Feeding Ecology and Bioturbation: Determining the Ecological Role of Euspira Lewisii
FEEDING ECOLOGY AND BIOTURBATION: DETERMINING THE ECOLOGICAL ROLE OF EUSPIRA LEWISII
by
Nicola Ashley Cook B.Sc., University of British Columbia, 2001
THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF
MASTER OF SCIENCE
In the Department of Biological Sciences
© Nicola Ashley Cook 2008
SIMON FRASER UNIVERSITY
Spring 2008
All rights reserved. This work may not be reproduced in whole or in part, by photocopy or other means, without permission of the author. APPROVAL
Name: Nicola Ashley Cook
Degree: Master of Science
Title of Thesis:
Feeding ecology and bioturbation: Determining the ecological role of Euspira lewisii
Examining Committee:
Chair: Dr. F. Law, Professor
Dr. L. Bendell-Young, Professor, Senior Supervisor Department of Biological Sciences, S.F.V.
Dr. M. Hart, Associate Professor Department of Biological Sciences, S.F.V.
Dr. I. Cote, Professor Department of Biological Sciences, S.F.V. Public Examiner
27 February 2008 Date Approved
11 SIMON I:RASER UNIVERSITY LIBRARY
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Revised: Fall 2007 ABSTRACT
The burrowing, predatory snail Euspira lewisii is being removed from intertidal habitats due to its reputation as an economically damaging species to shellfish aquaculture. Here, the objectives were to examine feeding ecology and determine the functional role of a poorly understood species. Feeding experiments and shell assemblages showed distinct prey preferences, avoidance of the commercially valuable Manila clam, a low, species-dependent feeding rate and a limited yearly consumption of the clam population. Exclusion experiments demonstrated increased sediment compaction, silt content and nutrient accumulations and decreased water content when E. lewisii is absent.
Homogenized biological communities in cages resembled less diverse areas.
These results demonstrate that E. lewisii is a low impact predator and acts as an ecosystem engineer to the benefit of other organisms. These results can be used to advise shellfish growers that control measures are not necessary and they will benefit from the maintenance of a healthy ecosystem.
Keywords:
Euspira lewisii; intertidal; community structure and function; feeding ecology; bioturbation; shellfish aquaculture
iii DEDICATION
To Chris, for your tireless love and support. You inspire me to take my dreams lito the moon and back".
iv ACKNOWLEDGEMENTS
I am forever grateful to Leah Bendell-Young for inspiring this work and for her role in supervising and funding this project, and to Mike Hart for his enthusiasm and constructive advice along the way. I would like to acknowledge Isabelle Cote for making me challenge myself during my defence and for the helpful comments she provided. A huge thank you goes to Tracey L'Esperance for her assistance in the field. I am so grateful to our Hornby Island family, Frances and Garth Millan, for their role in finding and providing places to live.
Thank you to Carolyn Allen for her interest in and contributions to parts of this project. Jonathan Whiteley, Chris Kowalchuk, Bruno L'Esperance, Carlos Palomera,
Jenna Thomson, Mike White, Charlotte Voss and John Driftmier volunteered their time to help in the field. Wayne Kowalchuk provided invaluable help building equipment and preparing materials for the field. Many thanks to Jonathan Whiteley, Tracey
L'Esperance, Carolyn Duckham, Carlos Palomera, Jeff Christie, Joline Widmeyer, Wade
Brunham and all the lab-mates who made my time at SFU fabulous and for supporting me through all of my graduate school adventures.
This work would not have been possible without the unwavering love, support, and encouragement of my family, the Cooks and the Kowalchuks, who have taught me to live well, laugh often, and love much. Finally, the biggest thanks goes out to Chris for encouraging me to challenge myself in everything I do, for the love, support, guidance, and patience along the way, and for being there to share in all the adventures.
