AN ABSTRACT OF THE THESIS OF
Gonzalo C. Castillo for the degree of Doctor of Philosophy in
Fisheries Science presented on July 14, 2000 Title: Benthic
Biological Invasions in Two Temperate Estuaries and Their
Effects on Trophic Relations of Native Fish and Community
Stability.
Redacted for Privacy
Abstract approved:
Hiram W. Li
The extent of biological invasions, their role on the feeding of
native fishes and their impact on community stability were
investigated in Alsea Bay and Yaquina Bay, two estuaries on the
central Oregon coast, USA. Most nonindigenous species (NIS)
introduced in these intermediately invaded estuaries are
considered byproducts of culturing introduced Atlantic and
Pacific oysters. Secondary potential vectors of NIS in Yaguina
Bay are external fouling of ship hulls and ballast water. Native
benthic invertebrates and native fishes dominate in density,
catch per unit effort (CPUE) and richness in both estuaries.
Three of the 11 benthic NIS of invertebrates in Yaquina Bay and
one of the eight NIS in Alsea Bay are among the 10 most dominant
benthic invertebrate species. The NIS of invertebrates are
concentrated in habitats with above average water temperature,
salinity, and macrophyte density at high-tide. The CPUE of
fishes and decapod crustaceans are associated wi.th above average
water temperature, salinity and macrophyte density but are not
consistently correlated with invertebrate density in sediments.
Biological invasions have caused significant prey shifts in
intertidal food webs of Yaquina Bay. Diets of two species of
native juvenile flatfishes
(Pleuronectes veLulus
and
Platichthys
stellatus) included mainly polychaetes, crustaceans and bivalves
and each of these taxa are represented in the diet by native
species and NIS in each estuary. Both flatfish species are
generalist predators and had no consistently higher selection
for either native species or MIS. Prey selection experiments
indicated that two native and two introduced amphipod prey
(Corophium spp.) are acceptable prey for juvenile English sole.
Thus, predator-prey coevolution plays no significant role on
prey selection. Interspecific prey selection may depend on prey
exposure, water visibility, substratum type, and species
diversity of available prey. Modeling of functional-group
interactions for the intertidal benthic community of Yaquina Bay
suggested reduced community response to invasions or removal of
fish predators as indicated by the community tendency to zero
overall-feedback. However, the increased risk of stability
decline of invaded community models implies that further humanmediated biological invasions should be avoided.
Benthic Biological Invasions in Two Temperate Estuaries
and Their Effects on Trophic Relations of Native Fish
and Community Stability
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 presented on
July 14, 2000
APPROVED:
Redacted for Privacy
Major Professor, representing Fisheries Science
Redacted for Privacy
Chair of Department ofisheries and Wildlife
Redacted for Privacy
Dean of G. .'e School
I understand that my thesis will become part of the permanent
collection of Oregon State University libraries. My signature
below authorizes release of my thesis to any reader upon
request.
Redacted for Privacy
hnzalo C. Castillo, Author
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
Chapter 2: Distribution and Habitat Use by Noncoevolved
Assemblages of Macroinvertebrates and Fishes in Two
Temperate Estuaries
10
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
Page
Figure
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
Summer mean CPUS of fishes
samples from the Alsea Bay
(bars) and percent of CPUE
relative to native species
35
2.3
2.4
2.5
2.6
2.7
2.8
2.9
and decapods in seine
and Yaquina Bay estuaries
of nonindigenous species
(circle)
Clusters by taxa and sites based on invertebrate
densities in sediment samples
38
Clusters by taxa and sites based on CPUS of
fishes and decapods in seine samples
40
Mean summer densities of assemblages of native and
nonindigenous invertebrates in sediment samples
under various temperature-salinity combinations
43
Mean percent densities for summer assemblages of
native and nonindigenous invertebrates under
various temperature-salinity combinations
44
Mean percent richness for summer assemblages of
native and nonindigenous invertebrates under
various temperature-salinity combinations
45
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
3.1
3.2
3.3
3.4
48
Fish and invertebrate collection sites in Alsea
Bay and Yaquina Bay
68
Mean number and volume of prey species in the
diets of English sole and starry flounder
78
Percent volume of major prey by origin in English
sole and starry flounder diets
80
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
Percentages of prey frequency of occurrence,
number and volume of main dietary items of starry
flounder by estuary section
85
Total prey volume of native and nonindigenous
species and all taxa combined as a function of
fish weight in intertidal areas
90
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
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
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
3.5
3.6
3.7
3.8
3.9
4.1
4.2
4.3
4.4
4.5
4.6
4.7
Amphipod body length (from telson to eye, Y) with
length of 4th articLe 2nd antenna (X) by species and
sex
120
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
Mean number of Corophium consumed by Pieuronectes
vetulus, in single-species experiments
125
Strauss' selection index by prey size (4th article
2nd antenna) and Corophium species consumed by
Pleuronectes vetuius
126
Percent of eaten and urieaten Corophium by size
(4th article 2nd antenna) in 10 tanks with sand
substratum
127
Mean number of Corophium consumed by Pieuronectes
vetulus in mixed-species experiments
129
Percent of uneaten Corophium by size (4th article
2nd antenna) in sand and mud substrata
130
LIST OF FIGURES (Continued)
Figure
5.1
5.2
5.3
5.4
5.5
Page
Ecological interactions between guilds 1,
and 3 in numbered circles and attendant
community matrices
2
147
Basic guild structure of activity models for the
benthic community of the Yaquina Bay estuary
156
Basic guild structure for trophic models of the
benthic community in the Yaquina Bay estuary
158
Distribution of feedback in models for the
pre-invaded and the invaded benthic community of
Yaquina Bay
164
Percent of models with near-zero feedback
(-2 F
2)
165
LIST OF TABLES
Table
2.1
2.2
2.3
2.4
3.1
3.2
3.3
3.4
3.5
4.1
4.2
5.1
Page
Substrata, vegetation types and macrophyte density
of intertidal sites in the Alsea Bay and Yaquina
Bay during summer 1993
27
Summer density and occurrence (OC) of intertidal
benthic invertebrates in core samples from Alsea
Bay and Yaquina Bay
29
Summer catch per unit effort (CPUE) and occurrence
(OC) of fishes and decapods in the Alsea Bay and
Yaquina Bay
34
Percent of community variance explained,
eigenvalues and correlations for the three main
axes of CCA ordinations
50
Total number, mean total length and total weight
of juvenile English sole and starry flounder
74
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
Frequency of species occurrence in the diet of
juvenile English sole (E) and starry flounder (S)
in Alsea Bay and Yaquina Bay
76
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
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
Walking (WA), swimming (SW) and partially visible
(PV) Corophium spp. in sand substrate
122
Walking (WA), swimming (SW) and partially
visible (PV) Corophium spp. in sand and
mud substrata
128
Activity and trophic invertebrate guilds assigned
to qualitative models of Yaquina Bay
145
LIST OF TABLES (Continued)
Table
5.2
5.3
5.4
5.5
Page
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
Guild structure of trophic guild models and
assumptions; number of guilds and invasion
status of each community structure
151
Alternative activity guild models for the
intertidal benthic community of Yaquina Bay
160
Alternative trophic guild models for the
intertidal benthic community of Yaquina Bay
161
LIST OF APPENDIX FIGURES
Figure
B.1
B.2
B.3
B.4
0.1
0.2
C.3
Page
Total weight (W) and total length (L) of English
sole and starry flounder
224
Fulton's condition factor of English sole and
starry flounder
228
Percent of dietary overlap (DO±) and trophic
breadth (B1 )for flatfish
231
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
mphipod dry weight (Y) with length of 4th
article 2nd antenna (X) by species and sex
244
Mean number of surviving Corophium in single-species
predation treatments and in controls without fish ..
Mean number of surviving Corophiurn in mixed-species
predation experiments in sand and mud treatments and
in controls without fish
..
245
246
LIST OF APPENDIX TABLES
Table
A.1
A.2
A.3
A.4
A.5
B.1
B.2
B.3
B.4
0.1
D.1
Page
Summer mean density and overall occurrence of
intertidal invertebrates in sediment samples
from Alsea Bay
210
Summer mean density and overall occurrence of
intertidal invertebrates in sediment samples from
Yaquina Bay
214
Summer mean catch per unit effort (CPUE) and
occurrence of fishes and decapods in Alsea Bay as
determined from seine sampling
219
Summer mean catch per unit effort (CPUE) and
occurrence of fishes and decapods in Yaquina
Bay as determined from seine sampling
220
Life-mode and functional-groups of nonindigenous
invertebrates found in intertidal and subtidal
areas of Alsea and Yaquina Bay
221
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
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
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
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
Density and number of Corophium salmonis in
the benthos and the diet of juvenile English sole
collected in intertidal areas at high tide
247
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
D.3
Trophic models derived from models in Figure 5.3
.
.. .
250
.
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 humanmediated invasions of aquatic organisms can not be accounted for
by natural dispersal mechanisms. Examples of the latter include
species with life-cycles restricted to brackish-water systems
such as estuaries (e.g., Canton 1979, Cohen and Canton 1995)
and enclosed seas (Carlton 1979, Leppakoski 1994)
Mechanisms of Biological Invasions
The major recent phyletically and ecologically nonselective
vector for the inadvertent dispersal of aquatic organisms is the
release of ballast water from ships (Jones 1980; Carlton and
Geller 1993). The world's fleet has at least 35,000 ships
transporting ballast water (Canton 1999)
.
The sheer scale and
magnitude of this vector are such that it has been referred to as
"conveyor-belts" exchanging species among otherwise isolated
ecosystems around the world (Canton and Geller 1993)
.
The use of
ballast water dates from the 1850's and became significant by the
1880's (Stewart 1991). External fouling of ship's hulls also has
been recognized as important vectors for the inadvertent
introduction of many NIS (Elton 1958; Cohen and Carlton 1995)
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 six
sian copepods (Cordell and Morrison 1996)
and the European green crab Carcinus maenas which invaded many
Northeast Pacific estuaries since the mid 1990s (Cohen et al.
1995; Miller 1996; Beherens and Hunt, in press)
Estuarine fishes are typically assumed to rely on
opportunistic use of prey (e.g., Barnes 1974; Day et al. 1989).
However, juvenile salmonids may have adapted their spatiotemporal use of the Squamish estuary, British Columbia, to the
production of Eogammarus confervicolus (Levings 1980) .
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 functionalgroup 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
(Platichthys
(Chapter 3) .
stellatus) are considered in field analyses
Behavior of two native and two NI amphipods
(Corophium spp.) and consumption and selection of the latter prey
by juvenile English sole is further considered in laboratory
experiments (Chapter 3)
Stability patterns before and after NIS invasions in the
benthic community of Yaquina Bay are estimated using two types of
functional group interaction models. Namely, activity models,
which emphasize physical interactions among invertebrates, and
trophic models, which emphasize direct and indirect trophic
interactions among invertebrates. English sole, starry flounder
and the native fish staghorn sculpin (Leptocottus armatus) are
the major benthic predators included in all models along with
both native and NI benthic macroinvertebrate prey (Chapter 5)
10
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17
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, shiptraffic, and/or better conditions for NIS in the more disturbed
and polluted habitats of Yaquina Bay. Noncoevolved interactions
among similar taxa may not be more likely when compared to
distantly related taxa. Highest mean densities of NIS of
invertebrates at low- and high-tide coincided with: 1) high-mid
temperatures in both estuaries, 2) mid salinities in Alsea Bay
and 3) mid-low salinities in Yaquina Bay.
Most of the population
variations of invertebrates and fishes in intertidal areas at
high-tide are accounted for by macrophyte density, water
temperature and salinity. High values for the latter three
environmental factors are associated with greater NIS densities
in most invertebrates. The CPUE of native fishes and decapod
crustaceans do not vary with invertebrate densities in sediment
samples. Further species introductions should be prevented if
dominance of native species and their potential ecological
functions are to be maintained.
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), noncompetitive species interactions (e.g., Cohen and Carlton 1995),
and reduced community response to invasions (Castillo et al.
2000)
.
Considering that the type and degree of species
interactions in benthic communities is influenced by the
ecological similarity among species (e.g., Whitlatch 1980; Woodin
1983), the need for comparing the distribution of noncoevolved
taxa seems critical to infer which groups of organisms are more
likely to interact in estuaries.
Northeast Pacific estuaries are nursery grounds for many
native species of fishes and invertebrates (e.g., Haertel and
Osterberg 1967; Pearcy and Myers 1974; Bayer 1981; De Ben et al.
1990; Bottom and Jones 1990; Jones et al. 1990)
.
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 temperaturesalinity combinations?; and 5) Are native species and NIS
similarly distributed across
downstream to upstream areas?
environmental
gradients from
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: 20-29 0/
< 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
Site
Upstream
Substrata
Km
Aquatic vegetation
Algae
Macrophyte
Density
Eelgrass
Alsea Bay
Al
4.5
A2
4.9
A3
5.3
A4
7.6
A5
8.6
A6
10.6
Sand/Polychaete tubes
Sand/Mud/Clay
Sand/Mud
Sand/Mud/Clay
C
C
3
C
A
4
A
C
4
C
C
3
Sand/Mud
Mud/Clay
R
R
2
R
N
1
Yaquina Bay
Yl
4.3
Sand
A
C
4
Y2
9.2
Sand/Mud
A
5
Y3
12.2
A
Y4
14.9
C
Y5
18.6
Sand/Mud/Cobble
Sand/Cobble/Clay
Mud
A
A
A
C
C
3
Y6
23.3
Mud/Woody debris
R
N
I
5
4
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 and
paucibranchiata
Pseudopolydora
and 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).
Origin/Vector
T axa
Alsea Bay
Yaquina Bay
No.m2 (OC)
No.m2 (OC)
Annelida
Polychaeta:
Abarenicola
pacifica
Arnaena occidentalis
Armandia brevis
Boccardia proboscidea
Dorvillea annulata
Eteone californica
Eteone columbiensis
Eteone spilotus
Eupolymnia heterobranchia
Exogone lourei
Glycera americana
Glycinde armigera
Glycinde polygnatha
Heteromastus filiformis
Hobsonia florida
Leithoscoloplos pugettensis
Lumbrineris
zonata
Magelona hobsonae
Manayunkia aestuarina
Mediomastus californiensis
Nephtys
caeca
limnicola
Owenia fusiformis
Paraonella platybranchia
Phyllodoce hartmanae
Platynereis bicanaliculata
Polydora cornuta
Pseudopolydora kempi
Nereis
Pseudopolydora paucibranchia La
Pygospio californica
Pygospio elegans
Sphaerosyllis californiensis
Streblospio benedicti
N
8
N
N
N
N
C
N
N
N
C
N
N
N
2
3
2
77
(33)
-
(11)
(11)
4
-
1
(11)
(78)
-
29
(A/B4O)"3
(A/O?)2
N
N
N
2
758
1
-
N
4
N
N
137
N
C
N
N
N
401
2
1
-
(A/O,B,F)3
(I/O,B,F)3
(J?/O,B,F)3
N
C
N
8
254
-
(11)
(44)
(22)
(11)
(22)
(56)
(11)
(78)
(11)
-
(22)
(67)
-
-
-
-
(A/B, F, O)"
84
3
2
2
-
18
55
(100) 1684
-
3050
3
404
-
1
1
19
(67)
12
6
3
724
426
3
151
7
26
5
3
15
388
2
2
315
-
3
-
987
(11)
(11)
(22)
(33)
(22)
(11)
(89)
(11)
(11)
(11)
(44)
(67)
(67)
(44)
(11)
(11)
(56)
(89)
(11)
(67)
(11)
(33)
(11)
(22)
(44)
(78)
(11)
(11)
(78)
(11)
(89)
Crustacea
Arnphipoda:
Allorchestes angusta
Arnpithoe lacertosa
Ampithoe valida
Corophium acherusicum
Corophium brevis
Corophium salmonis
Corophium spinicorne
Eobrolgus
spines us
Eogammarus con fervicolus
Eohaustorius estuarius
Traskorchestia traskiana
N
2
N
6
(A/B, F, 0) 1,3
5
(A/F, 0)13
-
N
N
N
(A/O)
N
N
N
5917
99
2
3
52
7
-
(11)
(22)
(22)
-
99
15
66
49
288
(100) 4256
(67)
(11)
(89)
(33)
-
490
129
68
81
3
(11)
(22)
(33)
(44)
(11)
(100)
(100)
(33)
(67)
(33)
(11)
30
Table 2.2 Continued.
