property rapidly declines with complexity.
Near-zero feedback in benthic systems may be a more common community property than the negative feedback suggested for planktonic systems or pre-invaded pelagic communities (e.g., Li and Moyle 1981; Lane 1986) .The role of positive feedback interactions in natural ecosystems (e.g., De Angelis et al. 1986; Stone and Weisburd 1992) and, as a result, their effects on feedback could have been greatly underestimated. In particular, the vertically compressed habitat range of benthic communities seems to favor a greater number and variety of interactions among organisms when compared to communities in the water column. Our approach is highly permissive of large complex systems. 167
A major implication of zero feedback for the present models is that a guild response to input is undetermined since no predictions, as with an inverse matrix (Bender et al. 1984), can be mathematically derived when the determinant of the community matrix is zero. The tendency towards zero feedback has potentially important theoretical and practical implications as new stabilizing and destabilizing interactions in both pre- invaded and invaded communities tend to be accommodated or buffered by virtue of the limited system response to change. The lack of relation between the percent of models exhibiting near-zero F and community complexity (shown in our study to occur over community sizes ranging from 3 to 22 guilds) tends to reconcile May's (1972) paradox of decline in community stability with complexity. Our hypothesis is not mutually exclusive with that of McCann et al. (1998) who used models of simple food webs to suggest that community persistence and stability is favored by weak to intermediate strength interactions. Conceivably, both hypothesized factors could operate concurrently, particularly in highly complex communities.
Although no single model may fully represent all species interactions in a community as complex as the mid section of Yaquina Bay, our hierarchical modeling approach provided consistent stability pattern based on alternative and realistic community interactions. For instance, our trophic models consider interference competition, mutualism, amensalisin and cornmensalisrn, relationships that are seldom included in food web models (Hall and Raffaelli 1993) .Yet, our models sacrifice precision for greater realism and generality, three known model attributes which cannot be simultaneously optimized (Puccia and Levins
1985) .Moreover, our alternative activity and trophic guild classifications reveal the limitation of representing all relevant interactions within a given community using a single functional-group approach.
We estimate that the most ecologically realistic models have the most complex guild structure (i.e., A VII; A VIII and T X for 168 invaded communities and A VIand T 11* for preinvaded communities) .Although the least complex models more frequently met both R-H criteria, they generally had similar feedback patterns than more complex models (Tables 5.4 and 5.5; Figure 5.5).
Our results suggest that extreme resistance is the primary property of the studied intertidal benthic community in the Yaquina Bay estuary. Only 9.6% of all the activity and trophic models combined had positive feedback greater than 2 and most of them corresponded to invaded communities (Figure 5.4) .Species may be added to a system and such system will stillprovide great resistance due to its capacity to buffer changes in feedback. This property may be a result of selection acting on systems that are constantly changing both in the short-term (e.g., tidal exchange, wind-driven water and/or sediment transport and seasonality), or in the long-term (e.g., droughts, floods, El Niflo events) .Paradoxically, these communities may undergo a tremendous amount of challenges, either due to invasions or to "input" to birth and death rates of native species, and yet not react as a self-regulated system due to their implied lack of feedback.
Fish can be strong interactors, "keystone species", top-down or intermediate predators of importance in food webs (e.g., Deegan and Thompson 1985; Gerking 1994) .Yet, our models suggested that juvenile fish may not generally act as keystone predators for the intertidal benthic system of Yaquina Bay. The impact of keystone predators on system stability may simply be overwhelmed by the high level community interactions arising from the "stroback" invertebrate subsystem (i.e., the remaining community matrix after removing of the fish guild) .Our theoretical prediction, although counterintuitive, is fully consistent with numerous unexpected experimental outcomes from marine soft sediments reviewed by Peterson (1979) .Two interesting patterns in these previous studies are:(1) No obvious decline in species richness even after considerable 169 periods free from predation and (2) No species became dominant following predator removal. Despite the apparent lack of competitive exclusion in predator- free soft sediment communities (Peterson 1979), fish could be important in controlling the abundance of species in subsystems of a community. For example, a model composed of two self- regulated competing guilds and a common predator changes from zero feedback to positive feedback following the removal of the predator. Thus, native juvenile fish in Yaquina Bay to some extent may help to control the abundance of certain NIS.
Species with facultative feeding modes assigned to our trophic guild models do not undermine the value of our functional-group approach for representing important community level interactions because:(1) Variable flow conditions should still allow each species to exhibit its dominant feeding mode over the tidal cycle, (2) Most guilds were represented by several taxa with the same dominant feeding strategy. Moreover, the generality of our trophic functional group classification allows inclusion of a wide range of taxa which otherwise could not be consistently classified under more taxonomically restricted functional-groups (e.g., Fauchald and Jumars 1979), or under even more conditional groups based on particular flows in the environment (e.g., Miller et al. 1992).
Buffering mechanisms other than those resulting from guild interactions have also been implied for estuarine communities. Levinton (1972) suggested that deposit feeders should be less exposed to fluctuations in abundance than suspension feeders. He pointed out that only deposit feeders make use of organic matter within the sediment as a "sink" which would serve as a buffer against fluctuations in food supply. Support for a temporally stable structure of deposit-feeder communities was also provided by Kendall (1979) .Besides, Ott and Fedra (1977) suggested that unlike tropical marine and high latitude waters, a biomass storage" in productive estuaries and shallow temperate seas would serve as stabilizing mechanisms in these fliictuating 170 ecosystems. Thus, several buffering mechanisms may simultaneously contribute to the maintenance of estuarine communities.
Using Gershgorin disks (a measure of e.igenvalue distribution) and applying feedback as the criterion for stability, Haydon's (1994) simulation analysis of the stability-complexity paradox suggested that stability increased with connectivity (i.e., connectance) but declined with the number of variables. The latter conclusion is consistent with the positive association between feedback and number of guilds in all our pre-invaded and invaded community models. Yet, increased connectance seemed associated to increased stability (i.e., negative feedback) only in activity models (r = -0.34, P < 0.05)
Our results arise from analytical theoretical considerations and may not occur under quantitative analyses. Thus, instead of canceling out, positive or negative loops may overwhelm the other and destabilize or stabilize the system, respectively. The problems with this objection are two-fold: First is that quantitative community matrices are unknown for most if not all natural complex systems. Reliable measurement of all the relevant quantitative relations among variables in ecological systems is often not possible in practice. Consider a n x n community matrix with a maximum of three interactions (-1;
0; 1) . Even in a system with three variables such model boundaries result in a total of 3' 39qualitative configurations. Second is that the range of values permissible for stability becomes extremely small with increasing system size (Nay 1972) and the paradox between stability and diversity quickly strikes out this possibility (Goodman 1975; Pimm 1984) .Under our consideration, feedback values of the large loops that contribute to feedback need only be approximately equal for the system to be resistant. Such a system may not be stable in that it does not respond to disturbances. 171
Our hypothesized near-zero feedback patterns are not to be interpreted as implying that invasions in estuaries are not risky. Clearly, invasions are to be avoided for two reasons. First is that the risk of stability decline will be increased along with the risk of extinction. Unlike systems with negative or positive feedback, if a system is already at zero feedback, an introduction will have unpredictable effects on individual species as it will not likely trigger a system-level response, but it will likely change the density of native populations or possibly replace some. Second, Yaquina Bay is in an intermediate stage of invasion relative to other estuaries on the west coast of the United States. The possibility is always present that an unusual organism will not "fit in" by interacting at the quantitative levels implied in our analysis and dramatically change the system. For example, invading organisms such as crabs and bivalves and sea grass (Cohen and Carlton 1995) and fishes (Moyle 1986) have caused major ecological changes that can greatly exceed the resistance and resilience of some estuarine communities. Nevertheless, reduced or absent feedback may be one of the important factors that explains why some communities appear to be more resistant to impact. 172
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Chapter 6
Conclusions
Summary
Both benthic macroinvertebrate and fish communities from Alsea Bay and Yaquina Bay are in intermediate stages of biological invasion when compared to larger and more urbanized U.S. west coast estuaries such as San Francisco Bay, Coos Bay and Puget Sound. Conservation of intertidal and subtidal habitats of native fishes and invertebrates in Alsea Bay and Yaquina Bay are then critical for buffering the community from increasing risks of further biological invasions due to human-mediated invasions from both regional areas and from more distant donor areas.