v TABLE OF CONTENTS
Approval ii Abstract iii Dedication iv Acknowledgements v Table of Contents vi List of Figures viii List of Tables x Chapter 1 Introduction 1 1.1 Study Rationale 2 1.1.1 Community Structure and Function 3 1.1 .1 Predation 6 1.1.2 Bioturbation 9 1.1.3 Biology of Euspira lewisii 13 1.2 Research Objectives 15 1.3 Literature Cited 15 Chapter 2 Using Prey Preferences and Feeding Rates to Examine the Influence of Euspira lewisii on Bivalve Communities 20 2.1 Abstract 21 2.2 Introduction 22 2.3 Methods 24 2.3.1 Study Areas 24 2.3.2 Feeding Experiments 24 2.3.3 Density and Drill Collection 27 2.3.4 Community Impacts 29 2.4 Results 29 2.4.1 Prey Preference 29 2.4.2 Feeding Rates 30 2.4.3 Bivalve and E. lewisii Density and Abundance 31 2.4.4 Shell Assemblage Prey Preference 35 2.4.5 Impacts of E. lewisii Predation on Intertidal Clam Communities 37 2.5 Discussion 40 Acknowledgments 47 2.6 Literature Cited 47
vi Chapter 3 Effects of Bioturbation by Lewis's Moon Snail (Euspira lewisii) on Sediment Properties and Biological Communities in British Columbia 51 3.1 Abstract 52 3.2 Introduction 52 3.3 Methods 56 3.3.1 Study Areas 56 3.3.2 Cage Design 57 3.3.3 Sediment Characteristics 59 3.3.4 Sediment Chemistry 60 3.3.5 Biological Community 61 3.3.6 Analyses 61 3.4 Results 62 3.4.1 Physical Characteristics of the Sediment 62 3.4.2 Grain Size Analyses 63 3.4.3 Chemical Properties of the Sediment. 65 3.4.4 Biological Community 67 3.4.5 Control Cage Impacts 70 3.5 Discussion 70 Acknowledgments 74 3.6 Literature Cited 75 Chapter 4 Conclusions and Recommendations 78 4.1 Conclusions 79 4.2 Future Work 80 4.3 Recommendations 81 Appendices 83 Appendix A: Exclusion Experiment By-Tide-Height Results 83 Appendix B: Exclusion Experiment Supplementary Data 93
vii LIST OF FIGURES
Figure 2.1. E. lewisii (e) prey preference (± 95% C.I.). The dashed line represents zero preference (0.33). Values above the dashed line indicate prey preference, values below indicate avoidance. Where the C.1. does not overlap the line, preference is significant. 30 Figure 2.2. Medians and interquartile ranges of the feeding rates of E. lewisii on P. staminea, V. philippinarum and N. obscurata in clams/day/snail for each species 31 Figure 2.3. Density of clam species in number of individuals per m2 for Fillongley (A) and Shingle Spit (B) 33 Figure 2.4. The proportion of drilled shells collected from Fillongley (A) and Shingle Spit (B) compared to the proportion of clams available at each site (H-high, M-mid, L-Iow, T-total) 37 Figure 2.5. Electivity coefficients for E. lewisii feeding on the clam populations in the high (A), mid (B), low (C) and all three zones (D) at Fillongley. Negative values indicate avoidance, while positive values indicate preference 38 Figure 2.6. Electivity coefficients for E. lewisii feeding on the clam populations in the high (A), mid (B), low (C) and all three zones (D) at Shingle Spit. Negative values indicate avoidance, while positive values indicate preference 39 Figure 2.7. The number of clams consumed by E. lewisii at the rate of 0.09 clams/day at a density of 0.22 snails/m2 in 1 month, 6 months and over 12 months compared to the total number of clams available at Fillongley and Shingle Spit. 40 Figure 3.1. Map showing the location of the study sites on Denman and Hornby Islands (Based on http://atlas.nrcan.gc.ca/site/english/maps/reference/outlinecan ada/canada01, http://atlas.nrcan.gc.ca/site/english/maps/reference/outlinepro v_terr/bc_outline) 58 Figure 3.2. Compressive strength of the sediments at each study site under each treatment (Medians, error bars represent interquartile range) 62 Figure 3.3. Water content of the sediments at each study site under each treatment (Medians, error bars represent interquartile range) 63
viii Figure 3.4. Percentages of gravel, coarse sand, fine sand, and silt at each site under each treatment (Medians, error bars represent interquartile range) 64 Figure 3.5. Nutrient concentrations of ammonium, carbon and phosphorous for each treatment at each study site (Medians, error bars represent interquartile range) 66 Figure 3.6. Total invertebrate species richness for each tide height at both sites. * indicates a significant result (Medians, error bars represent interquartile range) 67 Figure 3.7. Tree diagram illustrating the Bray-Curtis similarities for the Fillongley community at all tide heights under each treatment. H =high, M =mid, L =low. E =Exclusion, CA =Control area, CC =Control cage 68 Figure 3.8. Tree diagram illustrating the Bray-Curtis similarities for the Shingle Spit community at all tide heights under each treatment. H =high, M =mid, L =low. E =Exclusion, CA = Control area, CC =Control cage 69
ix LIST OF TABLES
Table 2.1. Density of E. lewisii at Fillongley and Shingle Spit in 2 density/m ± 95% C.1. and in total abundance in the survey area ± 950/0 C.I 33 Table 2.2. Total clam abundance by species at Fillongley and Shingle Spit for each stratum ± 95% C.1. 34 Table 2.3. Raw numbers of drilled shells collected in each stratum at each site with totals 35 Table 3.1. Length of the three tide strata at each site 57 Table 3.2. Summary of the non-parametric Kruskal-Wallis analyses on the physical properties of the sediments between treatments at both sites. * indicates a significant result and ** indicates a marginally significant result. 63 Table 3.3. Summary of the non-parametric Kruskal-Wallis analyses on the grain size analyses between treatments at both study sites. * indicates a significant result and ** indicates a marginally significant result. 65 Table 3.4. Summary of the non-parametric Kruskal-Wallis analyses on the sediment nutrient characteristics between treatments at both sites. * indicates a significant result and ** indicates a marginally significant result. 66
x CHAPTER 1 INTRODUCTION 1.1 Study Rationale
The loss of biodiversity has come to the forefront of science recently for both scientists and the public. As the human population grows, more demands are put on our coastal and marine ecosystems that result in alterations in marine communities, habitat loss and bioinvasions. The most important part of this issue is to try to link and understand the interplay between the function and the structure of an ecosystem as we lose the structure in the form of biodiversity.