Origin/Cause
Taxa
Alsea Bay
No.m2 (OC)
Yaquina Bay
No.m2 (OC)
Bra chyura:
Cancer magister
Hemigrapsus oregonensis
Copepoda:
Hemicyclops subadhaerens
N
N
6
(44)
4
(11)
(22)
C
26
(44)
23
(67)
57
58
(78)
(67)
N
N
N
3
-
N
N
35
C
C
105
-
Cumacea:
Nippoleucon hinumensis
Cumella vulgaris
(J/B)4
N
Isopoda:
Gnorisrnosphaeroma insulare
Gnorismosphaeroma oregonensis
Lironeca californica
Macrura:
Neotrypaea californiensis
Upogebia pugettensis
Tanaidacea:
Leptochelia dubia
Sinelobus stanfordi
Mollusca
Bivalvia:
Clinocardium nuttalli
Cryptomya californica
Macoma baithica
N
N
N
N
(A/O)1
N
N
C
N
Macorna inquinata
Mya arenaria
Myseila tumida
Mytilus californianus
Mytilus edulis
Transenelia tantilla
Gastropoda:
Alderia modesta
Apiysiopsis enteromorphae
Littorina sitkana
Melanochiamys diomedea
Phoronida
Phoronis pailida
Sources: 1 Canton
1996);
(1979);
N
N
N
N
N
2
335 (100)
31
(56)
(22)
-
5
(33)
(33)
(11)
(44)
(22)
34
1
(67)
(11)
(22)
(11)
1507
15
(44)
(33)
(11)
(44)
15
55
61
(33)
(33)
185
(100)
268
(100)
6
(11)
(67)
(11)
(11)
(11)
-
-
2
1
2
28
4
2
3
-
3
1
78
27
-
36
28
-
(67)
(11)
(33)
(11)
18
-
(11)
(22)
-
7
(33)
(33)
(11)
(22)
269
(56)
140
(44)
5
6
-
8
2
J.T. Canton (personal communication
Canton and Geller (1993)
Cohen and Canton (1995);
2
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 polychaete
Yaquina Bay:
Leptochelia
C.
Hobsonia florida and in
salrnonis, H. florida and the cryptogenic tanaid
dubia
(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)
.
60 25
25
ALSEA BAY (Low Tide)
ST
___ CR
20 -
N
r
NA
- 50 20 -
U)
NI
-40
15
40
15
Ui
30
6, 1, 7,
Ui
30
10
- 20
20 Cl)
U)
- 10
10
0
I-
0
10, 1, 17, 4
Ui
>- 40
I
2
3
4
5
I
6
50
oS2
0
40
2
3
4
5
6
50
2
0
40 z
YAQUINA BAY (High Tide)
40
z
0
17,5, 26,8
CR
NA
30
NI
30
20
ST
>-
30!
9, 2, 21, 6
20
10
10
U)
z
15, 4,18, 8
20
8,1,7,2
15 4, 15, 8
')fl Ui
10 10
0
0
1
2
3
4
5
6
1
< Downstream SITES Upstream >Figure 2.2
zZ
2
3
4
5
6
Downstream SITES Upstream >
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)
(No103 m2)
CPUS
(00)
(No103 m2)
Fish Order
Species
Atheriniformes
Atherinops
affinis
parva*
Clupeiformes
Alosa sapidissima*
Ciupea paiiasii
Engraulis mordax
Gasterosteiformes
Auiorhynchus fiavidus
Gasterosteus aculeatus
Syngnathus ieptorhynchus
Perciformes
Cieveiandia ios
Cymatogaster aggregata
Hyperprosopon argenteum
Lepidogobius iepidus
Lucania
0.2
(11)
0.1
(11)
(67)
1
2
Lumpenus sagitta
Phanerodon
furcatus
Pholis ornata
Pholis schuitzi
Pleuronectiformes
Platichthys stellatus
Pleuronectes vetulus
Salmoniformes
Hypomesus pretiosus
Oncorhynchus kisutch
Oncorhynchus tshawytscha
Scorpaeniformes
Cottus asper
Leptocottus armatus
Oligocottus macuiosus
Crustacea Order
Species
Decapoda
Cancer magister
Cancer productus
Crangon franciscorum
Hemigrapsus oregonensis
Heptacarpus paiudicola
Puggetia producta
1
4016.7
-
-
-
(67)
-
0.5
(11)
102.4
2817.1 (100)
-
3.6
-
-
-
(67)
-
15.2
0.2
(78)
(11)
1.6
16.9
3008.3
(44)
(33)
(78)
0.2
18.0
(11)
(78)
(22)
1.1
1272.0 (100)
1.5
1.7
0.2
0.7
3.6
0.3
(22)
(22)
(11)
(22)
(44)
(22)
9.5
26.8
(67)
(56)
14.7
12.2
(89)
(67)
2.5
(56)
(11)
0.6
(33)
65.6 (100)
12.3
0.2
0.6
(22)
400.1 (100)
0.3
(22)
4.1
(11)
-
55.6
2.2
0.1
-
-
(44)
(67)
(11)
-
-
-
-
(89)
-
636.0 (100)
20.8
0.5
54.4
1.7
0.2
0.2
(78)
(11)
(89)
(56)
(11)
(11)
Possibly introduced with oysters or ballast water (Hubbs and
Miller 1965); 2 Intended introduction (Craig and Hacker 1940)
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
-0.40
0
NI FISH
1'-
NA FISH
NA DECAPODS
8
-0.30
11
0
- 0.20 W
- 0.10
C/)
C/)
I
2
I
3
4
6
5
I
2
3
4
5
6
ALSEA BAY (HighTide)
0.00 0
z
w
0.10 0
a
NA FISH
0.08
NADECAPODS
7
6
9
z
z
0.04
a-
I
I
2
i
3
IA 'p
4
5
0.02
Ii
.< Downstream SITES Upstream
6
0.00
1
2
3
4
5
< Downstream SITES Upstream
Figure 2.3
6
>
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 lowtide 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 upperand lower-case letters (Alsea: A; a, Yaquina: Y; y) followed by
site number. Species densities for each taxa are indicated in
Tables 2.2; A.1; and A.2.
w
40
0
z
20
Cd)
DISTANCE
40
20
0
0E
,-,
-
I
I
I
Taxa I Sites Al A2 al A3 a3 A4 Y5 Y4 y2 Yl Y2 y4 Y3 A5 y6 Y6 A6 a6
BivaJvia
r
Crustacea* y
Polychaeta**yyyYVvv y
Crustacea
PoJychaeta V YVY V
Polychaeta* "V'V
y
Bivajvia*
V
yyyyy
VVyyyyyyyyy Vv
Vyyyvyyyyyyyyy
VVVVVVV
V
Gastropoda V V V
VV
Bivalvia**
VVV
V
Crustacea** V
Phoronida
yy
V
V
vVVVV
V
V
V
VVv
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 of fishes
and decapods in seine samples. Species origin within taxa:
nonindigenous (*), native (no asterisk) . High- and low-tide sites
are represented respectively by upper- and lower-case letters
(IUsea: A; a, Yaquina: Y; y) followed by site number. The CPUE
for species within each taxa are indicated in Tables 2.3; A.3;
and A.4.
w
30
z
C-)
15
r
C,)
DISTANCE
30
15
0
nl1L1I
0
Taxa I Sites
Yl y6 y2 y4 A6 a3 a6 al Al A2 A3 A5 Y2 Y5 Y6 Y3 Y4 A4
V
Atheriniformes
Atheriniformes*
Clupeiformes*
Decapoda
I
V
'
Vv
V
7
VVVYv
777
V
VVVVVV
VVV YVY
V
VV
VVVV
Pleuronectiformes V
V
Salmoniformes
VVVYVVVVVVVYVYV VY
Gasterosteiformes
Clupeiformes
Perciformes
Scorpaeniformes
V
V
V
VV
V
VV V
V
V
V
V
yyVVyyVVVyyyyyyyy
V
V
VVVVVV
CPUE (No! 1000 m2): 7: >0-10; Y: 11-100; V:ioi-i3Ooo;
Figure 2.5
V
V
VVVVVVVVVV
V:i,00i-io,000; Y':10,OOl-37,OOO
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
B ALSEA BAY
A ALSEA BAY
25
2O
2O
15
.
Z
0
.
/
.
.
15
-
I
.
0
-.-.--------
10
24
u
5
21
18
<35
l5
25
15
5
4LINfl_r
C
124('
(°'oo)
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
B
ALSEA BAY
100f
ALSEA BAY
100
U)
80
d.
40./Lj
60
80
60
I
40
12
20i
X5,
S(,N/
C
12'
YAQUINA BAY
15
0
/Ty(%:24k
D YAQUINA BAY
100
cL
U,
0
80
z
LU
(0
z
z0
Z
60
40
(U
0
20
2512
Figure 2.7. Mean percent densities for summer assemblages of
native and nonindigenous invertebrates under various temperaturesalinity combinations. Low- and high-tide sediment samples are
included in intertidal areas of Alsea Bay and Yaguina Bay.
45
A
ALSEA BAY
B
ALSEA BAY
D
YAQUINA BAY
100
U)
z
6
z
0
z
U)
U)
24
20
21
18
<35
-
15
12 A'
C
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 =
Armandia brevis; 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 = Corophium spinicorne; Cvul = Cumella vulgaris;
Econ = Eogammarus con fervicolus; Eest = Eohaustorius estuarius;
Espi = Eteone spilotus; Gins = Gnorismosphaeroma insulare; Gpol =
Glycinde polygnatha; Hfil* = Heteromastus filiformis; Hflo* =
Hobsonia florida; Hore = Hemigrapsus oregonensis; Hsub** =
Hemicyclops subadhaerens; Ldub** = Leptochelia dubia; Lpug =
Leitoscoloplos pugettensis; Maest = Manayunkia aestuarina; Mare*
= Mya arenaria; Mbal = Macoma balthica; Mcal = Mediomastus
californiensis; Nhin* = Nippoleucon hinumensis; Nlim = Nereis
limnicola; Pcor* = Polydora cornuta; Pele** = Pygospio elegans;
Pkem* = Pseudopolydora kernpi; Ppal = Phoronis pallida; Ppla =
Paraonella platybranchia; Sben* = Streblospio benedicti; Stan** =
Sinelobus stanfordi.
47
N
Low-Mid Estuary
Correlation with Axes
.................................
2
1
Al
Salinity 0.98 -0.14
Temp -0.81 -0.51
Heter
-0.60
0.79
A3.
Y6
Ldub
Yl
A
/
Eest
A
Upper Estuary
A2 Gpo!....
Hflo*
NIimA
/
Abre
PeIe
A ASfan
A .......
Csal
A4
Eco
ACVU!HoreA A NcaIBP0
A PpIa
A PpaI
sPi
Aa
A
Mare1
Mba!
Cspi
Aent
A A PCOT*
Cryc
A
A
Salinity
Pkem
Nhin*
Maes
Axis I
A L.pug
Gins
Arnod
HfiI*
Temperature
A
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
Correlation with Axes
A3,
s........
2
1
Salinity -0.72 0.67
Temp -0.01 -0.95
Heter -0.84 -0.35
Al
Low-Mid Estuary
A2
\..
Hpret
Pveta
(Hore)
Salinity
Macrophyte D.
Pho
Axis I
Otsh
Cagg
Larm4
(Cfran)
(çmag)
Y4
Aaff
Low-Mid Estuary
(Yaquina Bay)
S
Gacu
Pste
Emor
£
A4
£
£
Asap*
Y
A5
Temperature
Upper Estuary
Y5
S
Figure 2.10
A6
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 intertidal areas sampled in Alsea Bay
and Yaquina Bay during summer 1993.
Invertebrates
Core samples
Axes
Statistics
2
Fishes and Decapods
Seine samples
Axes
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
0.27**
0.13**
0.08**
0.16**
0.13**
QQ4-H
0.93*
0.93*
Q93**
0.96**
Q9Q**
1
LC Correlation
2
0.72k
Monte Carlo test P-values: (<0.l2) ,
(< 0.10), *(< 0.05), **(< 0.01)
Variance extracted by main axes from the total variance in species
data for core samples (0.971) and seine samples (0.648)
Pearson correlations derived from species data and sample scores
that are linear combinations of the environmental variables.
1
2
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 oysterculture, 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
amphipod
Grandidierella 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 lowtide, 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 (> 22.5
0/)
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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64
Chapter 3
Trophic Contribution and Selection of Native and NonindigenouS
Prey by Native Fishes in stuarine Rearing-Habitats
G.C. Castillo
1
1,2
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.
65
Abstract
We determined the summer prey richness, prey composition and
selection for native and nonindigenous (NI) prey by native
juvenile pleuronectids (English sole: Pleuronectes vetulus and
starry flounder: Platichthys stellatus) . Study areas included
intertidal and adjacent subtidal fish rearing habitats in two
Northeast Pacific estuaries, the Alsea Bay and Yaquina Bay
(Oregon, USA) .
Major NI prey varied greatly among sites with the
polychaete Pseudopolydora kerripi, the clam Mya arenaria and the
cumacean Nippoleucon hinumensis being common in both estuaries.
Native species dominated in Alsea Bay both in prey numbers and
volumes for both fish species but in Yaquina Bay neither native
or NI prey dominated. Fish reliance on native prey seems higher
in intertidal areas of Alsea Bay and in subtidal areas of Yaquina
Bay. Unlike NI prey, the total volume of consumed native prey
increased with the weight of starry flounder in each estuary.
Intertidal CPUE OF fish were not correlated with total numbers of
prey in the fish diet. The similar richness ratios of native to
NI prey present in the fish diet and the benthos indicate that
fish do not distinguish between native and NI prey species.
Predominant selection for native or NI prey types is not apparent
by fish in either estuary. Thus, predator-prey coevolution is not
a critical determinant of prey selection by these species.
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 of NIS introductions in
Northeast Pacific estuaries include propagation of species
associated with imported oysters, ballast water release and
fouling on hull of ships (Carlton 1979; Cohen and Carlton 1998)
Potential impacts of conspicuous NIS in U.S. west coast
estuaries since the late 1980s (e.g., the clam
amurensis; the crabs
Eriocheir
Fotamocorbula
sinensis and Carcinus maenas) have
caused great concern (e.g., Kimmerer et al. 1994; Grosholz and
Ruiz 1995) . However, studies in virtually all these estuaries
lack the resolution to assess long-term changes in food web
dynamics resulting from earlier biological invasions. Although
the implications of such possible trophic changes are uncertain,
changes in prey composition and prey density can affect the
growth, survival and year-class strength of fish (Steele et al.
1970; Poxton et al. 1983). Changes in the latter population
parameters could be enhanced by further human impacts such as
pollution and habitat degradation (e.g., Cross et al. 1985;
Sogard 1994)
Anecdotal information from northern San Francisco Bay suggests
that native fish rely little on nonindigenous
(NI) prey (Carlton
1979). However, the food webs in the latter system are
overwhelmingly dominated by NIS and cryptogenic species (i.e.,
species of unknown geographic origin, Carlton 1996b; Nichols et
al. 1990; Cohen and Canton 1995) . In the Alsea Bay and Yaquina
Bay estuaries, Oregon, the juveniles of four species of native
fish: English sole (Pleuronectes vetulus); starry flounder
(Platichthys
stellatus); staghorn sculpin (Leptocottus armatus);
and chinook salmon (Oncorhynchus
tshawytscha)
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 coevolution is not a critical
determinant of prey usage, we would expect similar selection
patterns between native and NI prey types by generalist
predators, but the evidence is mixed. Prey selection by larvae of
the introduced striped bass
(Morone saxatilis)
seems higher for
at least one coevolved copepod when compared to two NI copepods
(e.g., Meng and Orsi 1991). In contrast, juvenile English sole
did not consistently select two native amphipods over two NI
amphipods (Castillo et al. In press)
Given the limited understanding on the implications of
coevolved and non-coevolved predator-prey interactions in invaded
estuaries, we address the following questions: 1) Is the richness
of native and NI invertebrates in the fish diet proportional to
their attendant richness in the environment?; 2) What are the
contributions of native species, NIS and cryptogenic species to
the food-base of native fish in terms of frequency of occurrence,
number and volume of prey?; 3) Does the number and volume of prey
vary with fish size and weight?; 4) Are daily dietary patterns in
number and volume of prey evident?; 5) Is the total number of
prey in the fish diet correlated with benthic densities of
invertebrates and with CPUE of fish in intertidal areas?; and 6)
Is the overall prey selection by native fish similar on native
and NI prey types? To address these questions we surveyed
intertidal and adjacent subtidal rearing-habitats of juvenile
English sole and starry flounder in the Alsea Bay and Yaquina Bay
estuaries (Oregon, USA, Figure 3.1) .