Although native macroinvertebrate species as a group still dominate in most intertidal areas of Alsea Bay and Yaquina Bay, many introduced benthic invertebrates have become important prey for native benthic estuarine-dependent fishes. Estimates of prey selection by native juvenile fishes (English sole and starry flounder) and the overall number of native species and NIS (nonindigenous species) in the environment and in the fish diet suggested no direct adverse trophic effects of biological invasions on juvenile fishes. However, whether the increased role of NI (nonindigenous) prey in the fish diet has resulted in enhanced total prey production at the expense of native prey production is uncertain. Nevertheless, models of functional-group interactions for the intertidal benthic community of Yaquina Bay suggests that further species introductions should be avoided.
Alsea Bay and Yaquina Bay have been invaded by benthic macroinvertebrates which comprise mainly polychaetes, crustaceans, and bivalves. The most likely mechanism of species introductions in the Alsea Bay and Yaquina Bay estuaries is culture of introduced Atlantic oyster (Crassostrea virginica, 181 native to the east coast of the U.S.) and Pacific oyster (C. gigas, native to the western Pacific) .Ballast water discharge may have been a secondary source of benthic macroinvertebrate invasions in Yaquina Bay. The previous conclusion is consistent with the small difference in numbers of NIS of macroinvertebrates between Alsea Bay (n = 8) and Yaquina Bay (n = 11) despite the absence of ballast water traffic in Alsea Bay.
More native macrobenthic invertebrate species are found in intertidal areas of Yaquina Bay (native = 47) when compared to Alsea Bay (n = 33) and all NIS found in Alsea Bay are present in Yaquina Bay. The total density of NI macrobenthic invertebrates in sediment samples is also higher in Yaquina Bay, both at high- and low-tide. The polychaetes Hobsonia florida and Pseudopolydora kerripi and the cumacean Nippoleucon hinuniensis are among the 10 invertebrate taxa with greatest density in Yaquina Bay both at low- and high-tide. In Alsea Bay only one NIS (H. florida) was found among the 10 dominant taxa at high-tide.
All species from Alsea Bay in beach-seine samples are native (15 fishes and four decapod crustaceans) .Among the 20 species of fishes found in Yaquina Bay, only two are NIS (Alosa sapidissima and Lucania parva) and all six decapod crustaceans are native. Total CPUE of fishes in beach-seine samples from either Alsea Bay or Yaquina Bay are at least 10 times greater when compared to CPUE of decapod crustaceans in the same samples. In term of species density and their distributions along estuaries, most similar taxa including native species and NIS are not distributed in the same assemblages. Hence, noncoevolved interactions among similar taxa may not be more likely in comparison to interactions among more distantly related taxa.
The highest intertidal total densities of NI invertebrates in sediments under low and high tide conditions occurred at: 1) high-mid temperatures in both estuaries; 2) mid salinities in Alsea Bay and 3) mid-low salinities in Yaquina Bay. Total 182 densities of NI invertebrates were higher than those of native invertebrates only in Yaquina Bay, and such dominance coincided with high temperatures and mid salinities. Native invertebrates dominate in species richness over NIS under all temperature- salinity combination in both estuaries.
Variations in species densities of invertebrates in sediment samples and in CPUE of fishes and decapod crustaceans in seine samples are mainly accounted for by water temperature, salinity and macrophyte density. High values for the latter three environmental factors are associated with greater densities for most NI invertebrates at high-tide. The latter habitat conditions also coincided with the distributional centers of most native fishes, decapods and the introduced American shad. Yet, the CPUE of fishes (either all species or benthic species) and decapods crustaceans along the estuaries are not correlated with invertebrate densities in sediment samples, irrespective of the species' origin.
Native invertebrates are the dominant prey items for juvenile English sole and starry flounder in Alsea Bay, both in number and volume of prey, but in Yaquina Bay native and NI invertebrate prey are equal in importance. In term of prey richness however, the ratio of NI to native invertebrate prey for flatfish in Yaquina Bay (8/25) was similar to that in Alsea Bay (5/20) The similar ratios of native species to NIS in the fish diet and their benthic habitats indicate that native fishes do not distinguish between these two prey types.
Major NI prey of English sole and starry flounder varied greatly among sites within each estuary in terms of overall item contribution. The polychaetePseudopolydora kempi,the bivalve Mya arenaria and the cumacean Nippoieuconhinuniensisare common NI prey in Alsea Bay and Yaquina Bay. The NI polychaete Streblospiobenedictiwas a major prey in Yaquina Bay but it was 183 not detected in Alsea Bay. Majornative prey for both flatfishin the two estuaries included theamphipod Corophium salrnonis and the bivalve Macomabalthica.
Similar selection for native and NI preyis suggested by juvenile English sole and starry flounderin Alsea Bay and Yaquina Bay. Hence, predator-preycoevolution at species-level is not a critical determinant of preyselection by these estuarine- dependent fishes. Juvenile English soleand starry flounder are generalist predators as indicated by:1) their reliance on a wide variety of common native and NI prey; 2) thesimilar proportions of prey in gut contents and sediments,suggesting little selectivity of prey ;and 3) the non-significant difference between interspecific and intraspecificdiet overlap in prey volume between both flatfish species. Unlike starry flounder, English sole hadlower number and volume of prey during morning hours (low tide)when compared to afternoon hours (high tide) .Such increasing prey volume throughout the day is consistent with a morevisually oriented predation mode by age-O English sole. Prey usage by juvenile English soleand starry flounder was correlated with the availability of individualnative species and NIS. However, the total number of preyin the fish diet was not correlated with the total prey density amongestuarine sites.
Laboratory prey selection experimentsin which juvenile English sole were exposed to equal numbers ofnative amphipods (Corophium salmonis and C. spinicorne) and NorthwestAtlantic amphipods (C. acherusicum and C. insidiosum)support field results in that prey origin was not a criticaldeterminant of prey selection. Thus, NIarnphipods are potentially capable of supporting higher trophic levels of nativespecies.
Prey consumption by English solein single-species experiments over sand substratum was greater onthe native C. spinicorne (native species) and C. acherusicum (NIS)than on C. insidiosum 184
(NIS) and C. salmonis (native species) .Prey selection in mixed species experiments was consistentlyhigher for NI prey than for native prey over mud substratumbut not over sand substratum. No sex-selective predation byEnglish sole occurred on any species in either substratum typewhich is further supported by field observations on English sole and byits generalist feeding on both epifauna and infauna.Prey size-selection was only suggested for C. acherusicum in both substratatypes. Interspecific prey selection may varywith visibility; substratum type and prey behavior. In thelatter case, the higher exposure of C. spinicorne and C. acherusicum oversand substratum in single-species experiments increasedpredation. The greater predation on allCorophiurnspp. over sand substratumthan over mud substratum is potentially due tohigher prey visibility on sand. Thus, predation risk increasedwith prey exposure and visibility. Other prey characteristics (e.g.,origin; size; sex; activity) were less consequential.
Both activity and trophic guild modelsof the benthic community of Yaquina Bay indicateincreased risk of stability decline following species invasions.Such patterns were detected even after accounting forpotential complexity effects on stability. Nevertheless, community modelssuggested reduced response of the benthiccommunity of Yaquina Bay to input from either invertebrate invasions or potentialremoval of native fish predators. Such implied resistance may to someextent buffer the community from impacts of introduced species. Although juvenile fishes may not generally act askeystone predators in the intertidal benthic systemof Yaquina Bay, they may help to control theabundance of certain NIS. Activityand trophic models revealed a strong tendencyof overall feedback towards zero in both pre-invasion andpost-invasion scenarios. The implied weak or absent overallfeedback in the benthic community of Yaquina Bay may be anecological property and not an intrinsic property of large systems assuch. 185
Recommendations for Future Research
Suggested areas for continued work on the ecology of estuarine invasions include field, laboratory and modeling work on short- and long-term effects of noncoevoiLved species interactions on reproduction, growth, survival and recruitment of native species and NIS. Such research should complement long-term prevention and monitoring programs to minimize species invasions in estuaries and coastal areas.
Basic information is needed on key factors controlling the survival and production of major native and NI prey items for estuarine-dependent fishes and the nutritional value and prey conversion efficiency under several temperature-salinity combinations. This research is critical to predict potential interannual changes in prey abundance and long-term effects of climate change on diversity of native species and NIS.
More research on the extent of biological invasions in planktonic and nectonic communities is required, including determination of densities and functional roles of species (e.g., their trophic and non-trophic interactions) .Such knowledge is critical to evaluate the interrelations of organisms between the water column and the benthos and to improve understanding of the connections between these systems.