Over 70% of the sea floor is soft-sediment habitats and hence can be considered one of the more important habitat types (Lohrer et al. 2004). These habitats provide nurseries, are sites for nutrient exchange with the water column, and provide food for marine organisms from all levels of the food chain. It is imperative to fill knowledge gaps and gain understanding on the role of individual species in these functions.
, studied how Euspira lewisii (Lewis's moon snail), an intertidal, soft sediment predator of bivalves, influences the populations of its prey species and other species that share its habitat in British Columbia (B.C.), Canada. The goal of the work is to determine the role of this species in structuring communities as both a predator and a bioturbator.
This work also allows the opportunity to study a species (E. lewisiI) that is thought to have negative impacts on shellfish aquaculture (Bernard 1967). The
B.C. Shellfish Growers Association (BCSGA) (2002) suggests that E. lewisii is a predator to the commercially valuable Manila clam, Venerupis philippinarum. The
BCSGA Code of Practice (2002) states:
2 "A few select species (including starfish, Japanese drills, moon
snails, crustaceans, and some birds) can have significant economic
impact depending on their frequency and the type of farm operation.
Farmers are entitled to take reasonable steps to prevent the destruction of
their crops by pests and predators."
This has led to shellfish farmers actively removing E. lewisii from the intertidal zone. This is of some concern because very little is known about the role or function of E. lewisii in the intertidal community. Thus, information generated by this study will fill the knowledge gap and can be used to advise shellfish aquaculture activities and ensure a sustainable industry.
1.1.1 Community Structure and Function
A community is made up of a group of populations that live and interact in a given area (Krebs 2001). Communities have a set of five characteristics unique to this level of organization that help to study and understand them. Krebs (2001) defines these characteristics as growth form and structure, diversity, dominance, relative abundance and trophic structure. Community structure can be defined as how the populations in a given area are organized (Krebs 2001). The structure of a community can be physical or biological. Species composition and abundance, temporal changes, and relationships between species are all involved in the biological structure of a community. Species composition and abundance can be put under the umbrella of biodiversity. Predation, competition, herbivory, and biological disturbance are the relationships that influence the structure of a community and may influence biodiversity at a local scale.
3 Organisms living in sediments create much of the structure in soft
sediment habitats (Thrush & Dayton 2002). Burrows, tubes, mounds and other alterations to the sediment comprise this physical structure. Organisms that provide this habitat structure often have important roles in sequestering and recycling processes essential to ecosystem function (Thrush & Dayton 2002).
Small-scale disturbances by benthic feeding organisms can increase 3-D structure of habitat (Thrush & Dayton 2002). The physical and biological structures are closely related and strongly influence each other.
Physical and biological attributes also influence the function of a community, i.e., how energy and nutrients are processed within a community.
Nutrient cycling and primary and secondary production are all ecosystem functions (Krebs 2001). How an ecosystem functions is in part an outcome of the metrics that define that structure such as species richness and evenness
(Raghukumar & Anil 2003). Soft-sediment marine organisms have functional roles crucial to many ecosystem processes: protein supply to ecosystems, sediment stability, water column turbidity, nutrient and carbon processing (Thrush and Dayton 2002).
It is important to recognize that many recent studies have focused on the importance of maintaining biodiversity and function in marine systems. Previous work has shown that decreased ecosystem function occurs when there is a decrease in biodiversity (Lohrer et al. 2004). Heterogeneity is important in ecosystem function and makes for stable communities (Thrush & Dayton 2002).
Losing one species could have large impacts on marine systems including
4 function (Lohrer et al. 2004). Duarte (2000) showed that similar seagrass species may have different functions, so the species involved in each ecosystem function are important, not just the number of species. Removing ecosystem engineers was found to influence both biological diversity and ecosystem function (Coleman
& Williams 2002).
When trying to understand the interplay between ecosystem structure and function, the contribution of individual species to a specific function is difficult to assess (Lohrer et al. 2004). For example, Chalcroft & Resetarits (2003) found that six different predators on anuran larvae each had different impacts on their measured response variables of prey biomass, total prey number, prey species richness and prey evenness. They concluded that grouping species by function might lead to poor understanding of communities and that losing one predator species might result in loss of ecosystem function but it is difficult to differentiate each predator's role in this system. Within the intertidal region, E. lewisii can reach large populations, but unique aspects of this species are its relative size, mobility and deeper burial depth relative to other invertebrates within the same region. Hence, it may be possible due to these attributes to discern the role this species has on ecosystem function, specifically in the intertidal. This becomes of acute importance, in light of the culling of the moon snail from beaches.