These surveys revealed
higher densities of NI benthic macroinvertebrates in Yaquina Bay
(Castillo 2000)
English sole and starry flounder are estuarine-dependent
species (e.g., Pearcy and Myers 1974; Monaco et al. 1990). Many
age-0 juveniles of these species recruit to shallow estuarine
nursery areas where older individuals reach highest densities
during summer before migrating to deeper habitats (Orcutt 1950;
Krygier and Pearcy 1986; Boehlert and Mundy 1987) . Juveniles of
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°/
(lower estuary) to 5°/
closer to the ocean
in upstream areas (upper estuary, Figure
3.1). A beach seine (32 mx 1.8 m and 0.8 cm stretched mesh size)
was used to collect juvenile English sole and starry flounder.
Seine was also used to collect fish from three to six subtidal
areas adjacent to intertidal sites at low-tide (Figure 3.1)
Seining area encompassed a semicircle of 163 m2 from the water
line (0 in depth) to deeper areas. The catch per unit effort
(CPUE) of each flatfish species caught with seine in a given site
was standardized to individuals per 1000 m2.
The 370 collected fish (Table 3.1) were immersed in a lethal
doses of MS-222 (200 mg/l)
.
A 10% solution of buffered forinalin
was then injected into the coelomic cavity to fix prey items
followed by fish preservation in 80% ethanol. Feeding habits of
English sole are based on stomach contents. All prey items in the
stomach and the anterior 1/3 of the intestine were analyzed in
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 (1-600
mm3)
.
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
k
is the number of fish containing prey k and n is the
total number of fish. We computed the mean prey frequency
of occurrence for all species of a particular origin
(i.e.,
native, NI and cryptogenic) in the fish diet (MO) using the
formula:
MO =
Where °kj is the corresponding °k value for prey species , of
origin.3 and m is the total number of prey of origin.
(MN)
We computed the mean number of all prey species of origin
in the fish diet using the formula:
[Nklj]
Mn = n_li
Where NkI is the number of prey species k of origin
for a total
prey species of origin
and for
in fish
fish analyzed.
Likewise, the mean prey volume of all species of origin
(MV) in
the fish diet is computed by substituting Nkl by Vkj! in the
previous formula.
We defined the overall relevance of prey item k to the fish
diet by the index of overall item contribution (OIC)
OICk =
[Ok
+ %Vk
+ %Nk]3'
which we defined by the percentages of prey frequency of
occurrence (Ok), volume (%Vk) and number (%Nk)
itemk. The OIC index can range between 0%
in all fish examined) to 100%
contributed by
(i.e., itemk is absent
(i.e., only itemk occurs in all
73
fish examined) . We used the OIC index to rank up to 15 dominant
items in the fish diet. Estuaries are divided into three sections
to determine the OIC values in fish from downstream to upstream
sites: lower estuary (sites 1-2), mid estuary (sites 3-4), and
upper estuary (sites 5-6)
Bivalves and their siphons are considered different items
since siphon cropping by fish is more predominant than
consumption of entire bivalves. Polychaete parts of unknown
species, plant matter and pieces of woody debris are also
included as prey items. Unidentifiable organic matter comprised
less than 4% of the mean percent volume and is not included in
the OIC index.
We estimate prey selection of macroinvertebrates in the fish
diet with Johnson's (1980) selection index(E1):
= n1
[r
s]
where rjj is the rank of usage of prey item
the abundance of prey1 in the fish diet),
availability of item
to fish
by fish
(based on
is the rank of
(based on the benthic density of
item1) and n is the total number of fish. The more used and/or
available an items is, the closer to one is its average rank.
Hence, the most selected item has the lowest
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 are compared
with t-tests for unequal variances using Statgraphics 2.1 Plus.
Differences in the proportions of native and NI prey between low
and high-tide were compared by X2 tests.
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
Estuary
English sole:
Alsea Bay
Yaquina Bay
Starry flounder:
Alsea Bay
Yaquina Bay
Weight
Mean
Length (cm)
Mean
SE
Number
of fish
(g)
SE
135
113
6.8
7.1
0.1
0.1
3.0
3.4
0.1
0.2
61
61
9.0
13.3
0.6
0.6
14.9
34.4
3.1
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
Species
No. Species
Environment
Native
NI
No. Species (Occurrence)
Fish Diet
Native
Alsea Bay
English sole
Starry flounder
Both species
31
30
34
8
18
8
8
Yaquina Bay
English sole
Starry flounder
Both species
49
50
50
11
11
11
8
NI
(13.3)
(22.1)
5
20 (16.0)
5
20
5
9.0)
(12.8)
(10.2)
(
(14.5)
(17.9)
7.8)
8
16 (10.7)
8
25
9 (15.7)
(
(
8.8)
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
Yaquina
Alsea
F
S
F
Origin /Vector
S
B iva lvi a
Clinocardium nuttallii
Cryptomya californica
Macoma baithica
Mya arenaria
Myselia tumida
Transenneiia eantilla
10.4
2.2
3.0
17.8
3.0
3.3
18.0
3.3
0.9
1.8
0.9
4.4
3.3
1.6
13.1
23.0
1.6
0.9
N
N
N
A/O'
N
N
Crustacea
- Araphipoda:
Aliorchestes angusta
Ampithoe lacertosa
Ampithoe valida
Corophium acherusicum
Corophium saimonis
Corophium spinicorne
Corophium brevis
Eobroigus spinosus
Eogammarus confervicolus
- Cumacea:
Cumella vulgaris
Nippoleucon hinumensis
- Decapoda:
Neotrypaea californiensis
Crangon franciscorum
Hemigrapsus oregonensis
Upogebia pugettensis
- Tanaidacea:
Leptochelia dubia
Pancolus californiensis
Sinolobus stanfordi
N
1.8
1.6
0.9
2.7
37.2
9.7
7.1
1.8
4.4
3.3
70.5
31.1
45.2
87.0
31.1
2.2
9.6
1.6
47.4
16.3
1.6
6.6
22.1
38.1
4.9
18.0
1.6
2.7
0.9
9.8
4.9
4.4
5.2
2.2
0.7
2.2
3.0
12.4
0.9
3.0
N
A/(B;F;O)1'4
3.3
1.6
A/(F;O)"4
N
N
N
A/O2
N
N
J/B3'4
1.6
N
N
N
N
3.3
C
C
C
77
Table 3.3. Continued.
Estuary
Yaquina
Aisea
Species
E
S
E
Origin /Vector
S
Polychaeta
Amaeana occidentalis
Armandia brevis
Boccardia proboscidea
Dorvillea annulata
Eteone spilotus
Eteone californica
Glycinde polygnatha
Glycinde armigera
Heteromastus filiformis
Hobsonia florida
Manayunkia aestuarina
Mediomastus californiensis
Nephtys caeca
Nereis limnicola
Owenia fusiformis
Paraonella platybranchia
Phyllodoce hartmana
Pseudopolydora kempi
Pygospio californica
Pygospio elegans
Streblospio benedicti
39.3
4.4
3.3
4.4
9.6
1.5
0.7
1.5
1.5
57.4
N
N
N
4.4
4.4
5.3
0.9
16.8
0.9
0.9
0.9
1.6
4.9
2.7
4.4
2.7
49.2
16.4
3.3
C
C
N
N
A/(B;O)1'
11.5
67.2
0.9
1.6
0.7
6.7
1.5
17.8
3.3
1.6
N
0.9
25.7
36.1
10.6
39.8
1.6
24.6
A/ O?)2)
C
N
N
N
C
N
N
P/(B;F';O)"4
N
C
Teleost
CymatogasLer aggregata
Engraulis mordax
1.6
1.6
1.6
N
N
Sources:
1
Canton (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
40
English sole (Alsea)
a
English sole (Yaquina)
B
No. Fish = 113
No. Fish = 135
40
30
30
20
20
>-
w
u-.=
10
10
NA
50
z
w
C
b
C
NI
CR
Starry Flounder (Alsea)
0
D
Starry Flounder (Yaquina)
No. Fish = 61
40
40
30
30
20
a
E
No. Fish = 61
a
20
10
10
0
a
b
NA
NI
C
CR
English sole (Alsea)
0
NA
NI
CR
English sole (Yaquina)
F
a
30
20
0
G
Starry flounder (Alsea)
H
250
250
200
200
150
150
100
100
50
50
0
0
Starry flounder (Yaquina)
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 (1.1)
NI
CR (0.4)
-ST (0.5)
CR )0.9k,
ST (0.2)
NI (1.1)
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 kempi
and 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 polychaetes
Pseudopolydora
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 and
Hobsonia
florida.
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)
*
brev7s
Ohgochaeta
-n
ChironoiTidae
0
(7'
0)
0)
R
-i
-I
I
I
Oligochaeta
* * elegans Pygospio
I
garisvu! Cumella
rmtter Rant
Rant
spinicorne Corophium
rrtter
vulgar/s Cumella
Pseudopolydora
peces Wood
* kempi
(s) balfhica Macoma
salmon,s Corophium
(7,
4
Harpacticoida
*
* flonda sofia Hob
kempi Pseudopolydora
sp Capitella
parts FIychaeta
0)
(71
1)J
p)
5)
5)
C.)
0
(71
-4
(7)
(71
0
(7 01
5)
0
5)
5)
(7
vulgaris Cumella
elegans Pygospio
0
VOLUME %
Co
spilotus Efeone
-4
I
(7
**
Ostracoda
I'.)
(71
I
salmonis Corophium
0
(Il
Harpacticoida
I
(71
-J
NUMBER %
(71
hinumensis Nippoleucon
I
californiensis Neotrypaea I
pieces Wood
Iimnicola Nereis
(71
benedicti Streblosp!o
h,numensis Nippoleucon
*
0
*
I
0
0
spilotus Eteone
(4
parts Rlychaeta
*
ta/is 0cc/den Amaena
(71
0
C)
benedict! Streblospio
hinumensis Nippoleucon
*
-4
C)
(71
N.)
(Cl
-4
0
salmonis Comphium
I
0
(7
(7
N.)
(71
brevis Corophium
NJ
J
pieces Wood I
0
(71
Harpacticoida
I
(7,
.4
01
dubia Leptochelia
(s) ba//h/ca Macoma
parts Fklychaeta
oscidea prob Boccardia
**
0
0
rrtter Rant
flu/ta/li! Clinocardium
fervicolus con Eogammarus
*
* * elegans Pygospio
hinumensis Nippoleucon
arenaria Mya
Otgochaeta
parts CapeIlid
*
parts Fklychaeta
vulgaris Cumella
salmonis Corophium
(7)
N.)
VOLUME %
0
brevis IArmandia
pieces Wood
I
(71
0
(s) balthica Macoma
(71
-J
Harpacticoida
I
0
0
NUMBER %
t8
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)
0
0
Pe
Ce
0
O
4
Ce
I Fant Paper
0
C"
A
Pa 0
a
C-,
cS
S
Ce
0
0
Coroph/um salmonis
0
Ce
-1
Ce
0
0
t-lobsonia florida *
Macoma balm/ca (s)
0
Ce
1
Ce
0
M'a arenaria *
Corophium salmon,s
Pseudopol1eiora kempi *
Ce
PS
Harpacticoida
Corophium spin/come
ttani nailer
A
Ce
Mya arenana *
CS
Ostracoda
I Rtychaeta parts
10
I Eteone spllo/tus
Hobsonia florida *
I Cap/telIld parts
I I-'arpacticolda
Corophium spin/come
Neotrypaea Californiensis
0
klychaeta parts
Macoma barth/ca
Ce
Ce
Crangon franciscorum
Macoma bait h/ca
pant trailer
ciiir000rridae
j j Nereis limnicola
Nippoleucon hinumensis *
Macoma balihica (s)
-4
I Wood pieces
0
z
0t
Ce
tiers/s limnicola
Ce
P-i
(I
I Wood pieces
L
0
a,
b2
(0
P4
Ce
P-i
Ce
I
% NUMBER
I Manayunkia aestuarina
0
Ce
p
0
0
0I
Ce
VOLUME
I Karpacticoida
Ce
A
Macoma baithica (s)
P-i
Ce
A
C.S
P-S
0
Ce
'S
m
m
(I,
--4
a,
Ce
A
(S
PS
%
Ce
Corophium sp/nicorne
Hobsonia florida *
Ce
PS
I Wod pieces
0
Cam siphon
al Cyclopoida
A
Ce
Macoma baithica (s)
,-aMyaarenor/a *
'S
PS
I Ostracoda
Ce
Macoma ba/thica
I Wood pieces
Corophiurn spinicorne
m
-v
C
F-rpaclicoida
f I Dptera pupae
Ce
-4
a,
0
Core ph/urn salmon/s
Nereis limo/cola
PS
Ce
Cttironorridae
g
0
g(j Hobson,a florida *
¶
PS
0
% NUMBER
Ce
Ce
-4
0
U
VOLUME
A
'C
PS
0
1
%
CJ
vi
Ce
0
C)
Ce
A
CS
PS
0
Ce
Ce
-1
Ce
Ce
A
C-i
PS
Ce
A
C-,
P-i
a
Ce
Ce
-S
Ce
Ce
A
PS
0
Ce
°
¶
.
B
a
Ce
at .,
Ce
P-i
...t
to
PS
0
Ce
0
0
Corophium salmon/s
I Osiracoda
Harpacticoida
Cyrrialogaster aggregate
I ianl patter
IMacoma barth/ca (S)
I Wood p/eces
F'seudopolymiora kempi *
lHobsoniaflorida *
Nip poleucon hinurnens/s *
0
Ce
-S
Ce
0
0
Leptochelia dub/a ..
Macoma ballhica (s)
5nt riot/er
Ce
PS
Eobro/gus spinosus *
J Mia arenaria *
f Llpogebia pugettensis
I Nippoleucon hinumensis *
Pseudopolysiora kempi *
I Pygospio elegans..
Chnocardium flute/I/i
I Neotiypaea cal/fern/ens/s
I Eob re/gus sp/trosus *
l-rpacticoida
Ffe/ychaeta parts
Engraulis mordas
Macmyra ba/thica
I
0
4
Ce
(.1.
0Ce
Myaarenaria(s) *
I
PS
Ce
I Cryptomya cal/fornica
[
0
% NUMBER
jjI Macoma ba/thica
VOLUME
C) - C-'
C-,
Ce
4
%
vi
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 sizerange is also indicated (SE = standard error)
Alsea Bay
Taxa
Origin
Total size range (cm)
NI
All
Harpacticoida
All
Macro-Crustacea
Native
NI
Cryptogenic
Supra-specific
All
Polychaeta
Native
NI
Cryptogenic
Supra-specific
All
Miscellaneous
All
All taxa
5
SE
No. prey /fish
Total size range (cm)
(7.1-11.0)
50/64
(7.1-11.0)
60/67
(3.0-7.0)
48.8
0.1
48.9
25.3
2.2
27.5
5.2
0.2
5.4
13.0
38.9
44.1
68.9
15.1
3.9
0.6
0.0
11.3
2.5
8.5
0.6
0.1
11.7
17.3
15.7
9.0
0.3
42.2
2.9
10.0
0.5
(3.0-7.0)
66/68
Bivalvia
Native
Yaquina Bay
1.4
5.9
0.4
0.7
0.4
12.7
40/49
0.2
13.2
1.0
0.6
4.5
0.1
6.2
11.9
0.2
13.9
0.1
13.5
7.5
19.8
1.1
0.3
28.8
0.1
1.8
0.5
0.7
100.0
100.0
100.0
100.0
108.7
13.3
24.4
3.3
64.9
11.8
57.4
9.7
0.8
1.0
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
Size interval (cm)
Taxa
Origin
(3.0-7.0)
31/31
Bivalvia
Native
NI
All
(7.1-11.0)
13/13
Yaquina Bay
Size interval (cm)
(11.1-24.0)
14/17
(3.0-7.0)
14/15
(7.1-11.0)
9/9
(11.1-24.0)
34/37
0.9
0.0
0.9
1.4
0.0
1.4
8.7
15.4
24.1
4.5
2.1
6.7
37.2
2.0
39.2
6.2
2.1
8.3
71.4
28.2
12.9
2.8
4.0
0.1
22.4
0.3
0.0
0.0
22.7
61.7
0.0
0.0
0.0
61.7
44.1
1.0
0.0
0.0
45.0
17.5
4.5
4.4
1.0
27.4
25.1
1.0
0.0
0.0
26.1
60.5
0.9
1.5
0.8
63.8
2.3
2.1
0.0
0.0
4.4
4.9
3.5
0.0
0.0
8.5
7.7
5.1
0.3
0.0
4.9
23.2
0.0
0.7
28.9
1.0
26.1
13.2
3.8
57.0
1.0
0.0
61.8
0.7
0.2
4.8
1.4
1.7
0.5
100.0
100.0
100.0
100.0
100.0
100.0
67.3
19.0
32.8
8.1
22.2
3.6
123.5
24.1
77.0
21.7
51.2
9.8
Harpacticoida
All
Macro-Crustacea
Native
NI
Crypt o geni c
Supra-specific
All
Polychaeta
Native
NI
Crypt o geni c
Supra-specific
All
Miscellaneous
All
All taxa
No. prey/ fish
SE
0.0
0.2
27.3
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,
x2 test; p < 0.001)
.