The systematics, genetics and ecology of cryptogenic species must be investigated to facilitate resolution of their origins. Further research needs are determination of species composition and density of meiobenthos, particularly harpacticoid copepods and assessing the role of different prey species on early stages of benthic juvenile fishes. 186
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Appendices 209
Appendix A
Complement of Chapter 2 Table A.l. Summer mean density and overall occurrence of intertidal invertebrates in sediment samples from Alsea Bay. Six high-tide sites (Al to A6) and three low-tide sites (al, a3 and a6) are included. Probable species. origin: nonindigenous (*), and cryptogenic (**).Noasterisk is indicated for native species and supraspecific taxa.
Mean Density (No.m2) Overall Taxa Al A2 A3 A4 A5 A6 al a3 a6 all sites occur.
Annelida Polychaeta: Abarenicola pacifica 21 35 14 8 33 Armandia brevis 14 2 11 Boccardiaproboscidea 28 3 11 Capitella sp. 1167 1492 173 41 35 656 235 7 423 89 Capitellidae(fragment) 41 62 7 35 28 19 56 Eteone (fragment) 7 1 11 Eteone columbiensis 14 2 11 Eteone spilotus 269 152 76 41 14 28 111 77 78 Fabricia sp. 14 2 11 Glycindearmigera 7 1 11 Glycindepolygnatha 14 14 214 21 29 44 Heteromastusfiliformis * 14 7 2 22 Hobsonia florida * 332 684 55 1913 2666 442 62 41 629 758 100 Leithoscoloplos pugettensis 7 1 11 Manayunkia aestuarina 28 7 4 22 Mediomastuscaliforniensis 111 55 69 857 145 137 56 Neomediomastussp. 14 2 11 Nephtyscaeca 14 2 11 Nereislimnicola 41 14 601 1091 677 83 1105 401 78 Paraonellaplatybranchia 7 1 11 Polychaeta (fragment) 48 35 41 41 14 35 24 67 Polydora cornuta * 69 7 8 22 Prionospio sp. 14 2 11 Table A.1. Continued.
Mean Density (No.m2) Overall Taxa Al A2 A3 A4 A5 A6 al a3 a6 all sites occur.
* Pseudopolydorakempi 207 1699 14 14 97 256 254 67 Pygosplo elegans ** 20841 2528 608 560 2052 863 3050 67 Scolelepis sp. 28 7 21 6 33 Oligochaeta 3820 2908 421 1078 1993 242 4255 269 159 1683 100 Arthropoda Crustacea Amphipoda: Aliorchestes angusta 14 2 11 Ampithoe lacertosa 41 14 6 22 Ampithoe valida * 14 28 5 22 Corophium salmonis 4704 5892 3150 12013 8462 8186 1768 2390 6687 5917 100 Corophium spinicorne 7 180 83 511 28 83 99 67 spinosus * Eobrolgus 28 3 11 Eogammarus confervicolus 138 28 28 14 83 69 28 83 52 89 Eohaustorius estuarius 14 28 21 7 33 Brachyura: Hemigrapsusoregonensis 14 14 14 14 6 44 Copepoda: Harpacticoida 55 14 14 10 111 7 35 48 33 89 Hemicyciops subadhaerens ** 28 55 28 124 26 44 Cumacea: Cumelia vulgaris 62 41 41 14 235 131 58 67 Nippoleucon hinumensis * 7 193 14 28 28 152 90 57 78 Vaunthornpsonia sp. 7 1 11 Isopoda: Gnorismosphaeroma insulare 14 14 3 22 Table A.l. Continued.
Mean Density (No.m2) Overall Taxa Al A2 A3 A4 A5 A6 al a3 a6 all sites occur.
Macrura: Neotrypaea californiensis 83 28 90 117 35 44 Upogebia pugeftensis 14 7 2 22 Ostracoda 14 21 4 22 Tanaidacea: dubia ** Leptochelia 939 7 105 22 Sinelobus stanfordi ** 7 1 11
Ins ecta: Chironomidae 83 14 31 566 14 7 760 164 78 Diptera 28 7 4 22 Insect (fragment) 14 2 11 Mollusca Bivalvia: Bivalve (fragment) 14 14 3 22 Clinocardiuin nuttalli 14 2 11 Cryptomya californica 41 21 318 111 55 44 Macama baithica 76 69 546 318 166 28 221 207 35 185 100 Macoma inquinat:a 55 6 11 Mya arenaria * 7 14 111 14 41 62 28 67 Myselia tumida 35 4 11 Mytilus californianus 14 2 11 Mytilusedulis ** 28 3 11 Gastropoda: Alderia modesta 48 5 11 Aplysiopsisenteroinorphae 28 28 6 22 Table A.l. Continued.
Mean Density (No.m2) Overall Taxa Al A2 A3 A4 A5 A6 al a3 a6 all sites occur.
Prosobranchia 21 2 11
Nematoda 90 14 14 21 14 17 56
Nemertea 62 28 152 76 124 14 62 35 14 63 100 Phoronida: Phoronis pailida 7 14 14 2217 166 269 56 Platyhelminthes Prolecithophora 35 14 69 14 55 69 28 124 45 89 Table A.2. Summer mean density and overall occurrence of intertidal invertebrates in sediment samples from Yaquina Bay. Included are six high-tide sites (Yl to Y6) and three low-tide sites (y2, y4 and y6). Probable species origin: nonindigenous(*), cryptogenic (**). No asterisk is indicated for supraspecific taxa. Samples were collected on July 5 and 18, August 3 and September 17, 1993.
Mean Density (No.rrr2) Overall Taxa Y1 Y2 Y3 Y4 Y5 Y6 y2 y4 y6 all sites occur. Annelida Polychaeta: Abarenicolapacifica 7 1 11 Ainaenaoccidentalis 173 19 11 Arinandiabrevis 14 14 3 22 Boccardia proboscidea 2715 477 449 404 33 Branchiomaldanesp. 28 3 11 Capitella sp. 159 988 283 338 829 753 193 394 78 Capitellidae (fragment) 14 338 62 179 66 44 Dorvillea annulata 21 14 4 22 Eteone (fragment) 7 1 11 Eteone californica ** 7 1 11 Eteone spilotus 7 138 173 297 14 14 28 83 84 89 Eupolymniaheterobranchia 28 3 11 Exogone lourei 14 2 11 Glycera americana 14 2 11 Glycinde polygnatha 35 28 14 83 18 44 Heteromastusfiliformis * 21 28 124 55 7 262 55 67 Hobsonia florida * 14 1161 1534 4673 48 7723 1684 67 Leitoscoloplospugettensis 28 28 28 21 12 44 Lumbrinerissp. 7 1 11 Lumbrineriszonata 55 6 11 Magelona hobsonae 28 3 11 Manayunkia aestuarina 470 2114 273 138 3523 724 56 Mediomastuscaliforniensis 1458 1589 242 41 14 421 14 55 426 89 Table A.2. Continued.
Mean Density (No.m2) Overall Taxa Yl Y2 Y3 Y4 Y5 Y6 y2 y4 y6 all sites occur.
Mesospio sp. 14 2 11 Myriochele sp. 14 2 11 NephLys caeca 14 14 3 22 Nereis limnicola 262 117 193 539 14 235 151 67 Orbinia sp. 14 2 11 Owenia fusiformis ** 62 7 11 Paraoneiia platybranchia 62 90 83 26 33 Phyllodoce hartmanae 41 5 11 Platynereis bicanaliculata 14 14 3 22 Polychaeta (fragment) 152 318 166 408 14 55 55 159 76 156 100 Polydora cornuia * 28 55 28 28 15 44 Prionospio sp. 14 2 11 Pseudopolydora kempi * 131 138 449 995 83 62 1637 388 78 * Pseudopolydora paucibranchiata 14 2 11 Pygospio californica 14 2 11 Pygospio elegans ** 1167 373 311 35 470 366 117 315 78 Rhynchospio sp.l 62 7 11 Rhynchospio sp.2 1292 7 144 22 Scolelepis (fragment) 14 2 11 Scolelepis sp. 14 14 3 22 Sphaerosyllis californiensis 28 3 11 Spionidae (fragment) 14 2 11 Streblospio benedicti * 14 2024 1444 3523 69 276 1520 14 987 89 Tharyx sp. 7 207 428 318 7 108 56 Oligochaeta 2432 6625 2577 290 3053 2318 3495 387 1340 2502 100 Table A.2. Continued.