Predators and predation activities are likely to influence the structure and function of their community (Thrush & Dayton 2002). The manual removal of E. lewisii has been recommended without knowing the functional role of the moon snail in the intertidal. For any system, biodiversity is important as is
5 understanding the role of each species in ecosystem processes (RClghkumar &
Anil 2003). Increasing concern about alterations to diversity of various life forms makes it necessary now for management to understand the relation between biodiversity and ecosystem functioning in our coastal and offshore waters
(Raghkumar & Anil 2003).
1.1.1 Predation
Moon snails are predators on clams in the intertidal. The impacts of E. lewisii as a predator on clam populations are thought to be quite large demonstrated by their inclusion on the SCSGA (2002) list of species of economic threat. Several studies have shown that predation influences the abundance, composition, distribution, and productivity of infaunal prey species (Seal 2006;
Seal et al. 2001; Como et al. 2004; Gee et al. 1985; Palomo et al. 2003; Peitso et al. 1994; Quijon & Snelgrove 2005; Wiltse 1980). In soft-sediment communities, in the absence of a predator Menge et al. (1994), Peterson (1979), and Wiltse
(1980) found that prey biomass and abundance increased. In Maine, predation was shown to be the most important factor affecting the survival of juvenile clams
(Seal 2006). Under severe predation pressures, the densities of all the prey populations would decrease, leading to a decrease in diversity in that community
(Virnstein 1977).
However, limited studies have been conducted on the basic feeding ecology of E. lewisii and its impacts on intertidal clam populations are unclear.
The response of a natural community to predation is influenced by the relationship between prey preferences and the abundance, competitive abilities
6 and rates of increase of the prey species (Wiltse 1980). To best understand the
functioning of intertidal communities and the role of a predator in those communities, a knowledge of the patterns of foraging activity and rates of feeding of the major predators is crucial (Moran 1985; Thiel et al. 2001). Higher feeding rates lead to higher community impacts (DeGraaf & Tyrrell 2004). Most predators increase feeding rate as the density of prey increases, but feed at a decelerating rate, reaching a plateau when prey are very dense (Moran 1985; Thiel et al.
2001). Other predators reduce their prey intake when offered low numbers of preferred prey (Thiel et al. 2001). Species and size of prey also affect feeding rates (DeGraaf & Tyrrell 2004; Moran 1985; Thiel et al. 2001). Abiotic conditions such as temperature, tide height, intensity of wave action and duration of submersion affect feeding rates, as do other predator activities such as breeding or sheltering (Moran 1985; Thiel et al. 2001). Predation is often most intense in warmer temperatures (Weissberger 1999). The physical conditions of the habitat also impact predation: in a physically stressful environment, prey populations prioritize adapting to the physical regime over adapting to biological interactions
(Byers 2005; Virnstein 1977).
The effects of predation decrease when the prey species is able to avoid or escape predation (Byers 2002; Smith et al. 1999; Tallqvist 2001). The three dimensional nature of soft-sediment habitats allows some burrowing species to escape predators by burrowing deep into the sediment while other species, such as the razor clam, display elaborate escape responses (Byers 2002; Schneider
1982; Smith et al. 1999). With all these factors influencing the intensity of
7 predation, it is difficult to make generalizations about the effects of predation on infaunal communities. Predation needs to be understood to reliably evaluate ecological impact of predatory species (Savini & Occhipinti-Ambrogi 2006).
Not all effects of predation are negative. Predator effects on prey populations depend on the intensity of predation in that community (Palomo et a!.
2003; Virnstein 1977; Wiltse 1980). When predation pressure is low, predators can reduce the number of the dominant species that leads to the competitive release of other species (Ambrose 1984; Gee et a!. 1985; Quijon & Snelgrove
2005). Species densities may even be brought to a level below which competitive exclusion occurs (Virnstein 1977). Predators feeding on discrete areas of the intertidal would result in patches in the community that are at various stages of succession increasing community diversity (VanBlaricom 1982).
It is important to note that most soft-sediment predation studies have mainly focused on epibenthic predators such as birds, crabs and fish. It must be considered that many benthic infauna are themselves predators, and can influence abundances of other infaunal species in their own ways. Predatory infauna may influence abundances of other infaunal organisms by preying on adults, juveniles, or larvae, or by disturbing the sediment surface and reducing larval settlement and juvenile survivorship (Ambrose 1984). Infaunal predators may cause more damage to prey populations by injuring prey rather than consuming them (Gee et a!. 1985). These predators are usually small, inconspicuous and commercially important (Ambrose 1991). Their smaller size
8 tends to lead to lower feeding rates, which suggest that the role of infaunal
predators is less than that of epibenthic predators (Ambrose 1991).
1.1.2 Bioturbation
Bioturbation is the mixing of sediment from the action of infauna, epifauna, fish and mammals (Biles et al. 2002). Benthic infauna are major bioturbators of the sediment in marine and estuarine habitats. The burrowing and feeding activities of E. lewisii make it a bioturbator. It is well established that particular species of bioturbators have important roles in providing nutrient regeneration and structure to an otherwise homogeneous substratum (Dayton 1984). Infaunal species differ in their feeding behaviour and mode of movement consequently creating different levels of disturbance to the sediment structure. Bioturbators impact both the physical and chemical properties of the sediment and could therefore impact the distribution of organisms living within the sediments. The removal of E. lewisii from the intertidal by shellfish growers could have large implications to the properties of the intertidal as broad-scale losses of benthic bioturbators have been shown to impair marine ecosystem functioning (Lohrer et al. 2004).