C,
G,
E;
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 significant linear
relation (P < 0.05)
A
English sole (Alsea Bay)
600
(1:AlI taxa, R2 <0.00)
(2:Native, R2 = 0.09)
Starry flounder (Alsea Bay)
B
All taxa, R2 = 0.43*)
Native, R2 = 0.44*)
3000
n28
(3:NI,R20.01)
n29
400
'
NI, R2 = 0.03)
A
2000
,
C.,
E
1000
200
I
2
3
uJ
0 123 4 5 67
600
0
-
0
8
English sole (Yaquina Bay)
C
>Ui
3
D
3000
(I:A!I Taxa, R' <0.00)
(2:Native, R2 = 0.03)
20
400
2000
80
100 120
Starry flounder (Yaquina Bay)
(1:All Taxa, R2 = 0.16*)
'
A
(3:NI, R2 = 0.02)
60
40
(2:Native, R2=0.15*)
(3:NI, R2 = 0.03)
n41
'I
200
U
U
A
1000
.
U
2
'3
2
3
20
40
TOTAL FISH WEIGHT (g)
Figure 3.6
60
80
100 120
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- and high-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
200
English Sole (Alsea Bay)
Ii.' NA
NI
I
I
150
>LU
B
200
b
85.7 (n 29
CR
ST
150
0
LL
11.9n 106)
'-'Cl)
50
zoZ60
I
NI
EJ CR
a
10.6 (n= 28
LU
E
6O
ST
40
n 18
20
20
0
LOW
HIGH
80
NA
ST
40
h.iiIjiIhii(I:
D Starry Flounder (Yaquina Bay)
Starry Flounder (Alsea Bay)
a
5.7 n= 33
1.2 n=53
LOW
HIGH
LOW
mw C
b
0.9 n 60
IIIIIIIIIIIIIIIII
a
50
LUJJNA
E.iiii NI
CR
ST
100
100
a:
English Sole (Yaquina Bay)
0
HIGH
F
English Sole (Alsea Bay)
a
IIIIIIIIIIIIIIIII
LOW
HIGH
English Sole (Yaquina Bay)
0
100
DIIJNA
CR
ST
100
JCR
75
LU
NA
Nl
ST
b
9.5
50
o.
>Li
a
4.4
25
JJIIIIIIIJHIIUI
LOW
G
HIGH
Starry Flounder (Alsea Bay)
I
E 400
IIIIIIUIIIIIIIII
I NA
NI
H Starry Flounder (Yaquina Bay)
NA iINI
ICR
ST
400
<300
Ui
300
200
200
100
100
LOW
ICR
a
1.7
LOW
HIGH
TIDE
Figure 3.7
0.9
HIGH
ST
94
Figure 3.8. Mean number of invertebrates in the flatfish diet (AD), 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
250 A
250: B
EnglishSole Diet
200
-
All Taxa
NatIve
50
--0- Native
- - - Nonlndigenous
100
100.
-
>-
-_
All Taxa
200
- Non-IndIgenous
//
150
W
English Sole Diet
-
-___._&
I
3
4
6
5
200-
Starry Flounder Diet
C
- O-
All Taxa
NatIve
------
23456
0
0
Q200:
==
50
Starry Flounder Diet
.--AllTaxa
D
---c-- NatIve
150
- .- Non-Indigenous
Non-Indigenous
100
...........
50
50
0
'0
3
50 E
30
Z
20.0
I-
6
--
n
3
4
5
6
All Taxa
- a - NatIve
-.
Non-indIgenous
10
56
0
60- G
C
We
\
-
Intertidal Benthic Invertebrates
0 - Native
- Non-indigenous
3
E
2
All Taxa
0
N
1
F
Intertidal Benthic Invertebrates
N40
>-
4
-.
0
5
60- H
Intertidal Fish
50-
-.--
40
50-
gIlsh sole
Starry flounder
Intertidal Fish
bigllsh sole
Starry flounder
40
0.
30=
0 20-
20-
z
10-
0-
0-
I
4
5
6
5
2
ALSEA BAY SITES
YAQUINA BAY SITES
UPSTREAM -*
UPSTREAM -*
Figure 3.8
6
96
(A) Are individual prey selected differently? If so,
6.
(B) are
native and NI prey types selected differently?
Yes, prey selection varied among individual prey in the
diet of each fish species in Alsea Bay (Figure 3.9) and Yaquina
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 diet of
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 flatfish in 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)
-2
-3
0
-1
2
1
3
0
oleucon hinurnensis*
2
Cumella vu!
z
4Pseudo
dora kern
Eteone spilotus
Pygospio e!egans**
Capite!!a sp.
8
Corophium sa/rnonis
10
10
B
6
4
2
RANK AVAILABILITY
8
2
4
6
RANK USAGE
0
8
10
JOHNSONS SELECTION INDEX (STARRY FLOUNDER ALSEA BAY)
-3
-2
0
-1
2
1
3
Di .tera
a californica
Pseudo 0! dora kem.i
z
.r. phium salmonis
N...l
hi
Coro hium s inicorne
0
'0
Nereis /irnnico/a
Macoma baithica
Chironomidae
Macorna ba/thica S
10
Hobsonia florida*
12
10
6
4
2
RANK AVAILABILITY
8
0
Figure
2
3.9
4
6
8
RANK USAGE
10
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)
-6
-4
-8
-2
8
0
2
4
6
0
Cumella vulgaris
2
Chironomjdae
Eteo e
4
Il.
a balthica (S)
S
6
Ni000levcon hinumensis
Hnhcnni flnriria*
Fngammanic rnnfprvirnlus
8
Corophium brevis
10
Efnne piIrttc
I
12
pfncheIi di,hi
Boccardia Droboscideá
Streblospip benedicti*
Corophium sninico
Pseudopolydora kern pi*
Pyqospio eIejans1
14
16
Mediomatu.s ca1ifnmincis
18
Capitella sp.
Corophium salmonis
20
OIigQc haeta
22
15
10
5
RANK AVAILABILITY
10
5
RANK USAGE
0
15
JOHNSON'S SELECTION INDEX (STARRY FLOUNDER YAQUINA BAY)
-4
6
-6
-2
4
0
2
B
0
1
NentrvneR c.Iifnrniencjs
Mvi aranari*
Cumella vulaaris
2-
46810
12 14 16
ra kern pi*
CoroDhium sninicorne
Nereis limnicala
Eteone snilatus
Streblospio benedicti*
Macoçna haIfhiri 1S
CaDitella SD.
Corophium sa
NippoIiirrr1 hinnmncLc
*
Hobs
Macama baithica
Manay
18
15
10
5
RANK AVAILABILITY
0
Figure 3.10
5
RANK USAGE
10
15
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 are alike 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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sand sole in East Sound, Orcas Island, Washington. Journal of
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patterns of movement in juvenile estuarine fishes. In pages
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Monaco, N.E., D.M. Nelson, R.L. Emmett, and S.A. Hinton 1990.
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California Press, Berkeley and Los Angeles, California. 405
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Fishes 5:383-390.
.
113
Chapter 4
Predation on Native and Nonindigenous Pamphipod
Crustaceans by a Native Estuarine-Dependent Fish
G.C. Castillo
1
1,2
H.tL 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.
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.
acherusicum
and
C. insidiosum) . Single-species prey consumption in sand
substratum was greater on C.
spinicorne
and 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,
Corophiurn brevis, 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 2mm
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 L
is the selection for prey of size
proportions of prey of size
;
r1 and p are the
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 singlespecies 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
2
MALES
FEMALES
C. salmonis
C. salmon is
ii
V 0.758 + 7.386X
(R2 0.80,38 d.f.)
Y = 2.269X°4122 + 2.269X°4116
(R2
0
00
0.5
0.81, 37 cli.)
1.0
1.5
2.0
00
0.5
1.0
1.5
2.0
(R2
00
0.5
0.87, 39 d.f.)
1.0
2.0
1.5
C. acherusicum
C. acherusicum
V = 1.785 e14
Y= 1.396 + 1.797X
(R2= 0.80,39 d.f.)
(R2 0.40, 39 d.f.)
0
0
00
0.5
1.0
1.5
2.0
00
Y= 0.429 + 6.113X- 3.848X2
(R2= 0.69, 38 d.f.)
2.0
1.5
1.0
6
Y.2.05 + 25.03X -31.45X2
(R2= 0.52, 39 d.f.)
0
0
00
0.5
C. insidiosum
C. insidiosum
6
2.0
1.5
y .0.336+ 6.916X- 1.451X2
Y= 0.579 + 4.891X - 1.31 1X2
(R2= 0.86, 39 d.f.)
00
1.0
C. spinicorne
C. spinicorne
2
0.5
0.5
1.0
1.5
2.0
00
ANTENNA LENGTH
0.5
1.0
2.0
1.5
(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
single-species experiments is shown in each case.
(
I
)
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)
(3:00-9:00) a.m.
(7:00-12:00) p.m.
Species
WA
SW
PV
4
0.17
0.15
0.00
0.00
0.17
0.15
1.63
0.35
2
0.57
0.31
0.07
0.11
0.40
0.26
3.37
2.06
1
1.43
0.40
0.87
0.31
0.10
0.10
3
0.23
0.15
WA
SW
PV
0.53
0.21
0.03
0.06
0.30
0.10
2.70
2.25
0.37
0.21
2.40
0.95
0.57
0.06
Rank
(12:00-6:00) p.m.
WA
SW
PV
4
0.37
0.06
0.00
0.00
0.40
0.10
2
0.53
0.42
2
0.23
0.11
0.00
0.00
0.37
0.15
3
0.17
0.12
1.17
0.38
1
1.00
0.35
0.03
0.06
1.23
0.55
0.23
0.15
0.00
0.00
3
0.13
0.15
0.03
0.06
0.00
0.00
Rank
Rank
C. salrnonis
mean
sd (±)
C.
spinicorne
mean
sd (±)
C.
acherusicum
mean
sd (±)
C.
insidiosum
mean
sd (±)
4
123
B UNDISTURBED 24°C
A UNDISTURBED 14°C
12-
12
MALES
Ii'
FEMALES
U)
L()
E
SAL
C)
C
0
SPI
ACH
SAL
INS
DISTURBED 14°C
12-
8
8-
4
oiLI iiI .
SAL
SPI
ACH
i
INS
ACH
U
INS
DISTURBED 24°C
D
12
SPI
b
SAL
ii
SPI
ACH
I
INS
SPECIES
Figure 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 (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 (Figure 4.3:
A) irrespectively of sex (P < 0.05, ANOVA; Figures 4.3: B and C),
with predation on C. spinicorne and C. acherusicum nearly twice
as high as on C. salmonis and C. insidiosum (95% HSD test)
Significant differences in prey survival between treatments and
controls occurred in all species except C. sairnonis (1-tailed
test, P < 0.05, Figure C.2) .
X2
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-tailed
X2
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-tailed
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-
X2
125
A (MALES + FEMALES)
24
2016-
a
12-
T
80
SAL
B
SPI
ACH
INS
MALES
24
20
16
b
12
b
a
8
1
4
0
SAL
SPI
ACH
INS
FEMALES
2420
16-
b
12-
1
8-
b
a
40
1
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.
C. salmonis
0.10-
S
-0.10
r = 0.04 P >0.80
0.05-
.
I
0.00-
-0.05-
0.1 0
r-0.02 P>0.90
0.05-
I
.
.5
0.00-
S
S
-0.05-
I
02000.81012 14
.0.10
0.2 0.4 0.6
0.10-
0.10-
r=0.37 P=0.10
r0.26 P>0.40
0.05-
0.05-
0.00-
0.00- __
.0.10
0.8 1.0 1.2 1.4
C. insidiosum
C. acherusicum
-0.05-
spinicome
.
.
Is
-0.05S
r
02 04
0.6
0.8
1.0
12 14
-0.10
0.2 0.4 0.6
0.8 1.0 1.2 1.4
ANTENNA LENGTH (mm)
Figure 4.4. Strauss' selection index by prey size (4th article
2nd antenna) and by Corophium species consumed by Pleuronectes
vetulus.
127
Corophium salmonis
4O
Not Eaten (x = 0.42, se 0.01, n = 179)
Eaten
(x = 0.42, so = 0.01, n = 61)
3O
2010
I
0.2
I
04
03
Fii
0.5
Inn
0.6
0.7
-
0.8
0.9
1.0
1.1
1.3
1.2
Corophium spinicorne
40
Not Eaten (k = 0.64, se = 0.02, n = 57)
(x = 0.65, so = 0.02 n = 183)
Eaten
30
20
10
., .
0.2
0.3
In
0.4
I I I I
06
05
LI
1HIH
0.7
0.8
fl
0.9
Ill
1.0
H
1.1
1.2
1.3
Corophium acherusicum
40
Not Eaten (* = 0.40, so = 0.03, n = 57)
Eaten
(* = 0.60, se = 0.02 n = 183)
30
20
10
0.2
03
I.
0.4
40-
i
0.5
.
.
I
.
0.6
0.8
0.7
0.9
H In
1.0
1.1
1.2
1.3
Corophium insidiosum
Not Eaten ( = 0.34, se= 0.01, n = 151)
30-
(x = 0.35, se = 0.01, n = 89)
Eaten
20-
ID0
1-IFF
02
03
0.4
05
I
0.6
-
ri
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.
WA
(3:00-9:00) a.m.
SW
PV
WA
SW
PV
(12:00-6:00) p.m.
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
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
Mud
129
SAND
A
MUD
(MALES + FEMALES)
B
5
b
4
I
3-
3-
2
2
.c
(MALES + FEMALES)
b
b
ACH
INS
1-
SAL
SPI
ACH
SAL
INS
MALES
C
SPI
MALES
D
5-
4-
4-
3-
3
2
a
2
b
I
SAL
SPI
ACH
SAL
INS
FEMALES
E
5-
54-
ACH
INS
FEMALES
F
4-
SPI
3-
2
a
2
a
-r
I
i-
SAL
SPI
ACH
INS
bc
ab
a
SAL
SPI
C
rrI
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)
2010-
I
0.2
03
0.4
.n
.11 .
05
0.8
0.7
0.6
0.9
1.0
1.1
1.2
1.3
Corophium spinicorne
Sand(x0.70, se
81)
0.02, n
Mud (x = 0.72, so = 0.01, n = 127)
>-
10-
0
0
U-
o
40
I-
30
LU
20-
z
0.3
II
.
.n
0.2
05
0.4
I
H .H
0.7
0.6
I I
In.nln
In
0.8
0.9
1.1
1.0
1.2
1.3
1.2
1.3
Corophium acherusicum
Sand (x = 0.43, se = 0.03, n = 50)
Mud (x = 0.46, se = 0.02, n 95)
C)
LU
10
0.2
03
0.4
I
05
IR.
H
0.6
0.7
H. I
0.8
0.9
1.0
1.1
Corophium insidiosum
Sand (
0.34, so = 0.01, n = 75)
EJ Mud (* 0.36, se = 0.02, n = 93)
10-
0.2
03
I
0.4
0.5
Ir in
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. acherusicum in 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 used in
our study. The rank of predation in single-species experiments
was not consistently related to species origin as seen in mixedspecies 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 composition of
the prey populations.
The two most consumed species in single-species experiments
(C. spinicorne and C. acherusicum) were the most visible over the
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 for both
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 sparid
lithognathus),
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.
tshawytscha
and 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 remained in a fixed area on the bottom
or partially buried in the sediment for several hours.