Mean Density (No.rrr2) Overall Taxa Yl Y2 Y3 Y4 Y5 YE y2 y4 y6 all sites occur.
Arthropoda Crustacea Pmphipoda: Allorchestes angusta 891 99 11 Ampithoe lacertosa 124 14 15 22 Ampithoe valida * 207 373 14 66 33 Caprellidea 14 2 11 Corophium acherusicurn * 166 28 221 28 49 44 Corophium brevis 2590 288 11 Corophium salmonis 373 5105 3917 7751 677 5281 3033 3281 8883 4256 100 Corophium spinicorne 7 304 1402 511 41 421 97 145 1478 490 100 Eobrolgus spinosus * 1133 14 14 129 33 Eogaminarus confervicolus 14 283 262 14 21 21 68 67 Eogammarus sp. 14 2 11 Eohaustorius estuarius 649 21 55 81 33 Gammaridea 14 2 11 Traskorchestia traskiana 28 3 11 Bra chyura: Cancer magister 14 2 11 Hemigrapsus oregonensis 14 21 4 22 Copepoda: Calanoidea 21 14 4 22 Harpacticoida 21 21 48 90 28 23 56 Hemicyclops subadhaerens ** 35 83 7 14 28 41 23 67 Cumacea: Cumelia vuigaris 90 35 28 41 83 31 56 Nippoleucon hinuxnensis * 14 262 387 808 262 48 525 698 14 335 100 Table A.2. Continued.
Mean Density (No.nr2) Overall Taxa Yl Y2 Y3 Y4 Y5 Y6 y2 y4 y6 all sites occur.
Isopoda: Gnorisrnosphaeroma insulare 14 14 14 5 33 Gnorismosphaeroma oregonensis 14 7 7 3 33 Lironeca californica 7 1 11 Macrura: Neotrypaea californiensis 117 90 41 21 7 28 34 67 Upogebia pugettensis 7 1 11 Tanaidacea: Leptochelia dubia ** 13470 62 7 28 1507 44 Sinelobus stanfordi ** 55 69 7 15 33 Ostracoda 21 41 14 14 41 7 15 67 Insecta: Aphididae 14 5 11 Chironomidae 14 21 138 41 28 21 55 35 78 Collembola 28 3 11 Diptera 7 1 11 HesperoconopaSp. 14 5 11 Hymenoptera 55 6 11 Acarina 55 6 11 Aranea 55 6 11 Mollusca Bivalvia: Clinocardium nuttaili 76 14 41 15 33 Cryptomya californica 297 28 221 61 33 Macoma baithica 76 256 394 718 276 41 193 401 55 268 100 Mya arenaria * 28 166 138 111 21 235 78 67 Myseila tumida 242 27 11 Table A.2. Continued.
Mean Density (No.m2) Overall Taxa Yl Y2 Y3 Y4 Y5 Y6 y2 y4 y6 all sites occur.
Mytilus edulis ** 221 28 76 36 33 Transenella tantilla 256 28 11 Bivalve (fragment) 41 14 6 22 Gas tropoda: Aicieria modesta 41 28 97 18 33 Aplysiopsis enteromorphae 14 41 14 8 33 Gastropoda (fragment) 28 3 11 Littorina sitkana 14 2 11 Melanochiamys diomedea 41 21 7 22
Nematoda 1851 7 21 21 414 55 69 7 272 89
Nemertea 159 117 111 48 338 62 93 67 Phoronida Phoronis pailida 28 857 359 14 140 44 Platyhelminthes Prolecithophora 7 7 55 83 14 180 38 67 Turbellaria 14 7 2 22 Table A.3. Summer mean catch per unit effort (CPtJE) and occurrence of fishes and decapods in Alsea Bay as determined from seine sampling. Six intertidal sites (Al to A6) and three subtidal sites (al, a3 and a6) are included. Species are ordered in decreasing mean CPUE for all sites combined. All species are native (D = decapod crustaceans)
Mean CPUE (No.1000m2)
Species Al A2 A3 A4 A5 A6 al a3 a6 All Sites Occurr.
Engraulis inordax 1 1 36108 38 1 2 4016.7 67 Cymaogaster aggregata 259 931 6323 1470 2663 636 22 5196 7853 2817.1 100 Leptocottus armatus 20 57 27 137 730 64 1176 58 1331 400.1 100 Gasterosteus aculeatus 3 34 403 69 35 378 102.4 67 Oncorhynchustshawytscha 32 85 20 22 43 57 112 36 182 65.6 100 Crangon franciscorum (D) 44 221 130 106 55.6 44 Pleuronectes vetulus 6 1 24 201 9 26.8 56 Platichthysstellatus 4 2 16 9 11 43 9.5 67 Cancer magister (D) 37 4.1 11 Pholis ornata 8 2 2 2 14 4 3.6 67 Hypomesus pretiosus 16 1 3 1 1 2.5 56 Hemigrapsus oregonensis (D) 5 5 3 1 3 2 2.2 67 Cottus asper 5 1 0.6 22 Clevelandia 0.5 11 Oligocottus maculosus 1 2 0.3 22 Atherinops affinis 2 0.2 11 Oncorhynchuskisutch 2 0.2 11 Ciupea pailasii 1 0.1 11 Heptacarpus paludicola (D) 1 0.1 11 Table A.4. Summer mean catch per unit effort (CPUE) andoccurrence of fishes and decapods in Yaquina Bay as determined from seine sampling. Six intertidal sites (Ylto Y6) and three subtidal sites (y2, y4 and y6) are considered. Species are ordered in decreasingmean CPUE for all sites combined. Species code: nonindigenous species (*), decapod crustaceans (D).
Mean CPUE (No.1000 m) Species Yl Y2 Y3 Y4 Y5 Y6 y2 y4 y6 All Sites Occurr.
Engraulis mordax 3 669 617 107824415 291 2 3008.3 78 Cyrnatogaster aggregata 1599 637 507 2149 1690 1381 778 404 2302 1272.0 100 Leptocottus armatus 100 90 647 3740 322 459 205 80 81 636.0 100 Crangon franciscorum (D) 2 15 59 28 5 320 40 20 54.4 89 Cancer magister (0) 9 3 69 37 2 63 4 20.8 78 Gasterosteus aculeatus 1 1 4 48 2 1 104 18.0 78 Clupea pailasii 22 62 67 16.9 33 Atherinops affinis 49 2 8 6 6 39 27 15.2 78 Piatichthys steilatus 4 5 1 21 43 44 6 7 14.7 89 Oncorhynchus tshawytscha 24 1 2 8 9 63 2 2 12.3 89 Pleuronectes vetulus 35 3 20 20 25 6 12.2 67 Pholis ornata 11 1 7 13 3.6 44 Lepidogobius lepidus 4 11 1.7 22 Hemigrapsus oregonensis (D) 3 1 3 2 6 1.7 56 Alosa .sapidissiina * 1 1 2 10 1.6 44 Hyperprosopon argenteuni 3 10 1.5 22 Syngnathus leptorhynchus 8 2 1.1 22 Phanerodon furcatus 4 2 0.7 22 Hypomesus pretiosus 2 2 1 0.6 33 Cancer productus (D) 4 0.5 11 Pholis schuitzi 1 2 0.3 22 Aulorhynchus fiavidus 2 0.2 11 Heptacarpus paludicola (0) 2 0.2 11 Puggetia producta (0) 2 0.2 11 Lucania parva * 2 0.2 11 Luinpenus sagitta 2 0.2 11 221
Table A.5. Life-mode and functional-groups of nonindigenous invertebrates found in intertidal and subtidal areas of Alsea and Yaquina Bay. Life modes (U = burrow, T = tube, F = free-living) Activity functional-groups (Si = sedentary large size, Si = sedentary intermediate size, Ml = mobile large size, Mi = mobile intermediate size, Ms = mobile small size) .Trophic functional- groups (Su = suspension feeder, Sr = surface-deposit feeder, H = herbivore, D = detritivore)
FunctionalGroup
Taxa Life Mode Activity Trophic
Bivalvia:
Mya arenaria U' Si' Su Sr Crustacea: Ampithoe valida T3 Si '' H 5;D 6 Corophium acherusicum T7 Si Su 7;Sr Eobrolgus spinosus F9 Mi Sr?2 Nippoleucon hinumensis F7 Ms " Sr?9,10
Poiychaeta: Heteromastus filiformis T7 Ml" Sb" Hobsonia florida T Si 7,12 Sr " Polydora cornuta T ' Si Su 14; Sr 14