Bioturbation leads to particle redistribution and resuspention (Escapa et al.
2004; Katrak & Bird 2003; Widdows & Brinsley 2002). Larger animals, for example E. lewisii, playa particularly important role in influencing sediment reworking rates. Typically animals increase particle exchange between water and sediment by a factor of 2-10 (Thrush & Dayton 2002). Burrowing crabs trap fine grained and cohesive sediments that stabilize the sediment decreasing the
9 bedload transport. These fine sediments may act to protect the sediments against evaporation during low tide, increasing water content and humidity in the sediment (Escapa et al. 2004). The fine sediment also increases sediment softness (Palomo et al. 2003). Bioturbation affects stability and composition of marine sediments and influences their role as geochemical sources and sinks
(Thrush & Dayton 2002).
Porosity generally decreases with depth due to sediment compaction; however, burrowing organisms mix the sediment and increase porosity by altering the size of interstitial spaces within the sediments and breaking up the cohesive sediment structure (Katrak & Bird 2003). This leads to increased water content and permeability while sediment hardness decreases, which enhances the movement of water between sediment grains (Escapa et al. 2004; Lohrer et al. 2004; Palomo et al. 2003; Snelgrove 1999; Widdows & Brinsley 2002).
Infaunal organisms that construct burrows increase porosity of the sediments by pumping water through their burrows and manipulating sediments (Katrak & Bird
2003).
Studies have shown that large, deposit-feeding, bioturbating organisms dominate sediment reworking processes and related effects on sediment biogeochemistry (Katrak & Bird 2003). Benthic habitats supply up to half the nutrients for primary production in coastal seas, with ammonium being particularly important to nitrogen-limited marine waters (Lohrer et al. 2004).
Sediment disturbance and particle erosion through burrowing, feeding and movement enhance both (1) the direct release of nutrients sequestered in
10 porewater and (2) nutrient cycling back to the water column (Biles et al. 2002;
Katrak & Bird 2003; Thrush & Dayton 2002). Bioturbation also helps to decrease sulphide and ammonium concentrations in the sediments (Katrak & Bird 2003).
The activities of the burrowing urchin, Echinocardium led to the release of NH4-N from the sediments, which is linked to increases in primary production (Lohrer et al. 2004). Active burrowers such as E. lewisii would lead to the highest release of nutrients (Biles et al. 2002).
Burrowing and burrow construction increase the oxygen levels in the sediment (Biles et al. 2002; Coleman & Williams 2002; Katrak & Bird 2003;
Snelgrove 1999; Thrush & Dayton 2002). Increased oxygen levels as well as enhanced microbial activity caused by increased sediment surface area from burrowing help with the breakdown and recycling of organic matter (Biles et al.
2002; Coleman & Williams 2002; Katrak & Bird 2003). Bioturbation improves the conditions for production by microphytobenthos and increases the concentrations of chlorophyll a in the sediment which leads to increased photosynthesis (Katrak
& Bird 2003; Lohrer et al. 2004).
The disturbance caused by bioturbation or burrow construction leads to an accumulation of organic matter (Escapa et al. 2004; VanBlaricom 1982). This means that there is more food available for the organisms in the bioturbated habitat (Escapa et al. 2004; Palomo et al. 2003). This also makes the food availability more homogeneous across a bioturbated area (Escapa et al. 2004).
Bioturbation also increases the available habitat of the intertidal by extending the depth of hospitable living conditions. For example, the irrigation of
11 burrows extends the oxic water-sediment interface into the sediments (Katrak &
Bird 2003). The deeper bioturbator organisms live, feed or burrow in the sediments, the more impact they will have on the physical and chemical characteristics of the sediment (Katrak & Bird 2003).
Physical and chemical characteristics of the sediment control the movement and zonation patterns of infauna, and habitat use by shorebirds and their consumption rate (Escapa et a!. 2004). Ambrose (1991) and Widdows &
Brinsley (2002) found that nutrient enrichment from faecal material and creation of biogenic structures can be expected to enhance densities of some infaunal species. The biogeochemical changes driven by spatangoid urchins shifted microphyte community composition towards species with high productivity per amount of pigment. Biological disturbances such as bioturbation may increase diversity (Thrush & Dayton 2002). For example, burrowing crabs can moderate the physical harshness of the upper intertidal allowing some organisms to extend their distribution to higher intertidal levels. Bioturbator activities have a positive effect on infaunal polychaetes and nematodes, increasing their prey density and availability (Escapa et al. 2004).
Bioturbatory effects on infaunal populations can influence higher level predators at the surface. The activities of bioturbators can increase the amount of area available for predation. Many species of shorebird benefit from the presence of bioturbators, providing another link between infaunal and surface communities
(Escapa et a!. 2004; Palomo et a!. 2003).