Preliminary observations in which prey were added from the top of
the tank showed eye movements in the settled fish. Once the prey
reached the bottom of the tank, the fish used 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 to successfully ingest
large C. .spinicorne of either sex. In one case a large C.
spinicorne male oriented its second antennae sideways when
confronted by a fish, a defense mechanism also reported by
Reimers et al. (1979)
Foraging juvenile and adult P. vetulus use their pointed
snouts as a shovel to extract infauna (Ambrose 1976; Hulberg and
Oliver 1978) . Despite such behavior, the percent of the tank
bottom area free of sand at the end of the experiment was not
25)
correlated to total prey consumption (r = 0.17, P > 0.05, n
suggesting that overall sediment disturbance by fish does not
enhance prey consumption.
Juvenile P. vetulus seem visually oriented predators.
Stomach
fullness of juvenile P. vetulus increases throughout daylight
hours in nearshore areas (Rogue and Carey 1982) and estuaries
(Castillo 2000) .
Fish-induced turbidity in the mud treatment
reduced visibility and could explain the relatively low predation
on Corophium in mud.
Due to turbidity, the tank bottoms of most mud replicates were
not visible from the water surface at the end of the experiments.
However, predation on amphipods was not significantly higher in
mud replicates where the bottom was visible on completion of the
experiment (mean consumption = 7.1 prey, n = 7 fish) than in
tanks in which the bottom was not visible (mean consumption
prey, n = 18 fish, P > 0.50, ANOVA). Thus, differences in
5.8
134
amphipod behavior between sand and mud as well as reduced prey
visibility in mud could account for the greater consumption by
fish in the sand replicates.
Although the Strauss' prey size selection index was not
significantly correlated to prey size for any species, C.
acherusicum showed the highest correlation (Figure 4.4) . Thus,
this index is partially consistent with the significantly greater
size distribution of consumed prey for the latter species (Figure
4.5).
Our experiments support field observations in the sense that
some NI invertebrate prey can be highly selected by juvenile P.
vetulus (Castillo 2000) . With the exception of C. insidiosum,
field data for Yaquina Bay and Alsea Bay estuaries have shown
that P. vetulus and starry flounder (Platichthys stellatus) prey
on all Corophium species considered here (Castillo et al. 1996)
The low abundance of NI Corophium in the fish diet in that study
may be due to the high intertidal distribution of introduced
Corophium. Although direct negative trophic effects of NI
Corophium spp. on P. vetulus are not implied from our study, we
can not predict the effect of an increasing number of NIS on the
abundance of P. vetulus. Yet, additional guilds in the food-base
of native predators could lead to a decline in community
stability (Castillo et al. 2000)
135
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138
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
1
H.W. Li
2
P.A. Rossignol
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 polychaetes
P.seudopolydora
Hobsonia florida
kempi and the cumacean
and
Nippoleucon 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 polychaetes
florida and
Pseudopolydora
Hobsonia
kempi 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 midsection 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 of Yaquina 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 selfregulation 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);
benthic fish guild was present in all models;
(2) A predatory
(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
Mode
Vertical
Range
Guild
Activity
Trophic
Bivalvia:
Su 1,2
Cryptomya californica
Macoma baithica
U1
Mya arenaria 1-
U 1,33
ab
ab
ab
u'
ab
Mll
Sb 2,
ab 6
ab 6
Sil 6, 7
Sil 6,7
Sr
F 6
a6
Ml2 6
Sr'
u 6,35
ab
Mu
J-iarpacticoida
F
ab 36
Msl '°
Nippoleucon hinumensis t
F 6
ab?
Msl 6,12
Sr? 12,13
T 1,6
ab '
ab '
Sb 1,15
ab? 17
ab '
Mi5
Mi6
Mi2
M14
a6
Si2 6,20
U 1,33
1
Sr 3'4;Su 31
Su ';Sr 29
Crustacea:
Neotrypaea californiensis
Corophium salmonis
Corophium spinicorne
Eogammarus con fervicolus
Eohaustorius estuarius
Folychaeta:
Capitella sp.
Capitellidae
Eteone spilotus
Heteromastus filIformis t
Hobsonia florida t
(T;F)6
(T;F)6
(T;U) 17
F6
T 6
T 6
9
14
12
17
19
6
8;Su
29
Sr 6;Su 29
Sr? 30
(Sb;Su) "
Sb 12
Ip? 16,17
Sb '
Sr 21
Table 5.1. Continued.
Taxa
Manayunkia aestuarina
Mediomastus californiensis
Nereis limnicola
Paraonella platybranchia
Pseudopolydora kempi t
Pygospio elegans *
Streblospio benedicti t
Tharyx sp.
Miscellaneous:
Nemertea
Oligochaeta
Life
Mode
Vertical
Range
T 6,37
T 6
T '
a6
Ssl 6,20
Sr 22.23; Su? 17
b 24
Mi3
M13
Mi4
Sl2
Si2
Sb 24
a
a
T 6,38
T 6.17
a
a
a
T 6,17
F 6,23
(F;U)
Trophic
1
ab
(U;F)
'
T 27;U 40;F
Guild
Activity
26
17
Si2 6
(Mi7;Sil) 17
24
17
(a;b) 32
a 39;b
24
25
17
20
20
12
Sr'
Sr? 17
(Sr;Su) 26
Sr 3;Su 17
(Sr;Su) 24
Sr? 17
M1327
Ip'
M18 28
Sb'2
Sources: 1 Rudy and Hay (1991); 2 Peterson (1977);
Brey (1991);
Tunnicliffe
Posey (1987); 6 Personal Observation;
and Risk (1977);
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);
Zwarts and Wanink (1989);
32 Blake (1993);
Hedgpeth (1975);
Meador et al.
Light (1969); 38 Lindsay and Woodin (1996);
(1993); 36 Yingst (1978);
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
Fn = -1
0
0 -1
C
-1
0
Links 2
Loops = 1
Fn = 1
Links = 1
0 -1
D
Loops =0
00
Fn = 0
EOJ
Links =2
Loops = 1
Fn = 1
Links = 1
Loops = 0
F
Fn =0
G
010
-1
-1
-1
010
Figure 5.1
Links =5
Loops =3
Fn = 0
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 depthrange 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 = surfacesubsurface)
P represents fish predators. Asterisks denote both non-invaded communities and
native guild. Invaded communities and introduced guilds lack asterisks.
.
No.Guilds
Community
Structure
Guild
Assumptions1
Al *
All
Alli *
A
A
3
B
6
AIV
B
10
AV
B;C
11
AVI*
D
11
AVII
D
16
AVIII
D;E
22
Basic Guild Structure
(Variables)
5
*; M*; 5*
*; M*; S*; M; S
*; Ml*; Mi*; Ms*; 5j*; Ss*
*; Ml*; Mi*; Ms*; 5j*; Ss*;
Ml; Ms; Sl; Si
*; Ml*; Mi*; Ms*; Si*; Ss*;
Ml; Ms; Sl; Si; Mi
*; Mllab*; Ml2a*; Ml3ab*; Milab*;
Mi2ab*; Mi3b*; M14a*; Mslab*; Silab*; Ssla*
*; Mllab*; Ml2a*; Ml3ab*; Milab*;
Mi2ab*; Mi3b*; Mi4a*; Mslab*; Silab*; Ssla*;
Ml4ab; Mslab; Sllab; S12a; Si2a
*; Mllab*; M12a*; Ml3ab*; Milab*;
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.
Alternative
Models
3
5
6
6
6
6
6
6
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
Structure
TI*
1
Guild
Assumptions
No. Guilds
2
A
5
T II *
T III
T IV
B
6
C
8
B
D
9
T
T
T
T
B;
C;
B;
D;
B;
TV
VI
VII
VIII
IX
TX
1
2
9
D
E
E
E
D; E
Basic Guild Structure
(Variables)
10
9
10
10
11
5*;
5*;
5*;
5*;
*;
5*;
*;
5*;
5*;
5*;
Ip*;
Ip*;
ip*;
Ip*;
Ip*;
Ip*;
Ip*;
Ip*;
Ip*;
Ip*;
Sr*; Sb*; Su*
Sr*; Sb*; Su*; Fr*
Sr*; Sb*; Su*; Sr; Sb; Su
Sr*; Sb*; Su*; Sr; Sb; Su; Fr *
Sr*; Sb*; Su*; Sr; Sb; Su; Sb *
Sr*; Sb*; Su*; Sr; Sb; Su; Fr*; Sb *
Ip; Sr*; Sb*; Su*; Sr; Sb; Su
Ip; Sr*; Sb*; Su*; Sr; Sb; Su; Fr *
Ip; Sr*; Sb*; Su*; Sr; Sb; Su; Sb *
Ip; Sr*; Sb*; Su*; Sr; Sb; Su; Fr*; Fb*
Each community structure represents 6 alternative models.
Guild 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 D
is 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)
If F0 = 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
variable
of the community matrix denotes the effect to
from variable
and
(for
from 1 to n) .
For instance,
a community matrix of three variables is represented as:
11
21
31
a12
a13
a22
a23
a32
a33
where diagonal elements a11, a22 and a33 represent the respective
self-effects of variables 1, 2 and 3. Additional elements in
row
represent the effect to variable
Three possible effects to variable
to each matrix element: positive
and no effect (a
from other variables.
from variable
(a1, = 1), negative
are assigned
= -1)
= 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:
interactions,
models,
(1) ecological
(2) number of trophic levels in the case of trophic
(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
27) and all combined invaded community models (r
0.05; n = 77)
.
0.01; n =
0.28; p <
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 including native species
have an asterisk. Derivation of the remaining 36 activity models
is shown in Table D.2.
157
0
AIV
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 = surfacedeposit 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 for the intertidal
benthic community of Yaquina Bay. Ecological interactions
considered are: PP = predation; IC = interference competition
and AN = amensalism. Conn = connectance, F = overall feedback
strength. R-H criteria denotes which Routh-HurwitZ criteria are
met. Models with and without an asterisk represent pre-invaded
and invaded communities, respectively. All models include two
trophic levels (Figure 5.2)
Model
A1*
A2*
A3*
A4
A5
A6
A7
A8
A9*
A10*
All*
Al2*
A13*
A14*
A15
A16
A17
AiB
A19
A20
A21
A22
A23
A24
A25
A26
A27*
A28*
A29*
A30*
A31*
A32*
A33
A34
A35
A36
A37
A38
A39
A40
A41
A42
A43
A44
Community
Structure
Ecological
Interaction
Guilds
Links
Conn
F
0.66
0.83
1.00
0.40
0.60
0.80
0.90
1.00
0.60
0.83
0.73
0.70
0.93
0.83
0.49
0.79
0.70
0.60
0.90
0.82
0.51
0.81
0.70
0.58
0.92
0.82
0.36
0.83
0.69
0.43
0.88
0.74
0.39
0.79
0.67
0.48
0.88
0.75
0.35
0.86
0.72
0.44
0.92
0.80
-2
-1
Al
Al
Al
PP
3
6
PP; AM
PP; IC
3
7
3
8
All
All
All
All
All
Alil
Aill
Alil
Alil
AIII
Alil
PP
5
5
12
16
20
5
22
5
24
6
AIV
AIV
AIV
AIV
AIV
AIV
AV
AV
AV
AV
AV
AV
AvJ:
AVI
AVI
AV]I
AVI
AVE
AVII
AVII
AVII
AVII
AVII
AVII
AVIII
AVIII
AVIII
AVIII
AVIII
AVIII
PP;
PP;
PP;
PP;
PP;
AM
IC
IC
IC
AM; IC
5
IC
10
PP; AM; IC
PP; AM; IC
10
10
IC
IC
P9; AM; IC
PP; AM; IC
10
10
10
11
PP; AM; IC
11
23
30
27
26
33
30
53
80
72
63
90
83
66
99
PP; AN; IC
PP; AM; IC
11
11
87
74
PP; AM; IC
11
111
PP; AN; IC
11
100
PP; AM; IC
11
50
99; AM; IC
P9; AM; IC
PP; AN; IC
11
11
11
101
99; AM; IC
11
11
86
57
108
92
16
16
16
16
16
16
22
22
22
22
22
22
109
205
177
131
227
197
184
406
341
225
448
383
PP; AM; IC
6
PP; AM; IC
6
PP; AM; IC
6
PP; AM;
IC
PP; AM; IC
PP;
PP;
PP;
99;
AM;
AM;
AM;
AM;
IC
PP; AM; IC
99;
AM;
IC
PP; AM; IC
P9; AM; IC
99;
AM;
IC
PP; AM; IC
PP; AM; IC
PP; AN; IC
P9; AM; IC
P9; AM; IC
99;
AM;
IC
PP; AM; IC
6
6
R-H
Criteria
1;2
1;2
0
2
-4
1;2
0
2
4
2
0
2
0
--
0
2
2
2
0
1
6
2
0
2
0
2
2
2
0
--
0
19
0
0
--
0
2
0
--
0
16
0
0
0
0
0
-2
---
14
2
0
0
---
45
2
0
2
0
--
0
0
0
--
0
2
0
0
---
-23
1
0
--
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)
Model
Ti *
T2 *
T3
T4
T5
T6
T7 *
T8 *
T9
T10
Til
T12
T13
T14
T15
T16
T17 *
T18 *
T19
T20
T21
T22
T23 *
T24 *
T25
T26
T27
T28
T29 *
T30 *
T31
T32
T33
T34
T35 *
T36 *
T37
T38
T39
T40
Community
Structure
Ecological
Interaction
T I
PP; SC
T III
2?; SC
22; EC
T II
Cuilds
Links
5
6
17
8
32
37
35
40
18
20
TV
2?; EC
PP; SC
9
9
T VI
PP; EC
10
T II
T III
PP; EC; AM
22; SC; AM
P2; SC; AM
EC;
AN
TV
PP; SC;
AM
5
6
8
9
9
PP;
AM
10
44
22; EC; IC
9
10
46
51
49
54
19
22
40
45
43
48
5
6
18
21
T VI
TI
T IV
T VI
T VII
T VIII
T IX
TX
TI
T II
T III
T IV
TV
T VI
TI
T II
T III
T IV
TV
T VI
TI
T II
T III
T IV
TV
T VI
TI
T II
T III
T IV
TV
T VI
PP;
EC;
22; SC; IC
10
22; EC; IC
10
22; SC; IC
11
PP; EC; MU
5
MU
22; EC; MU
P2; SC; MU
22; EC; MU
6
8
22; SC;
9
9
PP; EC; MU
22; SC; CO
22; EC; CO
P2; EC; CO
P2; EC; CO
P2; EC; CO
PP; EC; CO
PP; SC; AM;
21
36
41
39
8
36
9
9
41
39
10
44
MU
5
PP; EC; AM; MU
6
20
23
22; SC;
AM;
MU
EC;
EC;
AM;
AM;
MU
MU
22;
PP;
PP; SC; AM; MU
22; SC; AM; CO
PP; SC; AM; CO
P2; SC; AM; CO
P2; SC; AM; CO
22; EC; AM; CO
P2; SC; AM; CO
8
9
44
9
47
10
6
8
9
9
52
16
22
40
45
43
10
48
5
49
Connectance Fn
0.70
0.60
0.46
0.47
0.42
0.42
0.75
0.63
0.55
0.53
0.47
0.47
0.56
0.53
0.49
0.47
0.80
0.67
0.59
0.58
0.53
0.51
0.75
0.63
0.54
0.53
0.47
0.47
0.85
0.70
0.68
0.64
0.58
0.56
0.80
0.66
0.61
0.58
0.53
0.51
-1
R-H
crit
i;2
0
2
-1
1;2
0
0
0
2
2
2
-i
1;2
1
2
2
2
2
-1
0
0
0
0
--
2
0
2
0
0
0
2
2
2
2
1
3
0
0
0
-1
0
-1
0
--
-2
2
1;2
--
0
0
0
2
3
0
0
2
2
2
2
--
0
2
-1
1;2
1
2
-1
1;2
0
0
0
2
2
2
162
Table 5.5. Continued.