Pseudopolydora kempi T , 15 Si 12 Sr 11;Su15 17 Pseudopolydora paucibranchiata T 16 Si Su 17; Sr
Streblospio benedicti T7,18 Si Sr 19; Su 19
Sources:1 Rudy and Hay (1991);2 J.W. Chapman (personal communication 1998); Borowsky (1983); Duffy and Hay (1994); Duffy (1990);6 Nelson (1979); ' G.C. Castillo (personal observation);8 Bousfield (1973); Kozioff (1990); '° L. Watling (personal communication 1997); Neira and Hopner (1993); 12Hobson and Banse (1981);13 Hentschel and Jumars (1994); 14 Dauer et al. (1981);" Taghon and Greene (1992); 16Crooks and Khim (1999);17 Shimeta (1996);18 Fauchald and Jumars (1979); Kalke and Montagna (1991) 222
Appendix B
Complement of Chapter 3 223
Weight-Length Relations
The relations between weight (W) and length (L)of English sole and starry flounder are compared between estuaries using the linearized relation log W = log a + b log L. Starry flounder attained sizes twice as large as English sole
(Figure P.1) . The weight-length relations of each fish species is represented by a single equation as the regression coefficients did not differ between estuaries (P < 0.05, Figure B.l: A and B) Both mean length and weight were similar between estuaries for English sole (P > 0.05, ANOVA, Figure B.l: A) and both mean length and weight were greater for starry flounder in Yaquina Bay (P < 0.001, ANOVA, Figure P.1: B) Iean sizes of English sole were similar among sites (Table B.l) and the largest starry flounder were found in downstream and mid estuarine sites at low-tide (Table B.2)
Condition Factor
Differences in fish condition between estuaries is evaluated using the Fulton condition factor (C)
C =[W L3]K where W is the total fish weight (g), L is the total fish length (cia) and K is an arbitrary scaling factor = 100 (Anderson and Gutreuter 1983)
Condition factor for each fish species did not differ between estuaries (t-test, P > 0.05, Figure B.2): English sole (Alsea 5C = 0.845, SE = 0.007; Yaquina C = 0.861, SE = 0.008); starry flounder (Alsea C = 1.094, SE = 0.016; Yaquina C = 1.060, SE = 0.012) .Condition factor was not correlated with fish size excepting for English sole in Alsea Bay (r = -0.27; P < 0.01) 224
Figure B.l. Total weight (W) and total length CL) of English sole and starry flounder. Fish were collected at intertidal and subtidal areas in the Alsea Bay and Yaquina Bay estuaries in summer 1993. 225
A English sole 10-W = 0.00877L2983 (R2 = 0.96, n = 246)
8
6 :j Mean 6Jsea Yaquina 'I L (cm) 6.8 7.1 - W(g) 3.0 3.4 I 2- 0 0 5 10 15 20 25
B Starry flounder W = 0.01041L3°°8 (R2 = 0.99, n = 122) '-150 Mean Aisea Yaquina L (cm) 9.0 13.3 W(g) 14.9 34.4 100
50
10 15 20 25
TOTAL LENGTH (mm)
Figure B.1 Table Bi. Ratio of English sole withprey (No. of fish with prey in their stomach analyzed); stomach fullness index; / No. of fish mean fish length and weight andmean prey richness (No. taxa) per fish. Prey origin: NA (native); NI (nonindigenous); CR (cryptogenic) and ALL (NA+ NI + CR + supraspecific taxa). Sites ending in H and L were sampled at high-tide in intertidalareas and at low-tide in subtidal areas, respectively.SE = standard error.
Estuary/Site Ratio Fish Stomach Total Fish Total Fish Mean Prey Richness With Prey Fullness Length (cm) Weight (g) (No. taxa per fish) Mean SE Mean SE Mean SE NA NI CR ALL Alsea Bay: A1H 6/6 3.0 0.7 6.3 0.5 2.5 0.6 1.8 1.2 1.3 6.0 A2H 1/1 4.0 6.9 2.8 2.0 3.0 1.0 9.0 A3H 22/22 3.6 0.7 5.1 0.1 1.3 0.1 2.0 0.5 0.5 5.3 AlL 78/88 1.4 0.1 7.6 0.1 3.8 0.2 3.1 0.5 0.1 7.8 A2L 8/9 1.9 0.2 5.7 0.1 1.6 0.1 2.0 1.0 0.2 6.2 A3L 8/9 1.2 0.5 4.7 0.2 1.0 0.1 1.8 0.3 0.0 5.2 Yaquina Bay: Y1H 19/26 1.6 0.3 7.7 0.2 4.4 0.3 2.2 0.2 0.7 6.2 Y2H 2/2 4.0 0.0 6.6 0.7 2.8 0.7 2.0 2.5 0.0 6.0 Y3H 19/19 3.8 0.2 7.6 0.2 3.9 0.3 3.7 2.8 0.5 10.4 Y4H 12/13 3.0 0.3 7.8 0.2 4.7 0.4 2.2 1.8 0.0 4.7 Y2L 18/18 1.3 0.3 5.5 0.3 1.6 0.2 1.2 0.8 0.1 4.8 Y3L 24/35 1.2 0.2 6.6 0.2 2.5 0.2 0.6 0.8 0.0 3.1 Table B.2. Ratio of starry flounder with prey (No. of fish with prey in theirgut / total No. of fish analyzed); stomach fullness index;mean fish length and weight andmean prey richness (No. taxa) per fish. Prey origin: NA (native);NI (nonindigenous); CR (cryptogenic) CR + supraspecific taxa). Sites ending in and ALL (NA + NI + H and L were sampled at high-tide inintertidal areas and at low-tide in adjacent subtidalareas, respectively. SE = standarderror.
Estuary/Site Ratio Fish Stomach Total Fish Total Fish Mean Prey Richness With Prey Fullness Length (cm) Weight (g) (No. taxa per fish) Mean SE Mean SE Mean SE NA NI CR ALL Alsea Bay: A2H 4/4 2.2 1.1 11.2 2.6 23.0 11.4 2.0 1.7 0.0 5.7 A4H 5/5 3.2 0.8 6.7 0.8 3.3 1.3 2.4 0.8 0.0 3.8 A5H 6/6 0.7 0.7 12.9 1.5 29.7 9.1 2.7 0.7 0.0 4.0 A6H 11/13 2.3 0.6 9.0 1.1 12.0 3.5 2.1 0.5 0.0 5.3 AlL 5/5 1.2 0.8 16.6 2.4 65.8 22.9 1.4 0.8 0.0 4.4 A4L 1/1 4.0 5.6 -- 1.8 -- 4.0 0.0 0.0 6.0 A6L 27/27 3.2 0.2 6.4 0.1 3.1 0.2 2.3 0.8 0.0 6.4 Yaquina Bay: Y1H 3/4 2.2 0.7 15.7 3.2 57.3 22.8 4.2 0.7 0.5 8.5 Y2H 2/2 2.0 2.0 11.5 5.2 25.3 22.9 6.0 0.5 0.5 10.0 Y4H 17/17 2.1 0.4 13.1 1.0 30.8 5.5 4.7 2.9 0.0 11.1 Y5H 6/6 2.7 0.7 10.7 1.1 15.6 4.8 3.7 1.8 0.0 Y6H 13/14 9.5 1.1 0.4 14.7 1.2 41.2 7.1 4.2 0.9 0.0 7.6 Y1L 1/1 4.0 21.5 109.5 -- 3.0 0.0 0.0 6.0 Y3L 3/3 1.0 0.6 19.2 2.6 76.2 29.0 4.3 2.0 0.3 10.0 Y4L 4/4 0.7 0.7 11.2 2.0 19.6 9.9 2.3 1.3 0.0 5.8 Y5L 1/1 4.0 -- 6.5 -- 2.6 -- 3.0 1.0 0.0 6.0 Y6L 9/9 2.2 0.6 11.1 1.9 23.1 8.0 4.2 1.0 0.0 8.0 228
Figure B.2 Fulton's condition factor of English sole and starry flounder. Fish were collected at intertidal and subtidalareas of Alsea Bay and Yaquina Bay estuaries in summer 1993. 229
English Sole
1.1- 135 113
1.0- Median 75 % Percentile 0.9- c 0 / I- Mean 25 % Percentile
LJ z 0.6 0 Alsea Yaquina I- z Starry Flounder 0 C.) 1.4- 61 Cl) 1.3- 61 z 0 I- - J 1.1- U 1.0- 0.9- 0.8- 0.7 Alsea Yaquina ESTUARY
Figure B.2 230
Diet Overlap and Trophic Breadth