12 Not all bioturbation activities are positive. Ambrose (1991) found that sediment modification by surface active predators or predators that plough through the surface can have a negative effect on infaunal densities in some communities. Beal et al. (2001) found that bioturbatory disturbance by predators affects the growth rates of some clam species. Disturbance can cause non selective mortality of other infaunal species (Gee et al. 1985). The effects of disturbance include the burial of newly settled larvae, juveniles, and adults
(Ambrose 1984).
1.1.3 Biology of Euspira lewisii
E. lewisii is a large, infaunal snail that inhabits the west coast of North
America from southeastern Alaska to southern California (Harbo 2001; Harbo
2002). It usually inhabits protected mud, sand, gravel or cobbles beaches in the intertidal to 50 m deep in the subtidal (Harbo 2001; Sept 1999; Snively 1978). It is the largest species of moon snail in the world and can have a shell that measures up to 14 cm high. This species displays sexual dimorphism, males being smaller than the females in larger size classes (Bernard 1967). It is thought that males grow at a slower rate than females. Males also have thicker shells.
Approximately six whorls make up E. lewisils shell, one very large whorl and the remainders being small. Its muscular foot is very large and almost completely surrounds its shell. It can pull its foot completely inside its shell for protection.
Water is squeezed out of small pores along the edge of its foot and a horny operculum seals the opening (Sept 1999; Snively 1978).
13 It is a long-lived species, living 11 to 14 years (Bernard 1967). E. lew;s;; begins breeding when snails are larger than 55 mm. This species lays its eggs in a distinctive sand collar (Harbo 2001; Harbo 2002; Sept 1999; Snively 1978). The collar is formed by the curvature in the shell as it is released from the body
(Bernard 1967). The eggs are found in a central jelly layer sandwiched between two thick mucous-bonded sand coats. When in the collar, the eggs measure approximately 250 IJm in length. Much of the development occurs in the collar. It is thought that the jelly layer may provide food for the developing snails. Up to
10% of the eggs disintegrate in the collar and might also provide a food source to the larvae. The collars are laid on the intertidal and deeper waters in the spring and summer with a peak in density occurring in May and June (Harbo 2001;
Harbo 2002; Sept 1999; Snively 1978). Each collar contains close to a million eggs and close to half a million hatch out of the collar. The collar disintegrates approximately 6 weeks after being constructed, and the larvae are released as a veliger larva during high tide (Bernard 1967). There is some discrepancy at this point as to what happens to the larvae. Some say that the larvae are often associated with VIva spp. which serves as a food source for the developing larvae. After this, they enter their carnivorous stage.
E. lew;s;; is a predator of bivalves that ploughs through the sediment in search of its prey (Bernard 1967; Harbo 2001; Harbo 2002; Sept 1999; Snively
1978). They attack by drilling through the shell of their prey using a toothed radula assisted by secretions from an accessory boring organ. This leaves a distinct counter-sunk hole unique to this species. Protothaca stam;nea (the
14 Pacific littleneck clam), Saxidomus gigantea (the butter clam), Mya arenaria (the
softshell clam), and Macoma nasuta (the bent-nose macoma) are species that
are commonly found with drill marks (Bernard 1967).
1.2 Research Objectives
The goal of this research was to broaden the knowledge base on the ecology of E. lewisii. This was achieved by addressing three objectives. The first objective was to determine the feeding ecology of E. lewisii through an examination of prey preference and feeding rates. Secondly, I examined the role of E. lewisii as a bioturbator. This was done through an exclusion experiment to look at how this species influences the physical and chemical properties of the sediment. The final objective was to use the information collected in each section to determine the impacts that E. lewisii predation and bioturbation have on infaunal community structure. This information is especially pertinent now given that shellfish growers are removing E. lewisii. The information collected will fill knowledge gaps on this species and demonstrate the importance of understanding the role of each species in an ecosystem to better comprehend ecosystem function. Such information can then be used to advise the shellfish industry.
1.3 Literature Cited
Ambrose WG,: ,Jr. 1984. Role of predatory infauna in structuring marine soft bottom communities. Marine Ecology Progress Series 17(2):109-15.
Ambrose WG,: ,Jr. 1991. Are infaunal predators important in structuring marine soft-bottom communities? American Zoologist 31 (6):849-60.
15 BCSGA. British Columbia Shellfish Farming Industry - Environment management system code of practice.
Beal BF. 2006. Biotic and abiotic factors influencing growth and survival of wild and cultured individuals of the softshell clam (Mya arenaria L.) in eastern Maine. Journal of Shellfish Research 25(2):461-74.
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17 Moran MJ. 1985. Effects of prey density, prey size and predator size on rates of feeding by an intertidal predatory gastropod Morula marginalba Blainville (Muricidae), on several species of prey. Journal of Experimental Marine Biology and Ecology 90(2):97-105.
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18 Snelgrove PVR. 1999. Getting to the bottom of marine biodiversity: Sedimentary habitats - Ocean bottoms are the most widespread habitat on earth and support high biodiversity and key ecosystem services. Bioscience 49(2): 129 38.