Model
T41
T42
T43
T44
T45
T46
T47
T48
T49
T50
T51
T52
T53
T54
T55
T56
T57
T58
T59
T60
Community
Structure
T VII
T VIII
T IX
TX
T Vu
T VIII
T IX
TX
T VII
T VIII
T IX
TX
T VII
T VIII
Ecological
Interaction
PP;
PP;
PP;
PP;
PE;
PP;
PP;
PP;
PP;
PP;
PP;
PP;
PP;
EC;
EC;
EC;
EC;
AM; IC
AM; IC
AM; IC
EC;
EC;
EC;
EC;
EC;
EC;
EC;
IC; MU
IC; MU
IC; MU
IC; MU
IC; CO
IC; CO
IC; CO
AN; IC
EC; IC; CO
EC; IC; AN; MU
Guilds
9
10
10
11
9
10
10
11
9
10
10
11
9
PP; SC; IC; AM; MU
PP; SC; IC; AN; MU
10
11
T VII
PP; EC; IC; AM; MU
PP; EC; IC; AM; CO
T IX
P9; EC; IC; AM; CO
T IX
TX
T VIII
TX
PP; SC; IC; AM; CO
99; SC; IC; AM; CO
10
9
10
10
11
Links
50
55
53
58
54
59
57
62
50
55
53
58
58
63
61
66
54
59
57
62
Connectance En
0.61
0.58
0.53
0.51
0.66
0.62
0.58
0.54
0.61
0.58
0.53
0.51
0.72
0.66
0.62
0.57
0.66
0.62
0.58
0.54
0
0
0
0
0
0
0
0
0
0
0
0
0
0
R-H
crit
2
-2
-2
2
--
2
2
2
2
-2
--
0
2
0
2
2
0
0
0
0
-2
--
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
Pre-inv.
13
22
Trophic Models
12
Invaded
Pre-inv.
46
-2to2
OVERALL FEEDBACK
Figure 5.4. Distribution of feedback in models for the preinvaded 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
18
3
24
10012
'I''
C')
-J
w
%11I
0
o
I
80-
I
%
S
I
,
4fl
'.0
I
I
S
I
%
%
S
%
..
S
%
S
'6:
6
-. COMMUNITY MODELS
cDZ 60LU
LU
LLW
o__
LU
--RANDOM MODELS
LU
z
2
4
6
8 101214 16 18 20 22
NUMBER OF VARIABLES (GUILDS)
Figure 5.5. Percent of models with near-zero feedback
Included are all community models of Yaquina
2) .
(-2 F1,
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 sigmoidal decline in near-zero
feedback with increasing number of variables (Figure 5.5) . The
proportion of near-zero feedback models with three variables and
greater differed significantly between community models and
random models (X2 test; P < 0.001)
Discussion
We confirm that as complexity of a benthic estuarine community
increases due to biological invasions, stability as measured
through feedback tends to decrease. Although this conclusion 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 as community
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 preinvaded 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 VI
and 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 still provide 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 predatorfree 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'
39 qualitative
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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180
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 highand 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 temperaturesalinity 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
Pseudopolydora kempi, the bivalve
cumacean Nippoieucon hinuniensis are common
contribution. The polychaete
Mya arenaria and the
NI prey in Alsea Bay and Yaquina Bay. The NI polychaete
Streblospio
benedicti
was a major prey in Yaquina Bay but it was
183
not detected in Alsea Bay. Major native prey for both flatfish in
the two estuaries included the amphipod Corophium salrnonis and
the bivalve Macoma
balthica.
Similar selection for native and NI prey is suggested by
juvenile English sole and starry flounder in Alsea Bay and
Yaquina Bay. Hence, predator-prey coevolution at species-level is
not a critical determinant of prey selection by these estuarinedependent fishes. Juvenile English sole and starry flounder are
generalist predators as indicated by: 1) their reliance on a wide
variety of common native and NI prey; 2) the similar proportions
of prey in gut contents and sediments, suggesting little
selectivity of prey ; and 3) the non-significant difference
between interspecific and intraspecific diet overlap in prey
volume between both flatfish species.
Unlike starry flounder, English sole had lower 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 more visually oriented
predation mode by age-O English sole.
Prey usage by juvenile English sole and starry flounder was
correlated with the availability of individual native species and
NIS. However, the total number of prey in the fish diet was not
correlated with the total prey density among estuarine sites.
Laboratory prey selection experiments in which juvenile
English sole were exposed to equal numbers of native amphipods
(Corophium salmonis and C. spinicorne) and Northwest Atlantic
amphipods (C. acherusicum and C. insidiosum) support field
results in that prey origin was not a critical determinant of
prey selection. Thus, NI arnphipods are potentially capable of
supporting higher trophic levels of native species.
Prey consumption by English sole in single-species experiments
over sand substratum was greater on the 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 consistently higher for NI prey
than for native prey over mud substratum but not over sand
substratum.
No sex-selective predation by English sole occurred on any
species in either substratum type which is further supported by
field observations on English sole and by its generalist feeding
on both epifauna and infauna. 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. In the latter case, the higher exposure
of C. spinicorne and C. acherusicum over sand substratum in
single-species experiments increased predation. The greater
predation on all Corophiurn spp. over sand substratum than over
mud substratum is potentially due to higher prey visibility on
sand. Thus, predation risk increased with prey exposure and
visibility. Other prey characteristics (e.g., origin; size; sex;
activity) were less consequential.
Both activity and trophic guild models of the benthic
community of Yaquina Bay indicate increased risk of stability
decline following species invasions. Such patterns were detected
even after accounting for potential complexity effects on
stability. Nevertheless, community models suggested reduced
response of the benthic community of Yaquina Bay to input from
either invertebrate invasions or potential removal of native fish
predators. Such implied resistance may to some extent buffer the
community from impacts of introduced species.
Although juvenile fishes may not generally act as keystone
predators in the intertidal benthic system of Yaquina Bay, they
may help to control the abundance of certain NIS. Activity and
trophic models revealed a strong tendency of overall feedback
towards zero in both pre-invasion and post-invasion scenarios.
The implied weak or absent overall feedback in the benthic
community of Yaquina Bay may be an ecological property and not an
intrinsic property of large systems as such.
185
Recommendations for Future Research
Suggested areas for continued work on the ecology of estuarine
invasions include field, laboratory and modeling work on shortand 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 (**). No asterisk is indicated for native
species and supraspecific taxa.
Taxa
Annelida
Polychaeta:
Abarenicola pacifica
Armandia brevis
Boccardia proboscidea
Capitella sp.
Capitellidae
(fragment)
Eteone (fragment)
Eteone columbiensis
Eteone spilotus
Fabricia sp.
Glycinde armigera
Glycinde polygnatha
Heteromastus filiformis *
Hobsonia florida *
Leithoscoloplos pugettensis
Manayunkia aestuarina
Mediomastus californiensis
Neomediomastus
Nephtys caeca
Nereis
Al
A2
21
14
35
1167
1492
41
62
Mean Density (No.m2)
A4
A5
A6
al
a3
a6
14
2
3
173
35
35
41
7
656
235
7
28
1
152
76
41
14
28
14
2
111
77
14
2
7
14
14
14
7
332
684
55
111
28
55
14
1913
2666
442
21
62
41
29
629
41
601
1091
14
41
2
22
100
1
11
4
22
56
145
137
677
2
83
1105
401
1
41
14
11
78
11
11
44
7
7
35
69
33
11
11
89
56
11
758
2
14
Overall
occur.
7
857
69
1
214
14
48
423
19
7
269
all sites
8
28
sp.
limnicola
Paraonella platybranchia
Polychaeta (fragment)
Polydora cornuta *
Prionospio sp.
A3
35
24
7
8
2
11
11
78
11
67
22
11
Table A.1. Continued.
Taxa
Pseudopolydora
kempi
Pygosplo elegans **
Scolelepis sp.
Oligochaeta
*
Arthropoda
Crustacea
Amphipoda:
Aliorchestes angusta
Ampithoe lacertosa
Ampithoe valida *
Corophium salmonis
Corophium spinicorne
Eobrolgus spinosus *
Eogammarus confervicolus
Eohaustorius estuarius
Al
A2
207
20841
1699
2528
3820
2908
A3
Mean Density (No.m2)
A4
A5
A6
al
14
14
97
608
28
421
560
2052
1078
1993
242
a6
256
863
7
21
4255
269
4704
14
14
254
159
3150 12013
8462
8186
7
180
83
511
28
28
14
83
1768
69
**
Gnorismosphaeroma insulare
14
55
28
55
14
14
62
41
7
7
193
14
41
28
10
111
14
28
14
2
5
6687
5917
100
28
83
99
67
11
3
83
21
14
14
100
2390
28
28
14
14
6
1683
11
22
22
28
5892
Overall
occur.
67
67
33
6
28
138
all sites
3050
14
41
Brachyura:
Hemigrapsus oregonensis
Copepoda:
Harpacticoida
Hemicyciops subadhaerens
Cumacea:
Cumelia vulgaris
Nippoleucon hinumensis *
Vaunthornpsonia sp.
Isopoda:
a3
7
89
33
6
44
52
7
35
33
89
28
124
26
44
235
152
131
58
57
1
67
78
11
3
22
48
90
14
Table A.l. Continued.
Taxa
Macrura:
Neotrypaea californiensis
Upogebia pugeftensis
Ostracoda
Tanaidacea:
Leptochelia dubia **
Sinelobus stanfordi **
Al
A2
83
A3
Mean Density (No.m2)
A4
A5
A6
al
a3
90
117
28
14
14
a6
7
21
all sites
35
44
2
22
22
4
939
105
7
7
Overall
occur.
1
22
11
Ins ecta:
Chironomidae
Diptera
Insect (fragment)
Mollusca
Bivalvia:
Bivalve (fragment)
Clinocardiuin nuttalli
Cryptomya californica
Macama baithica
83
28
14
31
566
14
760
14
2
14
41
76
21
69
546
7
14
111
318
14
14
166
28
318
221
14
41
111
207
55
62
3
2
35
35
44
100
6
3
11
67
11
11
11
5
6
11
22
2
48
28
22
11
55
4
28
78
22
11
185
28
14
28
164
4
Macoma inquinat:a
Mya arenaria *
Myselia tumida
Mytilus californianus
Mytilus edulis **
Gastropoda:
Alderia modesta
Aplysiopsis enteroinorphae
7
7
Table A.l. Continued.
Taxa
Al
A2
A3
Mean Density (No.m2)
A4
A5
A6
al
Prosobranchia
Phoronida:
Phoronis pailida
Platyhelminthes
Prolecithophora
a6
21
Nematoda
Nemertea
a3
90
14
62
28
152
7
14
14
35
14
76
69
124
14
14
21
14
14
62
35
2217
166
69
28
55
14
124
all sites
Overall
occur.
2
11
17
56
63
100
269
56
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.
Taxa
Y1
Y2
Mean Density (No.rrr2)
Y3
Y4
Y5
Y6
Overall
y2
y4
y6
all sites
occur.
Annelida
Polychaeta:
Abarenicola
pacifica
occidentalis
brevis
Ainaena
Arinandia
Boccardia proboscidea
Branchiomaldane
7
14
14
2715
477
159
988
Lumbrineris sp.
Lumbrineris zonata
Magelona hobsonae
Manayunkia aestuarina
Mediomastus californiensis
283
14
338
338
829
62
394
66
14
4
1
7
173
1
297
14
14
28
83
84
28
35
28
28
21
28
56
89
3
193
179
7
138
3
404
753
21
7
724
426
449
28
Heteromastus
filiformis *
Hobsonia florida *
Leitoscoloplos pugettensis
19
3
sp.
Capitella sp.
Capitellidae (fragment)
Dorvillea annulata
Eteone (fragment)
Eteone californica **
Eteone spilotus
Eupolymnia heterobranchia
Exogone lourei
Glycera americana
Glycinde polygnatha
11
11
22
33
11
78
44
22
11
11
89
11
11
11
44
67
67
44
11
11
11
1
173
3
2
14
14
83
14
28
124
55
14
1161
1534
7
4673
28
2
18
262
48
7723
21
55
1684
12
7
1
55
28
6
1458
1589
242
470
2114
41
14
273
421
138
3523
14
55
Table A.2. Continued.
Mean Density (No.m2)
Taxa
Mesospio sp.
Myriochele sp.
NephLys caeca
Nereis limnicola
Orbinia sp.
Owenia fusiformis **
Paraoneiia platybranchia
Phyllodoce hartmanae
Platynereis bicanaliculata
Polychaeta (fragment)
Polydora cornuia *
Prionospio sp.
Pseudopolydora kempi *
Pseudopolydora paucibranchiata
Pygospio californica
Pygospio elegans **
Rhynchospio sp.l
Rhynchospio sp.2
Scolelepis (fragment)
Scolelepis sp.
Sphaerosyllis californiensis
Spionidae (fragment)
Streblospio benedicti *
Tharyx sp.
Oligochaeta
Yl
Y2
Y3
Y4
Y5
Y6
y2
y4
y6
14
14
14
117
193
539
14
235
14
62
62
41
14
*
83
166
28
14
55
28
131
138
408
14
55
55
28
159
76
449
995
83
62
1637
388
2
373
311
35
470
366
14
2
117
315
7
7
144
14
2
14
7
156
15
2
62
2432
67
11
11
5
3
14
14
14
22
26
14
318
1292
3
151
7
152
1167
11
11
2
90
14
28
3
3
2
2024
207
6625
1444
428
2577
3523
69
290
3053
2318
276
318
3495
1520
14
7
387
1340
Overall
occur.
2
2
14
262
all sites
987
108
2502
33
11
22
100
44
11
78
11
11
78
11
22
11
22
11
11
89
56
100
Table A.2. Continued.
Taxa
Arthropoda
Crustacea
Pmphipoda:
Allorchestes angusta
Ampithoe lacertosa
Ampithoe valida *
Caprellidea
Corophium acherusicurn *
Corophium brevis
Corophium salmonis
Corophium spinicorne
Eobrolgus spinosus *
Eogaminarus confervicolus
Eogammarus sp.
Eohaustorius estuarius
Gammaridea
Traskorchestia traskiana
Yl
Y2
Y3
Mean Density (No.rrr2)
Y4
Y5
YE
Overall
y2
y4
y6
891
124
207
99
15
66
14
14
373
14
166
2590
373
5105
7
304
3917
1402
7751
511
14
283
262
677
41
2
221
28
5281
421
3033
3281
145
14
21
28
1133
97
14
49
288
14
4256
490
129
21
68
55
81
8883
1478
14
649
all sites
2
21
14
2
28
3
occur.
11
22
33
11
44
11
100
100
33
67
11
33
11
11
Bra chyura:
Cancer magister
Hemigrapsus oregonensis
Copepoda:
Calanoidea
Harpacticoida
Hemicyclops subadhaerens **
14
2
21
14
21
21
35
4
14
21
83
48
90
7
14
4
28
28
23
23
41
11
22
22
56
67
Cumacea:
Cumelia vuigaris
Nippoleucon hinuxnensis *
90
35
28
41
14
262
387
808
83
262
48
525
698
14
31
56
335
100
Table A.2. Continued.
Taxa
Yl
Y2
Isopoda:
Gnorisrnosphaeroma insulare
Gnorismosphaeroma oregonensis
Lironeca californica
Macrura:
Neotrypaea californiensis
Upogebia pugettensis
Tanaidacea:
Leptochelia dubia **
Sinelobus stanfordi **
Ostracoda
Mean Density (No.nr2)
Y3
Y4
Y5
Y6
14
14
117
90
Overall
y2
14
7
41
y4
21
7
y6
all sites
14
5
7
3
7
1
28
34
7
13470
62
55
21
occur.
33
33
11
1
67
11
28
1507
44
69
7
33
7
15
15
5
6
11
78
11
11
11
11
11
11
15
61
33
33
268
100
78
27
67
11
7
41
14
14
41
14
21
138
41
67
Insecta:
Aphididae
Chironomidae
Collembola
Diptera
Hesperoconopa Sp.
Hymenoptera
Acarina
Aranea
Mollusca
Bivalvia:
Clinocardium nuttaili
Cryptomya californica
Macoma baithica
Mya arenaria *
Myseila tumida
14
28
21
55
28
3
7
1
14
55
55
55
76
297
76
28
242
256
166
35
5
6
6
14
41
28
394
138
221
193
21
718
111
276
41
401
235
55
Table A.2. Continued.
Mean Density (No.m2)
Taxa
Mytilus edulis **
Transenella tantilla
Bivalve (fragment)
Yl
Y2
221
256
Gas tropoda:
Aicieria modesta
Y3
41
14
14
28
41
Overall
y2
y4
y6
6
97
18
14
8
28
21
21
21
Nemertea
159
117
111
48
28
857
55
7
55
83
33
11
22
33
33
11
11
7
22
69
7
272
89
338
62
93
67
14
140
44
180
38
67
2
22
359
7
14
414
occur.