Diet overlap is computed between species 1 and 2(DO1,2) within, and between, estuaries (or diet overlap of a given species between estuaries 1 and 2) using Levins' (1968) index of resource overlap:
\1[B1] DO1,2 = E[(P 2j IJ where P1and P2are the proportions of volume for prey in predator species 1 and 2(or the proportions of volume for prey in a given predator species in estuaries 1 and 2) for a number of prey taxa and B1 is the trophic breadth of predator species 1 as defined by the term:
B1= [ (P2)] where P2is the squared proportion of volume for prey
Starry flounder had a greater trophic breadth in Yaquina Bay and English sole had similar trophic breadths between estuaries
(Figure B.3) . The highest and lowest diet overlaps occurred, respectively, between starry flounder and English sole in Yaquina Bay and between starry flounder in Alsea Bay and English sole in
Yaquina Bay (Figure B.3) . Non-significant differences in diet overlap are detected among fish between estuaries (intraspecific mean = 31.7; intraspecific mean = 28.7) and within estuaries (interspecific mean = 38.5);(ANOVA,P > 0.65) 231
Figure B.3. Percent of dietary overlap (DO1,) and trophic breadth
(B ) for flatfish. Fish were collected in the Alsea Bay and Yaquina estuaries in 1993. Rank 1 denotes highest diet overlap. DO and DO were positively related for the six possible pairwise comparisons (r = 0.66, P <0.15) DO English sole English sole Starry flounderStarry flounder Alsea Bay Yaquina Bay Alsea Bay Yaquina Bay (Rank) B= 8.1 B= 7.4 B1= 4.7 B1=9.1 English sole Alsea Bay English sole Yaquina Bay Starry flounder 16 13 Alsea Bay (12) Starry flounder Yaquina Bay
Percent of Diet Overlap (DO1 ):
Figure B.3 233
Individual Prey Volume
The number of native species and NIS along the prey volume spectra are estimated from the volume of prey species in predator species (IPV1), which is defined as:
-1 IPV = [ V] [ N] where V1and are the volume and number of prey species in the diet of n predators of species
Native and NI prey in the diet of each fish species had similar lower ranges for individual prey volumes (Figure B.4). However, native prey had a consistently higher range for individual prey volumes when compared to NI prey. The volume for most species ranged from 1 to 100 mm3 in both fish species. The larger size range and average size of starry flounder accounted for its wider prey volume spectra in both estuaries (Figure B.4) 234
Figure B.4. Number of native and nonindigenous speciesby volume of individual prey in the diet of juvenile English sole and starry flounder in the Alsea and Yaquina estuaries. No. NI SPECIES No. NATIVE SPECIES No. NI SPECIES No. NATIVE SPECIES Nippoleucon hinurrensis a' Cii 0-a C-) c, -S Mya arenaria (a) Arrnandia bresis 0-s 01> Eobrolgus spinosus Hobsonia florida Eteone spilotus Corophium spinicornesalmonis 0 Eobrolgus spinosusNippoleucon hmnurnensisHobsonia florida MacomoCurnelia uvigoris bait/rica (a) Pseudo polydora kempi CrangonCryptomyaMacurnsCurrella silgarisfranciscorumbaithica californica (a) Pseudopolydora kempi Mya arenaria NephthysGlycinde polygnathacoeca zD Mya arenaria MacomaNereis Iimr,icola balthica Nereis hmnico)a Ca Mya arenaria (a) Cymatogasfer aggre gata 0 'ii- Taxa Alsea Bay Yaquina Bay Prey Mean Mean Prey Prey Mean Mean Prey Occurr. No. Prey Vol. (mm ) Occurr. No. Prey Vol. (mm3) Anne lida Polychaeta: Amaena occidentalis 4.4 0.186 1.373 Armandia brevis 39.3 1.807 1.727 4.4 0.265 0.186 Boccardiaproboscidea 5.3 2.381 3.765 Capitella sp. 11.1 0.163 0.233 4.4 0.062 0.085 Capitellid part 14.8 0.356 0.327 23.0 0.168 0.247 Dorvillea annulata 0.9 0.009 0.022 Dorvilleidae 0.7 0.007 0.004 Eteone californica 0.9 0.009 0.088 Eteone sp. 3.5 0.035 0.007 Eteone spilotus 4.4 0.059 0.533 16.8 0.301 1.259 Exogone sp. 0.7 0.007 0.007 0.9 0.009 0.001 Glyceridae part 0.7 0.007 0.007 Glycinde armigera 0.9 0.009 0.619 Glycinde polygnatha 4.4 0.044 0.611 0.9 0.009 0.133 Hobsonia florida * 9.6 0.200 0.367 2.7 0.044 0.032 Manayunkia aestuarina 4.4 0.071 0.011 Mediomastuscaliforniensis 1.5 0.022 0.022 2.7 0.035 0.178 Nephthys caeca 0.7 0.007 0.296 Nerejs limnicola 1.5 0.015 0.504 Owenia fusiformis 1.5 0.015 0.015 0.9 0.009 0.044 Phyllodoce Hartman 0.7 0.007 0.007 0.9 0.009 0.004 Polychaeta part 29.6 0.570 0.345 26.5 0.867 0.743 Pseudopolydorakempi * 6.7 0.141 0.774 25.7 1.619 5.797 Pygospio californica 1.5 0.015 0.015 Table B.3. Continued. r axa Aisea Iay Yaquina Bay % Prey Mean Mean Prey Prey Mean Mean Prey Occurr. No. Prey Vol. (mm ) Occurr. No. Prey Vol. (mm3) Pygospio elegans ** 17.8 2.556 1.430 10.6 0.487 0.396 sp.l Rhynchospio 1.8 0.018 0.018 Spionidae (juvenile) 1.5 0.022 0.004 Streblospio benedicti * 39.8 7.593 17.878 Tharyx sp. 0.9 Oligochaeta 0.009 0.004 19.3 0.274 0.221 9.7 0.292 0.235 Arthropoda Acarina 0.9 0.009 0.001 Crustacea Brachyura: Hemigrapsus oregonensis 5.2 0.044 0.052 Caridea: Crangon franciscorum 0.9 0.009 Amphipoda: 0.022 Allorchestes angusta 1.8 0.018 0.018 Amphipoda part 0.7 0.007 0.001 Ampithoe sp. 0.7 0.007 0.007 Ampithoe valida * 0.9 0.009 Corophium acherusicum * 0.022 2.7 0.018 0.031 Corophium brevis 7.1 0.832 2.201 Corophium salmonis 45.2 1.963 3.737 37.2 4.619 10.839 Corophium spinicorne 0.7 0.007 0.007 9.7 spinosus * 0.434 2.177 Eobroigus 2.2 0.030 0.063 1.8 0.044 0.053 Eogammarus confervicolus 9.6 0.119 0.156 4.4 0.071 Copepoda: 0.157 Cyclopoid copepod Hemicyciopssubadhaerens 1.5 0.022 0.015 Harpacticoida 87.7 26.615 1.573 66.4 24.283 1.548 Table B.3. Continued. 1 axa Alsea Bay Yaquina Bay % Prey Mean Mean Prey Prey Mean Mean Prey Occurr. No. Prey Vol. (mm ) Occurr. No. Prey Vol. (mm3) Crustacean part 3.0 0.037 0.023 3.5 0.062 Cumacea: 0.025 Cumacean part 1.8 0.018 0.018 Cumella vulgaris 47.4 1.281 0.804 his umensis * 22.1 0.327 0.205 Nippoleucon 16.3 0.363 0.189 38.1 Macrura: 7.407 4.777 californiensis Neotrypaea 4.4 0.044 0.230 2.7 0.035 0.155 Upogebia pugettensis 2.2 0.030 0.052 Tanaidacea: Leptochelia dubia ** 0.7 0.030 0.004 12.4 3.062 3.040 Pancolus californiensis ** 2.2 0.022 0.027 0.9 Synelobus stanfordi ** 0.009 0.004 3.0 0.030 0.030 3.0 0.030 0.030 Ostracoda 20.7 0.778 0.127 8.8 0.142 0.033 Insecta Diptera: Chironomidae 0.7 0.007 0.007 4.4 0.053 0.059 Insect parts 0.7 0.007 0.007 Mollusca Bivalvia: Bivalve siphon Bivalve part 0.7 0. 