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19 CHAPTER 2 USING PREY PREFERENCES AND FEEDING RATES TO EXAMINE THE INFLUENCE OF EUSPIRA LEWISII ON BIVALVE COMMUNITIES1
1 The following chapter has been submitted to the Journal of Experimental Marine Biology and Ecology under the co-authorship of Leah Bendell-Young. 20 2.1 Abstract
The predatory naticid snail Euspira lewisii, native to the west coast of
North America, is stated to be an economic threat to the shellfish aquaculture
industry in British Columbia (B.C.). This species is being manually removed from the intertidal ecosystem, yet little is known about the ecology of this species.
Enclosures and beach shell assemblages were used to determine the prey preference, feeding rates and community impacts of E. lewisii. Protothaca staminea, the native little neck clam, was found to be the preferred prey, while the commercially valuable Manila clam, Venerupis philippinarum, was avoided.
Drilled shells collected from the intertidal revealed similar feeding preferences.
The feeding rate on a variety of species was found to be 0.09 clams/day or 1 clam every 14 days. The feeding rate was dependent on prey species and was highest for the preferred species and significantly lower on avoided species. The overall impact of E. lewisii to the bivalve community was found to be extremely low. Based on these results, E. lewisii consumed only approximately 3% of the clam population over one year, assuming maximal feeding rates and typical population densities found on the west coast of B.C. E. lewisii has minimal impacts to the Manila clam industry in B.C. and control measures are not necessary for this species. Baseline ecological field studies are important for gaining understand of poorly understood species, especially those considered threats to industry.
21 2.2 Introduction
Predation is one of the most important factors effecting community
structure in intertidal communities. It can affect the distribution pattern, size and age composition and abundance of prey species (Beal 2006; Peitso et al. 1994).
Recent studies have stressed the importance of a full understanding of predation such that we can evaluate the ecological impact a predator has on a community
(Savini & Occhipinti-Ambrogi 2006). The key to understanding the role a predator plays in a community includes knowing its prey preferences and feeding rates
(Moran 1985; Thiel et al. 2001). From an applied aspect, without a full understanding of predation it is difficult to manage intertidal communities or know if antipredator practices, such as predator removal, are effective (Miron et al.
2005).
Many intertidal predators demonstrate prey preferences and select prey that is the quickest to handle and consume to maximize their net energy intake
(Savini & Occhipinti-Ambrogi 2006). The effects of selective predation on community structure vary with relative abundance of prey species (Moran 1985) and the escape abilities of the prey species. Selective feeding on non-dominant species can have adverse effects on the community such as decreasing species diversity by removing rare species (Wiltse 1980b).
Feeding rates of predators depend on a number of biotic and abiotic factors. Biotic factors include prey biomass, density, species, quality, and predator and prey size (DeGraaf & Tyrrell 2004; Moran 1985; Thiel et al. 2001).
Time spent on other activities such as mating or predator avoidance also
22 influences feeding rates (DeGraaf & Tyrrell 2004; Thiel et al. 2001). Abiotic factors such as temperature, season, wave action, and duration of submersion
(Moran 1985; Weissberger 1999) also affect feeding rates. Greater feeding rates can lead to greater impacts on the prey community such as reduced abundance of the prey species (DeGraaf & Tyrrell 2004; Savini & Occhipinti-Ambrogi 2006).
Moon snails are infaunal, predatory snails that feed on bivalves. Several species of moon snails have shown both size and species preferences while feeding (Bernard 1967; Commito 1982; Dietl & Alexander 1997; Peitso et al.
1994; Rodrigues et al. 1987; Wiltse1980b). Through drilling activities, very clear artefacts of the predation of these species are left in intertidal habitats. For this reason, they are considered pest species, especially to shellfish aquaculture
(BCSGA 2002; Beal 2006; Bernard 1967; Peitso et al. 1994). However, little is known about the predation pressure of the moon snail on bivalve populations. A review of the literature suggests that bivalve mortality attributed to moon snails may in fact be overestimated (Beal et al. 2001; Miron et al. 2005; Peitso et al.
1994; Wiltse 1980a).
On the west coast of B.C., the native moon snail, E. lewisii is being actively eliminated from shellfish farms, based on the assumption that they are effective predators. Hence, the objectives of this study are to assess the impacts of predation by E. lewisii on bivalve communities with special emphasis on the commercially valuable Manila clam, Venerupis philippinarum. We use both field experiments and the collection of drilled bivalve shells to determine moon snail prey preference, feeding rates, and impacts on the prey community.
23 2.3 Methods
2.3.1 Study Areas
Field research was conducted in southern B.C. at Fillongley Provincial
Park, on Denman Island (49°31'59"N, 124°49'0"W) and Shingle Spit, on Hornby
Island (49°31 'O"N, 124°37'59"W). Both sites are home to known populations of E. lewisii. Venerupis philippinarum, the commercially valuable Manila clam and
Protothaca staminea, the native Pacific littleneck clam dominate the bivalve community at these sites. Nuttallia obscurata, the varnish clam, a recent introduction to southern B.C., as well as several other clam species are also found at these sites.