2
3
41
7
all sites
36
28
14
1851
Platyhelminthes
Prolecithophora
Turbellaria
Y6
76
Nematoda
Phoronida
Phoronis pailida
Y5
28
41
Aplysiopsis enteromorphae
Gastropoda (fragment)
Littorina sitkana
Melanochiamys diomedea
Y4
14
7
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.1000 m2)
Species
Al
A2
259
20
57
Engraulis inordax
Cymaogaster aggregata
Leptocottus armatus
Gasterosteus aculeatus
Oncorhynchus tshawytscha
Crangon franciscorum (D)
Pleuronectes vetulus
Platichthys stellatus
Cancer magister (D)
Pholis ornata
Hypomesus pretiosus
Hemigrapsus oregonensis (D)
Cottus asper
Clevelandia
Oligocottus maculosus
Atherinops affinis
Oncorhynchus kisutch
Ciupea pailasii
Heptacarpus paludicola (D)
A3
A4
1
1
931
6323
36108
1470
27
137
3
32
85
20
6
1
24
22
2
4
A5
A6
38
1
2
2663
730
636
22
5196
64
1176
58
34
403
69
35
43
57
112
36
44
221
201
130
7853
1331
378
182
106
16
9
al
a3
a6
9
11
43
37
2
8
16
1
3
5
5
3
2
1
2
14
1
1
3
2
5
1
1
2
2
1
1
4
2
All Sites
4016.7
2817.1
400.1
102.4
65.6
55.6
26.8
9.5
4.1
3.6
2.5
2.2
0.6
0.5
0.3
0.2
0.2
0.1
0.1
Occurr.
67
100
100
67
100
44
56
67
11
67
56
67
22
11
22
11
11
11
11
Table A.4. Summer mean catch per unit effort (CPUE) and occurrence of fishes and decapods in Yaquina
Bay as determined from seine sampling. Six intertidal sites (Yl to Y6) and three subtidal sites (y2,
y4 and y6) are considered. Species are ordered in decreasing mean CPUE for all sites combined. Species
code: nonindigenous species (*), decapod crustaceans (D).
Mean CPUE (No.1000 m)
Species
Engraulis mordax
Cyrnatogaster aggregata
Leptocottus armatus
Crangon franciscorum (D)
Cancer magister (0)
Gasterosteus aculeatus
Clupea pailasii
Atherinops affinis
Piatichthys steilatus
Yl
Y2
Y3
3
669
637
90
15
617
507
647
59
1599
100
2
9
1
4
Oncorhynchus tshawytscha
Pleuronectes vetulus
Pholis ornata
Lepidogobius lepidus
Hemigrapsus oregonensis (D)
Alosa .sapidissiina *
35
11
Y6
24415
1690
291
1381
459
322
28
69
37
1
4
5
2
48
y2
y4
y6
778
205
320
63
404
80
2302
2
1
104
6
39
27
63
6
2
6
7
2
2
62
67
2
8
6
5
1
1
21
43
44
2
8
9
3
1
20
20
13
25
4
11
3
1
3
1
2
7
1
2
2
1
4
2
2
2
2
2
2
6
10
8
2
81
20
4
10
4
1
40
2
3
Pholis schuitzi
Aulorhynchus fiavidus
Heptacarpus paludicola (0)
Puggetia producta (0)
Lucania parva *
Luinpenus sagitta
1078
2149
3740
Y5
22
49
24
Hyperprosopon argenteuni
Syngnathus leptorhynchus
Phanerodon furcatus
Hypomesus pretiosus
Cancer productus (D)
3
Y4
2
All Sites
3008.3
1272.0
636.0
54.4
20.8
18.0
16.9
15.2
14.7
12.3
12.2
3.6
1.7
1.7
1.6
1.5
1.1
0.7
0.6
0.5
0.3
0.2
0.2
0.2
0.2
0.2
Occurr.
78
100
100
89
78
78
33
78
89
89
67
44
22
56
44
22
22
22
33
11
22
11
11
11
11
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 functionalgroups (Su = suspension feeder, Sr = surface-deposit feeder, H =
herbivore, D = detritivore)
Functional
Life Mode
Activity
U'
Si'
T3
T7
F9
F7
Si
Heteromastus filiformis
Taxa
Group
Trophic
Bivalvia:
Mya arenaria
Sr
Su
Crustacea:
Ampithoe valida
H 5;
''
6
D
Si
Su 7;
Mi
Sr?
Ms "
Sr? 9,10
T7
Ml"
Sb"
Hobsonia florida
T
Si 7,12
Sr "
Polydora cornuta
T
Si
Su 14; Sr
14
Pseudopolydora kempi
T
Sr 11;
Su
15
Pseudopolydora paucibranchiata
T
Si
Su 17; Sr
17
Streblospio benedicti
T 7,
Si
Sr 19; Su
19
Corophium acherusicum
Eobrolgus spinosus
Nippoleucon hinumensis
Sr
2
Poiychaeta:
'
,
15
16
18
Si
12
Sources: 1 Rudy and Hay (1991); 2 J.W. Chapman (personal
Duffy and Hay (1994);
Borowsky (1983);
communication 1998);
G.C. Castillo (personal
Duffy (1990); 6 Nelson (1979);
Kozioff (1990); '° L. Watling
observation); 8 Bousfield (1973);
Neira and Hopner (1993);
(personal communication 1997);
12
Hobson and Banse (1981); 13 Hentschel and Jumars (1994);
14
Dauer et al. (1981); " Taghon and Greene (1992);
16
Crooks and Khim (1999); 17 Shimeta (1996); 18 Fauchald and
Kalke and Montagna (1991)
Jumars (1979);
'
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 5 C
= 0.845, SE = 0.007; Yaquina
flounder (Alsea
C = 0.861, SE = 0.008); starry
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
English sole
A
10- W = 0.00877L2983 (R2 = 0.96, n = 246)
8
:j
6
'I
I
6Jsea Yaquina
Mean
6.8
7.1
L (cm)
-
W(g)
20
0
5
150
10
15
20
25
Starry flounder
B
'-
3.4
3.0
W = 0.01041L3°°8 (R2 = 0.99, n = 122)
Mean Aisea Yaquina
9.0
13.3
L (cm)
100
W(g)
14.9
34.4
50
10
15
20
TOTAL LENGTH (mm)
Figure B.1
25
Table Bi.
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. 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 intertidal
areas and at
low-tide in subtidal areas, respectively. SE = standard error.
Estuary/Site
Alsea Bay:
A1H
A2H
A3H
AlL
A2L
A3L
Yaquina Bay:
Y1H
Y2H
Y3H
Y4H
Y2L
Y3L
Ratio Fish
With Prey
Stomach
Fullness
Total Fish
Length (cm)
Total Fish
Weight (g)
Mean
SE
Mean
SE
Mean
SE
6/6
1/1
22/22
78/88
8/9
8/9
3.0
4.0
3.6
1.4
1.9
1.2
0.7
0.5
0.7
0.1
0.2
0.5
6.3
6.9
5.1
7.6
5.7
4.7
2.5
2.8
1.3
3.8
1.6
1.0
19/26
1.6
4.0
3.8
3.0
1.3
1.2
0.3
0.0
0.2
0.3
0.3
0.2
7.7
6.6
7.6
7.8
5.5
6.6
0.2
0.7
0.2
0.2
0.3
0.2
4.4
2.8
3.9
4.7
1.6
2/2
19/19
12/13
18/18
24/35
0.1
0.1
0.1
0.2
2.5
Mean Prey Richness
(No. taxa per fish)
NA
NI
CR
ALL
0.6
1.8
0.1
0.2
0.1
0.1
2.0
2.0
3.1
2.0
1.8
1.2
3.0
0.5
0.5
1.0
0.3
1.3
1.0
0.5
0.1
0.2
0.0
6.0
9.0
5.3
7.8
6.2
5.2
0.3
0.7
0.3
0.4
0.2
0.2
2.2
2.0
3.7
2.2
1.2
0.6
0.2
2.5
2.8
1.8
0.8
0.8
0.7
0.0
0.5
0.0
0.1
0.0
6.2
6.0
10.4
4.7
4.8
3.1
Table 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. 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
intertidal
areas
and at low-tide in adjacent subtidal areas, respectively. SE = standard error.
Estuary/Site
Ratio Fish
With Prey
Stomach
Fullness
Mean
SE
Mean
SE
NA
NI
CR
ALL
1.1
0.8
0.7
0.6
0.8
11.2
23.0
3.3
29.7
12.0
65.8
11.4
1.3
0.2
6.4
2.6
0.8
1.5
1.1
2.4
-0.1
2.0
2.4
2.7
2.1
1.4
4.0
2.3
1.7
0.8
0.7
0.5
0.8
0.0
0.8
0.0
0.0
0.0
0.0
0.0
0.0
0.0
5.7
3.8
4.0
5.3
4.4
6.0
6.4
0.7
15.7
11.5
13.1
10.7
14.7
21.5
19.2
11.2
3.2
5.2
1.0
1.1
1.2
57.3
25.3
30.8
15.6
41.2
109.5
76.2
19.6
2.6
23.1
4.2
6.0
4.7
3.7
4.2
3.0
4.3
2.3
3.0
4.2
0.7
0.5
2.9
1.8
0.9
0.0
2.0
1.3
1.0
1.0
0.5
0.5
0.0
0.0
0.0
0.0
0.3
0.0
0.0
0.0
8.5
10.0
11.1
9.5
7.6
27/27
Yaquina Bay:
Y1H
Y2H
Y4H
Y5H
Y6H
Y1L
Y3L
Y4L
Y5L
Y6L
3/4
2/2
17/17
6/6
13/14
1/1
3/3
4/4
1/1
9/9
2.2
2.0
2.1
2.7
1.1
4.0
1.0
0.7
4.0
2.2
5/5
1/1
Mean Prey Richness
(No. taxa per fish)
SE
2.2
3.2
0.7
2.3
1.2
4.0
3.2
11/13
Total Fish
Weight (g)
Mean
Alsea Bay:
A2H
A4H
A5H
A6H
AlL
A4L
A6L
4/4
5/5
6/6
Total Fish
Length (cm)
2.0
0.4
0.7
0.4
0.6
0.7
--
0.6
6.7
12.9
9.0
16.6
5.6
6.5
11.1
2.6
2.0
--
1.9
1.8
3.1
9.1
3.5
22.9
-0.2
22.8
22.9
5.5
4.8
7.1
--
29.0
9.9
--
8.0
6.0
10.0
5.8
6.0
8.0
228
Figure B.2 Fulton's condition factor of English sole and starry
flounder. Fish were collected at intertidal and subtidal areas of
Alsea Bay and Yaquina Bay estuaries in summer 1993.
229
English Sole
113
135
1.11.0Median
75 % Percentile
0.9c
0
I-
/
Mean
25 % Percentile
LJ
z
0
0.6
Yaquina
Alsea
I-
z
0
Starry Flounder
C.)
1.4-
Cl)
1.3-
z
0
IJ
U
61
61
-
1.11.0-
0.90.80.7
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:
[(P
DO1,2 = E
where P1
and P2
\1J
2j I
[B1]
are the proportions of volume for prey
predator species 1 and 2
in
(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
where P2
=
[
(P2)]
is 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
(Rank)
English sole
Alsea Bay
B= 8.1
English sole
Yaquina Bay
Starry flounder
Alsea Bay
Starry flounder
Yaquina Bay
B= 7.4
B1= 4.7
B1=9.1
English sole
Alsea Bay
English sole
Yaquina Bay
Starry flounder
Alsea Bay
16
13
(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
predator species
(IPV1), which is defined as:
IPV
where V1
and
in
=
[
V] [
N] -1
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 species by volume
of individual prey in the diet of juvenile English sole and
starry flounder in the Alsea and Yaquina estuaries.
C,,
3
3
C
C-)
m
Cl)
m
C
0I-
-<
'ii
C
a
z
D
0
0
Pseudo potydora kenai
Heteromastus flhifomas
Mya arenaria
Eobrolgus spinosus
Hobsonia florida
Streblospio benedi cii
Corophium acherusicum
Mya arenoria (s)
Nippoleucon hinurnensis
61
Mya arenaria
Mya arenaria (a)
Hobsonia florida
Pseudo polydora kempi
Nippoleucon hinurrensis
Eobrolgus spinosus
a'
No. NI SPECIES
0
Engraulls mordas
Cymatogaster aggregata
Clinocardium nutlofiuii
Macnina bait/rica
Upogebia pugettensis
Nereis limnicola
Neotrypaea colforniensis
Mans yunkia aestuarina
Curnella ailgaris
Mysella turnido
Cymatogasfer aggre gata
Cryptomya californica
Crangon franciscorum
Nereis Iimr,icola
Macoma balthica
Currella silgaris
Macurns baithica (a)
Corophium salmonis
Corophium spinicorne
Eteone spilotus
Cii
0
0
-a
No. NATIVE SPECIES
0
C-)
0a
0
0
0
0
0
0
0
c,
Mya arenaria (a)
Hobsonia florida
SfrebioSpio benedicti /
Corophium acherusicum
Eobroigus spinosus
Myo arenaria
Pseudopoiydora kempi
Anpithoe 5Iida \
Nippoleucon hmnumensis
0
Eobrolgus spinosus
Hobsonia florida
Pseudopolydora kempi
Mya arenaria
Nippoleucon hmnurnensis
-S
No. NI SPECIES
0-s
armigera
/Armondia braids
Cumella vulgans
Macnina baithica (s)
Mona yunkia aestuarina
Phyliodoce hart rranae
Transenelta tan/rita
Glycinde polygnatha
IGiycinde
C,'
Nereis hmnico)a
Glycinde polygnatha
Nephthys coeca
Arrnandia bresis
Curnelia uvigoris
Macomo bait/rica (a)
0
-s
No. NATIVE SPECIES
rw
01>
Table 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.
N.onindigenous and cryptogenic species are denoted by one and two asterisks, respectively. Based on
135 fish from Alsea Bay and 113 fish from Yaquina Bay.
Alsea Bay
Taxa
Prey
Occurr.
Mean
No. Prey
Yaquina Bay
Mean Prey
Vol.
(mm )
Mean
No. Prey
Mean Prey
23.0
0.9
0.186
0.265
2.381
0.062
0.168
0.009
1.373
0.186
3.765
0.085
0.247
0.022
0.9
3.5
16.8
0.9
0.009
0.035
0.301
0.009
0.088
0.007
1.259
0.001
0.9
0.9
2.7
4.4
2.7
0.009
0.009
0.044
0.071
0.035
0.619
0.133
0.032
0.011
0.178
0.9
0.9
26.5
25.7
0.009
0.009
0.867
1.619
0.044
0.004
0.743
5.797
Prey
Occurr.
Vol.
(mm3)
Anne lida
Polychaeta:
Amaena occidentalis
Armandia brevis
Boccardia proboscidea
Capitella sp.
Capitellid part
Dorvillea annulata
Dorvilleidae
Eteone californica
Eteone sp.
Eteone spilotus
Exogone sp.
Glyceridae part
Glycinde armigera
Glycinde polygnatha
Hobsonia florida *
Manayunkia aestuarina
Mediomastus californiensis
Nephthys caeca
Nerejs limnicola
Owenia fusiformis
Phyllodoce Hartman
Polychaeta part
Pseudopolydora kempi *
Pygospio californica
39.3
1.807
1.727
11.1
14.8
0.163
0.356
0.233
0.327
0.7
0.007
0.004
4.4
0.7
0.7
0.059
0.007
0.007
0.533
0.007
0.007
4.4
9.6
0.044
0.200
0.611
0.367
1.5
0.7
1.5
1.5
0.7
0.022
0.007
0.015
0.015
0.007
0.570
0.141
0.015
0.022
0.296
0.504
0.015
0.007
0.345
0.774
0.015
29.6
6.7
1.5
4.4
4.4
5.3
4.4
Table B.3. Continued.
r axa
Pygospio elegans **
Rhynchospio
sp.l
Spionidae (juvenile)
Aisea Iay
% Prey
Occurr.
Mean
No. Prey
17.8
2.556
1.5
Streblospio benedicti *
0.022
Yaquina Bay
Mean Prey
Vol.
(mm
1.430
19.3
0.274
0.221
Arthropoda
Acarina
Crustacea
Brachyura:
Hemigrapsus oregonensis
Caridea:
Crangon franciscorum
Amphipoda:
Allorchestes angusta
Amphipoda part
Ampithoe sp.
Ampithoe valida *
Corophium acherusicum *
Corophium brevis
Corophium salmonis
Corophium spinicorne
Eobroigus spinosus *
Eogammarus confervicolus
Copepoda:
Cyclopoid copepod
Hemicyciops subadhaerens
Harpacticoida
5.2
0.7
0.7
0.044
0.007
0.007
Prey
Occurr.