007 0.007 0.9 0.009 Ciinocardium nuttaili 0.004 10.4 0.126 0.233 0.9 0.009 0.004 Cryptomya californica 2.2 0.022 0. 033 1.8 Macoma 0. 027 0.031 baithica 3.0 0.044 0.048 0.9 Macoma baithica (siphon) 0.018 0.018 77.8 29.548 6.670 43.4 5. 602 2.452 Table B.3. Continued. Taxa Alsea Bay Yaquina Bay Prey Mean Mean Prey % Prey Mean Occurr. No. Prey Mean Prey Vol. (mm ) Occurr. No. Prey Vol. (mm3) Bivalvia: Mya arenaria * 17.8 0.319 1.296 4.4 0.133 0.642 Mya arenaria (siphon) * 1.5 0.022 0.011 Mysella tumida 3.0 0.037 0.041 Transenella tantilla 0.9 0.009 0.004 Gastropoda 0.7 0.007 0.019 Phoronida Phoronjda part 37 0.044 0.056 Miscellaneous items: Feather 9.6 0.133 0.063 Invertebrate parts 2.7 0.053 0.451 Organic matter 47.4 0.785 0.630 27. 4 Plant matter 0.416 0.360 13.3 0.185 0.092 20.4 0.646 0.489 Plastic line 2.2 0.030 0.011 3.5 0.062 0. 012 Seed 1.5 0.015 0.052 Stone 32.6 0.763 0.467 Wood fragment 5.3 0.124 0.064 56.3 1.615 0.945 34.5 0.558 0.336 Table B.4. Frequency of preyoccurrence and mean number and volume of starry flounder in intertidal-subtidal prey consumed by juvenile areas of Alsea Bay and Yaquina Bay duringsummer 1993. Nonindigenous and cryptogenic speciesare denoted by one and two asterisks, 61 fish per estuary. respectively. Based on Taxa Alsea Bay Yaquina Bay Prey Mean Mean Prey Prey Mean Mean Prey Occurr. No. Prey Vol. (rnm) Occurr. No. Prey Vol. (mm3) Anne 1 ida Polychaeta: Capil:ella sp. Capitellid part 4.9 0.098 0.066 4.9 0.148 0.148 Eteone spilotus 3.3 0.033 0.007 Heteromascus filiformis * 4.9 0.082 0.148 florida * 1.6 0.016 0. 492 Hobsonia 57 . 4 1.082 5.149 49.2 Manayunkia aestuarina 7.852 13. 684 16.4 Mediomastuscall forniensis 1. 623 0.262 3.3 0.033 0.180 Nereis lirnnicola 67.2 1.492 78.010 Paraonellaplatybranchia 11.5 0.180 5.820 1.6 0.016 0.016 Polychaeta part 4.9 0.066 0.066 21.3 0.426 0.379 Pseudopolydorakempi * 3.3 0.148 1.148 36. 1 6.623 71.811 Pygospio elegans 1.6 0.016 0.016 Streblospio benedicti * 1.6 0.279 0. 098 24.6 Tharyx sp. 11.738 23. 992 01 igochaeta 1.6 0.049 0.033 1.6 0.016 0.016 6.6 0.230 Arthropoda Crustacea Caridea: Crangon franciscorum 1.6 0.016 0.328 4.9 0.098 0.869 Table B.4. Continued. Tax a Alsea Bay Yaquina Bay Prey Mean Mean Prey Prey Mean Prey Occurr. No. Prey Mean Vol. (mm ) Occurr. No. Prey Vol. (mm3) Amphipoda: Ampithoe lacertosa Corophium acherusicum * 1.6 0.016 0.016 3.3 0.164 Corophium salmonis 86.9 0.107 13.410 38.559 70.5 18.852 58.656 Corophi urn spinicorne 31.1 0.770 1.754 31.1 5.623 28.279 spinosus * Eobrolgus 1.6 0.033 0.016 Eogarnrnarusconfervicolus 3.3 0.049 0.049 1.6 0.033 0.246 Copepoda: Cyclopoid copepod 3.3 0.066 0.007 Hernicyciops subadhaerens 1.6 0.049 0.002 1.6 0.180 0.016 Harpacticoida 52.5 27.066 2.408 22.9 1.279 0.164 Crustacean part 3.3 0.049 0.180 Cumacea: Curne]ia vulgaris 1.6 0.016 0.082 4.9 0.115 0.066 Nippoleucon hinurnensis * 6.6 0.115 Macrura: 0.082 18.0 1.443 1.311 Neotrypaea californiensis Upogebia pugettensis 9.8 0.311 12.623 Tanaidacea: 1.6 0.033 0.656 Leptocheliadubia ** 3.3 1.672 0.639 Ostracoda 22.9 0.410 0.075 9.8 0.279 0.069 Insecta Diptera: Chironomidae 47.5 2.230 1. 154 11.5 0.393 0.557 Diptera pupae 42.6 1. 082 Insect parts 0. 670 3.3 0.033 0.197 Orthoptera 1.6 0.016 0.016 1.6 0.016 0.016 Table B.4. Continued. Taxa Alsea Bay Yaquina Bay o Prey Mean Mean Prey Prey Mean Mean Prey Occurr. No. Prey Vol. (mm ) Occurr. No. Prey Vol. (mm3) Mollusca Bivalvja: Bivalve siphon 3.3 0.066 0.049 Bivalvepart 1.6 0.016 0.008 1.6 0.016 0.008 1.6 Clinocardjumnuttallj 0.016 0.016 3.3 0.023 4.180 Cryptomya californica 3.3 0.131 2.049 Crtomya californica (siphon) 1.6 0.016 0.082 3.3 0.066 0.328 Macomabalthjca 6.6 0.197 13.934 Macoma 13.1 0.311 42.311 baithica (siphon) 18.0 0.508 0.852 34.4 Myaarenaria* 6.820 8.805 3.3 0.066 10.492 22.9 1.311 56.231 Myaarenaria (siphon) * 4.9 0.721 75.738 Mysella tumida 1.6 0.131 0.098 1.6 0.033 0.025 Phoronida Phoronida part 1.6 0.033 0.016 Fish: Cymatogaster aggregata 1.6 0.016 24.590 Engraulis mordax 1.6 0.016 6.557 1.6 0.016 16.393 Miscellaneous items: Feather Organic matter 3.3 0.033 0.018 8.2 0.213 2.680 13.1 0.131 0.189 Plantmatter 18.0 0.361 1.346 Plastic line 18.0 0.672 18.541 1.6 0.016 0.033 Stone 9.8 0.164 0.131 14.7 0.508 0.946 Wood fragment 32.8 0.721 0.987 50.8 2.148 7.649 243 Appendix C Complement of Chapter 4 244 Males Females C. salmonis C. salmonis 1500 Y-52.74+572.3X 1500 Y = 21.29 e6129x (R2= 0.71, 39d.f.) 2_ 0.64, 38 d.f.) 1000 1000 500 500 0 0 00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0 C. spinicorne C. spin/come 1500Y =- 334.5+ 997.6X 1500 Y=-208.3+358.3X+ 781.8X2 (R2 0.81,39 dJ.) 1000 1000 500 500 R2= 0.77,38 d.f.) 0 00 0.5 1.0 1.5 2.0 0 0 0.5 1.0 1.5 2.0 C. acherusicum C. acheruskum 1500 Y 54.90e4 1500 Y = 34.40 + 368X (R2 0.50, 36 d.f.) (R2= 0.16, 39 d.f.) 1000 1000 500 500 00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0 C. insidiosum C. insidiosum 1500- Y51.01 e18 1500 V = -15.99 + 485.5X (R20.37, 40 d.f.) (R2= 0.18, 39d.f.) 1000 1000 500 :. 500- 00 0.5 1.0 1.5 2.0 00 0.5 1.0 1.5 2.0 ANTENNA LENGTH (mm) Figure C.l. Amphipod dry weight (Y)with length of 4th article 2nd antenna (X) by species andsex. All correlations are significant (P < 0.05). 245 30- I Treatment C,) 0 a a Cl) zdl z w SAL SPI ACH INS SPECIES Figure 0.2. Mean number of surviving Corophium in single- species predation treatments and in controls without fish. (SAL = C. saimonis, SPI = C. spinicorne, ACH = C. acherusicum, INS = C. insidiosum), with standard error scale indicated over each bar and different letters above bars denoting significant differences in the proportion of survivors between treatment and control at P < 0.05 (One-tail X2 test) 246 A SAND Control Treatment a a I -I- SPI ACH INS B MUD Control Treatment SAL SPI ACH INS SPECIES Figure C.3. Mean nuither of surviving Corophium in mixed-species predation experiments in sand and mud treatments and in controls without fish.(SAL = C. saimonis, SPI = C. spinicorne, ACH = C. acherusicum, INS = C. insidiosum), with standarderror scale indicated over each bar and different letters above bars denoting significant difference in the proportion of survivors between treatment and control at P < 0.05 (One-tail X2 test). Table C.l. Density and number of Corophiuzn salmonis in the benthos and the diet of juvenile English sole collected in intertidal areas at high tide. Yaquina Bay sites: Y3H (July 7,1993) and Y4H (July 18 1993) and Alsea Bay site (A3H, July 7, 1993) . Fish stomachs analyzed included only those with 5 C. salmonis or more. Thirty core samples were collected per site (10 cores stratified at about 0,40 and 80 cm of water depth, 8.04 cm2 each core) . Original prey length includes all C. salmonis found in stomachs in the field. Adjusted prey length (4th article 2nd antenna) includes only the size range used in single-species feeding experiments (Figure 4.1) Significance of the difference in the proportion of males (M) and females (F)in the diet and in the benthos is based on the number of prey counted and is indicated by X2 two-tail test. Original Prey Length Adjusted Prey Length Site No. Prey Size Range (mm) Density No. Prey P Density No. Prey X2 P Fish Benthos Fish Diet Benthos Fish Diet M F M+F M/F M+F M/F M+F M/F M+F M/F (Norrr2) (No.m2) A3H 4 0.18-1.58 0.13-0.50 954 0.44 49 1.23 3.82 0.05 829 0.43 28 1.33 3.46 0.06 Y3H 13 0.20-1.68 0.15-0.53 2695 0.62 245 0.67 0.82 0.82 2405 0.53 127 1.12 5.34 0.02* Y4H 5 0.13-2.38 0.13-1.00 3027 0.70 45 0.61 0.13 0.72 1990 0.65 21 1.10 0.97 0.32 * Significant difference (P < 0.05). 248 Appendix D Complement of Chapter 5 249 Table D.1. Number of prey and their percent frequency of occurrence in stomachs of juvenile staghorn sculpin (Leptocottus armatus) . Thirty-six fish collected in Yaquina Bay sites 3 and 4 during summer 1993 were analyzed. Species origin: nonindigenous (*); cryptogenic (**); native (no asterisk). Taxa Mean Percent No. Occurrence Bivalvia Macomabalthicasiphon 0.56 19.4 Mya arenaria* 0.03 2.8 Mya arenaria* (siphon) 0.03 2.8 Polychaeta Nerei.s limnicola 0.11 16.7 Arthropoda Acarina 0.03 2.8 Crustacea 2mph ipoda Ampithoe lacertosa 0.17 16.7 Ampithoe sp. 0.25 22.2 Corophium salmonis 7.78 75.0 Corophium spinicorne 2.31 44.4 Eobrolgusspinosus * 0.03 2.8 Eogammarus confervicolus 1.55 36.1 Ca ride a Crangon franciscorum 0.14 11.1 Copepoda Copepoda, unidentified 0.05 8.3 1-larpacticoid 0.03 2.8 Hemicyclops sp. 0.03 2.8 Cumacea Nippoleuconhinumensis * 0.74 16.7 Decapoda Zoea 0.03 2.8 Neotrypaeacaliforniensis 0.08 5.6 Hernigrapsus oregonensis 0.03 2.8 Tanaidacea Pancoluscaliforniensis ** 0.64 16.7 Isopoda Gnorissmosphaeroma insulare 0.28 16.7 Gnorissmosphaeroma sp. 0.06 2.8 Fish Fish, fragment 0.19 11.1 Leptocottus armatus 0.03 2.8 Engraulis mordax 0.33 5.6 Cymatogaster aggregata 0.06 5.6 Miscellaneous Organic matter 0.47 30.6 Plant matter 0.14 13.9 Wood fragment 1.22 50.0 Stone 0.75 47.2 250 Table D.2. Activity models derived from models in Figure 5.2. Negative effect to guild i from guild jis denoted as (i,j)=-l. Reciprocal negative effect between guilds i and jis denoted as (i/j)=-l. Size-dependent interactions for different mobility types (Yl; Y2) and same mobility types (Xl; X2) are defined by relative size of guilds (s= small; i= intermediate; 1= large) Interactions Yl and Xi: InteractionsY2andX2: Relative sizes Relative sizes S i 1 S i 1 S -1 -1 -1 S -1 -1 -1 1 0 -1 -1 i -1 -1 -1 1 0 0 -1 1 0 -1 -1 Ba sic Model Model Additional Interactions A2 Al (M,S)=-1 A3 Al (M/S)=-i A5 A4 (M*,S*)=_i;(M*,S)=_l; (M,S)=-l; (M,S*)=_i A6 A4 (M*/S*)=i; (M*/S)=_l;(M/S)=-i; (M/S*)=_i A7 A4 (M*/S*)=_1;(M*/S)=_i;(M/S)=-l; (M/S*)_l; (M*/M)=_l A8 A4 (M*/S*)=_1;(M*/S)=_i;(M/S)=-l;(M/S*)=_1;(M*/M)=_i; (S*/S)=_i AlO A9 X2 All A9 Xl Al2 A9 Y2 A13 A9 Y2; X2 A14 A9 Y2; Xi A16 A15 X2 A17 A15 Xl A18 A15 Y2 A19 A15 Y2; X2 A20 A15 Y2; Xl A22 A21 X2 A23 A21 Xl A24 A21 Y2 A25 A21 Y2; X2 A26 A21 Y2; Xl A28 A27 X2 A29 A27 Xl A30 A27 Y2 A31 A27 Y2; X2 A32 A27 Y2; Xl A34 A33 X2 A35 A33 Xl A36 A33 Y2 A37 A33 Y2; X2 A38 A33 Y2; Xl A40 A39 X2 A41 A39 Xl A42 A39 Y2 A43 A39 Y2; X2 A44 A39 Y2; Xl 251 Table D.3. Trophic models derived from models in Figure 5.3. Negative effect to guild i from guild jis denoted as (i,j) - 1. Positive effect to guild ± from guild jis denoted as(i,j)= 1. Reciprocal positive effect between guilds i and jis denoted as (i/j)= 1. Ba s ± c Model Model Additional Interactions T7 Ti (Su*,Sr*)= -1 T8 T2 asinT7 T9 T3 (Su*,Sr*)= -1;(Su*,Sr)= -1;(Su,Sr*)= -1;(Su,Sr)= -i TiO T4 as inT9 Til T5 as inT9 T12 T6 as in T9 T17 Ti (Sr*/Sb*)= 1 T18 T2 as in T17 T19 T3 (Sr*/Sb*)= 1;(Sr*/Sb)= 1;(Sr/Sb*)= 1;(Sr/Sb)= 1 T20 T4 as inT19 T21 T5 as inT19 T22 T6 as inT19 T23 Ti (Sb*, Sr*)= 1 T24 T2 as inT23 T25 T3 (Sb*, Sr*)= 1;(Sb*, Sr)= 1;(Sb, Sr*)= 1;(Sb,Sr)= 1 T26 T4 as inT25 T27 T5 as inT25 T28 T6 as in T25 T29 Ti (Sr*/Sb*)= 1;(Su*,Sr*)=_i T30 T2 as inT29 T3i T3 as inT9andTi9 T32 T4 as inT9and Ti9 T33 T5 as inT9andT19 T34 T6 as inT9andTi9 T35 Ti as in T7 andT23 T36 T2 as in T7 andT23 T37 T3 as inT9andT25 T38 T4 as inT9andT25 T39 T5 as inT9andT25 T40 T6 as inT9andT25 '141 Ti3 as inT9 T42 T14 as inT9 T43 T15 as inT9 T44 T16 as inT9 T45 T13 as inTi9 T46 T14 as in T19 T47 Ti5 as in Ti9 T48 T16 as inTi9 T49 T13 as inT25 T50 T14 as inT25 T5i Ti5 as inT25 T52 T16 as inT25 T53 Ti3 as inT9andTi9 T54 T14 as inT9andT19 T55 T15 as inT9and T19 T56 Ti6 as inT9andTi9 T57 T13 as inT9andT25 T58 '114 as in T9andT25 T59 Ti5 as inT9andT25 T60 T16 as inT9andT25