10.6
Mean
No. Prey
Mean Prey
Vol.
(mm3)
1.8
0.487
0.018
0.396
0.018
39.8
0.9
9.7
7.593
0.009
0.292
17.878
0.004
0.235
0.9
0.009
0.001
0.9
0.009
0.022
1.8
0.018
0.018
0.9
2.7
0.009
0.018
0.022
0.031
0.004
Tharyx sp.
Oligochaeta
)
0.052
0.001
0.007
7.1
45.2
0.7
2.2
9.6
0.832
1.963
0.007
0.030
0.119
2.201
3.737
0.007
0.063
0.156
37.2
9.7
1.8
4.4
4.619
0.434
0.044
0.071
10.839
2.177
0.053
0.157
1.5
87.7
0.022
26.615
0.015
1.573
66.4
24.283
1.548
Table B.3. Continued.
1 axa
Alsea Bay
% Prey
Occurr.
Crustacean part
Nippoleucon
his umensis *
Macrura:
Neotrypaea
californiensis
Upogebia pugettensis
Tanaidacea:
Leptochelia dubia **
Pancolus californiensis **
Synelobus stanfordi **
Ostracoda
Insecta
Diptera:
Chironomidae
Insect parts
Mollusca
Bivalvia:
Bivalve siphon
Bivalve part
Ciinocardium nuttaili
Cryptomya californica
Macoma baithica
Macoma baithica (siphon)
Mean Prey
Vol.
(mm )
Prey
Occurr.
Mean
No. Prey
Mean Prey
Vol. (mm3)
3.0
0.037
0.023
3.5
0.062
0.025
1.8
47.4
16.3
1.281
0.363
0.804
0.189
22.1
38.1
0.018
0.327
7.407
0.018
0.205
4.777
4.4
2.2
0.044
0.030
0.230
0.052
2.7
0.035
0.155
0.7
2.2
3.0
0.030
0.022
0.030
0.004
0.027
0.030
12.4
0.9
3.0
3.062
0.009
0.030
3.040
0.004
0.030
20.7
0.778
0.127
8.8
0.142
0.033
0.7
0.7
0.007
0.007
0.007
0.007
4.4
0.053
0.059
0.7
10.4
2.2
3.0
77.8
0. 007
0.007
0.233
0.9
0.9
1.8
0.9
43.4
0.009
0.009
0.004
0.004
0.031
0.018
2.452
Cumacea:
Cumacean part
Cumella vulgaris
Mean
No. Prey
Yaquina Bay
0.126
0.022
0.044
29.548
0. 033
0.048
6.670
0. 027
0.018
5. 602
Table B.3. Continued.
Taxa
Alsea Bay
Prey
Occurr.
Bivalvia:
Mya arenaria *
Mya arenaria (siphon)
Mysella tumida
Transenella tantilla
Gastropoda
Phoronida
Phoronjda part
Miscellaneous items:
Feather
Invertebrate parts
Organic matter
Plant matter
Plastic line
Seed
Stone
Wood fragment
*
Mean
No. Prey
Yaquina Bay
Mean Prey
Vol.
(mm )
17.8
1.5
3.0
0.319
0.022
0.037
1.296
0.011
0.041
0.7
0.007
0.019
37
0.044
0.056
9.6
0.133
0.063
47.4
13.3
2.2
1.5
32.6
56.3
0.785
0.185
0.030
0.015
0.763
1.615
0.630
0.092
0.011
0.052
0.467
0.945
% Prey
Occurr.
Mean
No. Prey
Mean Prey
Vol.
4.4
0.133
0.642
0.9
0.009
0.004
2.7
20.4
3.5
0.053
0.416
0.646
0.062
0.451
0.360
0.489
5.3
34.5
0.124
0.558
0.064
0.336
27
. 4
0. 012
(mm3)
Table 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.
Nonindigenous and cryptogenic species are denoted by one and two asterisks,
respectively. Based on
61 fish per estuary.
Taxa
Alsea Bay
Prey
Occurr.
Mean
No. Prey
Yaquina Bay
Mean Prey
Vol.
(rnm)
Prey
Occurr.
Mean
No. Prey
Mean Prey
Vol.
(mm3)
Anne 1 ida
Polychaeta:
Capil:ella sp.
Capitellid part
Eteone spilotus
Heteromascus filiformis *
Hobsonia florida *
Manayunkia aestuarina
3.3
0.033
0.007
57 . 4
1.082
5.149
Mediomastus
call forniensis
Nereis lirnnicola
Paraonella platybranchia
Polychaeta part
Pseudopolydora kempi *
Pygospio elegans
Streblospio benedicti *
Tharyx sp.
01 igochaeta
Arthropoda
Crustacea
Caridea:
Crangon franciscorum
4.9
4.9
4.9
1.6
49.2
16.4
3.3
11.5
1.6
21.3
0.098
0.148
0.082
0.016
7.852
1. 623
0.016
1.6
24.6
1.6
6.6
0.033
0.180
0.016
0.426
6.623
0.279
11.738
0.049
0.230
0.328
4.9
0.098
67.2
1.492
78.010
4.9
3.3
1.6
0.066
0.148
0.016
0.066
1.148
0.016
1.6
0.016
1.6
0.016
36. 1
0.066
0.148
0.148
0. 492
13. 684
0.262
0.180
5.820
0.016
0.379
71.811
0. 098
23. 992
0.033
0.869
Table B.4. Continued.
Tax a
Prey
Occurr.
Prey
Amphipoda:
Ampithoe lacertosa
Corophium acherusicum *
Corophium salmonis
Corophi urn spinicorne
Eobrolgus
Alsea Bay
spinosus *
Eogarnrnarus confervicolus
Mean
No. Prey
Yaquina Bay
Mean Prey
Vol.
(mm )
Prey
Occurr.
Mean
No. Prey
Mean
Vol.
(mm3)
1.6
3.3
70.5
31.1
0.016
0.164
18.852
5.623
0.016
0.107
58.656
28.279
86.9
31.1
13.410
0.770
38.559
1.754
1.6
0.033
0.016
3.3
1.6
0.049
0.033
0.049
0.246
3.3
0.066
0.007
52.5
27.066
2.408
1.6
1.6
22.9
0.049
0.180
1.279
0.002
0.016
0.164
3.3
0.049
0.180
4.9
18.0
0.115
1.443
0.066
1.311
9.8
1.6
0.311
0.033
12.623
0.656
3.3
1.672
0.639
Copepoda:
Cyclopoid copepod
Hernicyciops subadhaerens
Harpacticoida
Crustacean part
Cumacea:
Curne]ia vulgaris
Nippoleucon hinurnensis *
Macrura:
1.6
6.6
0.016
0.115
0.082
0.082
Neotrypaea californiensis
Upogebia pugettensis
Tanaidacea:
Leptochelia dubia **
Ostracoda
Insecta
Diptera:
Chironomidae
Diptera pupae
Insect parts
Orthoptera
22.9
0.410
0.075
9.8
0.279
0.069
47.5
42.6
2.230
1. 154
0. 670
11.5
3.3
1.6
1.6
0.393
0.033
0.016
0.016
0.557
0.197
0.016
0.016
1. 082
Table B.4. Continued.
Taxa
Alsea Bay
o Prey
Occurr.
Mollusca
Bivalvja:
Bivalve siphon
Bivalve part
Clinocardjum nuttallj
Cryptomya californica
Mean Prey
Vol.
(mm
3.3
1.6
0.066
0.016
0.049
0.008
3.3
0.131
2.049
6.6
18.0
3.3
4.9
0.197
0.508
0.066
0.721
13.934
0.852
10.492
75.738
Crtomya californica (siphon)
Macoma balthjca
Macoma baithica (siphon)
Mya arenaria *
Mya arenaria (siphon) *
Mysella tumida
Mean
No. Prey
Yaquina Bay
Phoronida
Phoronida part
)
Prey
Occurr.
1.6
1.6
3.3
1.6
3.3
13.1
34.4
Mean
No. Prey
Mean Prey
Vol.
(mm3)
1.6
1.6
0.016
0.016
0.023
0.016
0.066
0.311
6.820
1.311
0.131
0.033
0.008
0.016
4.180
0.082
0.328
42.311
8.805
56.231
0.098
0.025
1.6
0.033
0.016
1.6
1.6
0.016
0.016
6.557
16.393
3.3
13.1
18.0
1.6
14.7
50.8
0.033
0.131
0.672
0.016
0.508
2.148
0.018
0.189
18.541
0.033
0.946
7.649
22.9
Fish:
Cymatogaster aggregata
Engraulis mordax
Miscellaneous items:
Feather
Organic matter
Plant matter
Plastic line
Stone
Wood fragment
1.6
0.016
24.590
8.2
18.0
0.213
0.361
2.680
1.346
9.8
0.164
0.721
0.131
0.987
32.8
243
Appendix C
Complement of Chapter 4
244
Males
Females
C. salmonis
1500
C. salmonis
Y-52.74+572.3X
1500
Y = 21.29 e6129x
2_ 0.64, 38 d.f.)
(R2= 0.71, 39d.f.)
1000
1000
500
500
0
00
0.5
0
1.0
1.5
00
2.0
C. spinicorne
0.5
1.0
2.0
1.5
C. spin/come
1500 Y =- 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.)
00
0.5
0
1.0
1.5
2.0
00
C. acherusicum
1500
(R2 0.50, 36 d.f.)
1000
1000
500
500
00
0.5
1.0
1.5
2.0
00
C. insidiosum
1500-
Y
1.5
2.0
Y = 34.40 + 368X
(R2= 0.16, 39 d.f.)
0.5
1.0
1.5
2.0
C. insidiosum
1500
51.01 e18
(R2 0.37, 40 d.f.)
V = -15.99 + 485.5X
(R2= 0.18, 39d.f.)
1000
500
1.0
C. acheruskum
Y 54.90e4
1500
0.5
1000
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 and sex. All correlations are
significant (P < 0.05).
245
30-
I Treatment
C,)
a
0
a
zdl
z
Cl)
w
SAL
SPI
ACH
INS
SPECIES
Figure 0.2. Mean number of surviving Corophium in singlespecies 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
I
a
SPI
ACH
a
-I-
INS
MUD
B
Control
SAL
SPI
Treatment
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 standard error
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
Site
No.
Prey Size Range (mm)
Fish
M
F
Density
Benthos
M+F
Adjusted Prey Length
No. Prey
Fish Diet
M/F
P
M+F M/F
Density
Benthos
M+F
(Norrr2)
M/F
No. Prey
Fish Diet
M+F
M/F
X2
P
(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
No.
Bivalvia
Macoma balthica siphon
Mya arenaria*
Mya arenaria* (siphon)
Percent
Occurrence
0.56
0.03
0.03
19.4
2.8
2.8
0.11
16.7
0.03
2.8
0.17
0.25
7.78
2.31
0.03
1.55
16.7
22.2
75.0
44.4
2.8
36.1
0.14
11.1
0.05
0.03
0.03
8.3
2.8
2.8
0.74
16.7
0.03
0.08
0.03
2.8
5.6
2.8
0.64
16.7
0.28
0.06
16.7
2.8
Fish
Fish, fragment
Leptocottus armatus
Engraulis mordax
Cymatogaster aggregata
0.19
0.03
0.33
0.06
11.1
2.8
5.6
5.6
Miscellaneous
Organic matter
Plant matter
Wood fragment
Stone
0.47
0.14
1.22
0.75
30.6
13.9
50.0
47.2
Polychaeta
Nerei.s limnicola
Arthropoda
Acarina
Crustacea
2mph ipoda
Ampithoe lacertosa
Ampithoe sp.
Corophium salmonis
Corophium spinicorne
Eobrolgus spinosus *
Eogammarus confervicolus
Ca ride a
Crangon franciscorum
Copepoda
Copepoda, unidentified
1-larpacticoid
Hemicyclops sp.
Cumacea
Nippoleucon hinumensis *
Decapoda
Zoea
Neotrypaea californiensis
Hernigrapsus oregonensis
Tanaidacea
Pancolus californiensis **
Isopoda
Gnorissmosphaeroma insulare
Gnorissmosphaeroma sp.
250
Table D.2. Activity models derived from models in Figure 5.2.
Negative effect to guild i from guild j is denoted as (i,j)=-l.
Reciprocal negative effect between guilds i and j is 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:
Relative sizes
S
i
1
S
-1
1
0
-1
-1
1
0
0
-1
-1
-1
Interactions Y2 and X2:
Relative sizes
S
i
1
i
-1
-1
1
0
-1
-1
-1
-1
-1
-1
S
Ba sic
Model
A2
A3
A5
A6
A7
A8
AlO
All
Al2
A13
A14
A16
A17
A18
A19
A20
A22
A23
A24
A25
A26
A28
A29
A30
A31
A32
A34
A35
A36
A37
A38
A40
A41
A42
A43
A44
Model
Al
Al
A4
A4
A4
A4
A9
A9
A9
A9
A9
A15
A15
A15
A15
A15
A21
A21
A21
A21
A21
A27
A27
A27
A27
A27
A33
A33
A33
A33
A33
A39
A39
A39
A39
A39
Additional Interactions
(M,S)=-1
(M/S)=-i
(M*,S*)=_i; (M*,S)=_l; (M,S)=-l; (M,S*)=_i
(M*/S*)=i; (M*/S)=_l; (M/S)=-i; (M/S*)=_i
(M*/S*)=_1; (M*/S)=_i; (M/S)=-l; (M/S*)_l; (M*/M)=_l
(M*/S*)=_1;
(S*/S)=_i
X2
Xl
Y2
Y2; X2
Y2; Xi
X2
Xl
Y2
Y2; X2
Y2; Xl
X2
Xl
Y2
Y2; X2
Y2; Xl
X2
Xl
Y2
Y2; X2
Y2; Xl
X2
Xl
Y2
Y2; X2
Y2; Xl
X2
Xl
Y2
Y2; X2
Y2; Xl
(M*/S)=_i; (M/S)=-l; (M/S*)=_1; (M*/M)=_i;
251
Table D.3. Trophic models derived from models in Figure 5.3.
Negative effect to guild i from guild j is denoted as (i,j)
1. Positive effect to guild ± from guild j is denoted as(i,j)=
1. Reciprocal positive effect between guilds i and j is denoted
as (i/j)= 1.
Ba s ± c
Model
T7
T8
T9
TiO
Til
T12
T17
T18
T19
T20
T21
T22
T23
T24
T25
T26
T27
T28
T29
T30
T3i
T32
T33
T34
T35
T36
T37
T38
T39
T40
'141
T42
T43
T44
T45
T46
T47
T48
T49
T50
T5i
Additional Interactions
Model
Ti
(Su*,Sr*)= -1
T2
T3
(Su*,Sr*)= -1; (Su*,Sr)= -1; (Su,Sr*)= -1; (Su,Sr)= -i
T4
T5
T6
asinT7
as in T9
as in T9
as in
T9
Ti
(Sr*/Sb*)= 1
T2
T3
(Sr*/Sb*)= 1; (Sr*/Sb)= 1; (Sr/Sb*)= 1; (Sr/Sb)= 1
T4
T5
T6
Ti
T2
T3
T4
T5
T6
Ti
T2
T3
T4
T5
T6
Ti
T2
T3
T4
T5
T6
Ti3
T14
T15
T16
T13
T14
Ti5
T16
T13
T14
Ti5
T52
T53
T54
T55
T56
T16
T57
T13
T58
T59
T60
'114
Ti3
T14
T15
Ti6
Ti5
T16
as in T17
as in T19
as in T19
as in T19
(Sb*, Sr*)= 1
as in T23
(Sb*, Sr*)= 1; (Sb*, Sr)= 1; (Sb, Sr*)= 1; (Sb,Sr)= 1
as in T25
as in T25
as in
T25
(Sr*/Sb*)= 1; (Su*,Sr*)=_i
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
as in
T29
T9
T9
T9
T9
T7
T7
T9
T9
T9
T9
T9
T9
T9
T9
and
and
and
and
and
and
and
and
and
and
Ti9
Ti9
T19
Ti9
T23
T23
T25
T25
T25
T25
Ti9
T19
Ti9
Ti9
T25
T25
T25
T25
T9
T9
T9
T9
T9
T9
T9
T9
and
and
and
and
and
and
and
and
Ti9
T19
T19
Ti9
T25
T25
T25
T25