2.3.2 Feeding Experiments
Cage Design
We used enclosure experiments, i.e. cages, to determine the prey preferences of E. lewisii. The cages were made of PVC pipe frame measuring
2 1x1xO.3m, and enclosed an area of 1m . All sides of the frame were covered with
2 plastic mesh with an aperture of 1cm . The cages were dug into the sediment to a depth of 0.2m, leaving 0.1 m exposed at the surface. Sediment was sieved back into the cage and all bivalves and drilled shells were removed. A 4 by 3 grid was created, using 12 cages, oriented parallel to the water line. The cages in the grid were spaced approximately 2m apart. Studies were conducted from May to
September in 2005 and 2006.
24 Prey Preference
Three clam species collected from Fillongley were used in the
experiments to analyze the prey preferences of E. lewisii: P. staminea, V.
philippinarum and Nuttallia obscurata. Twenty clams of each species were buried
in each cage, five individuals of each species in each corner. This led to 60
clams in each cage and 720 in all 12 cages. This was in the range of clam
densities found at this site. Two cages, selected at random, served as controls that contained only clams and no snail that tested for clam transplant
survivorship. In the ten remaining cages, a single moon snail, collected from the site, was measured and buried into the centre of the cage. All cages were sealed and left.
The cages were checked every other tide cycle, approximately once every three weeks, throughout the course of 4 months and all drilled and dead clams were removed and replaced with live individuals of the appropriate species. Only completely drilled shells were used in the prey preference analyses.
Manly's a was used as an index of preference for constant prey populations (see Krebs 1999).
where: aj = Manly's a (preference index) for prey type i
Tj, '1 = proportion of prey type i or j in the diet (i and j = 1,2,3,.. .m) nj, nj = proportion of prey type i or j in the environment m =number of prey types possible
25 Similar preference experiments have used this index (Dudas et al. 2005) and it is well established in the feeding preference literature (Krebs 1999; Manly 1974;
Manly et al. 1972). The interpretation of the 0 values for this index are:
OJ = 1/m = no preference for species i OJ> 1/m = preference for species i OJ < 1/m =avoidance of species i where m =number of prey species.
For these experiments, three species were used therefore an 0 value of
0.33 indicates no preference, >0.33 is an indication of preference and <0.33 is an indication of avoidance. These 0 values are considered significant if the 95% confidence intervals does not overlap the 0.33 prey types.
Feeding Rates
Feeding rates were determined in tandem with the prey preference data.
Feeding rates were calculated as the # clams consumed/# days the moon snail was contained within the cage.
By-species feeding rates were also determined. Due to time constraints in
August of 2006, a single trial was carried out where the 12 cages were randomly selected to contain one of each of the three species. Fifteen individuals of each species were buried in each of the four corners of the cage Le., 4 cages/species,
240 clams/species for a total of 720 clams. Snails were added as described above and cages were sealed for approximately 3 weeks. After the 3 weeks, all the cages were checked and any drilled shells were removed and tallied.
26 A Kruskal-Wallis test was applied to determine significant differences among species feeding rates on these three clam species.
2.3.3 Density and Drill Collection
Density surveys were conducted at both the Fillongley Provincial Park site and at Shingle Spit. To account for tidal influences a 60m wide strip representative of the intertidal communities was stratified into tide heights by dividing into a high (2.3-1.7m above chart datum), mid (1.7-1.3m above chart datum) and low (1.3-0.7m above chart datum) zone.
Survey Design
Within each stratum 4 and 3-60m long transects were randomly selected at Fillongley and Shingle Spit respectively. Along each transect 6 quadrat locations were selected at random. Random numbers were selected using a random number table. At each coordinate, a 0.5 by 0.5m quadrat was dug
2 (0.25m ) down to a depth of 0.2m. All sediment dug from the quadrat was sifted through a 6mm mesh and all infaunal bivalves were identified and counted to determine community composition and densities.
During the sifting process any shells containing the distinct counter-sunk
E. lewisii drill marks were removed and the clam species was identified (Peitso et al. 1994). All live organisms and drilled shells were replaced post sampling.
Euspira lewisii densities were determined using a mark-recapture technique. Fifty individual snails were marked by scratching a number into their shell then the snails were buried back into the sediment. After three weeks, we
27 returned and dug up 30 snails and determined the number of marked snails. The total E. lewisii population was calculated based on Bernard (1967) as follows:
T =M/(R/C)
Where: T = total population in the area M = # marked animals in 1st sample R =# marked animals in 2nd sample C = total caught in 2nd sample.
Prey Preferences from Beach Shell Assemblages
The density measurements and drills collected were used to determine if
E. lewisii prey preferences were also evident under natural conditions.
Proportions of the clams were calculated based on a stratified multi-stage design
(Krebs 1998; Schwarz 2005). The proportions of shells and species in the community were also used to calculate electivity coefficients (E) based on Ivlev
(1961):
E = (r- p)l(r + p)
Where: r =proportion of a food item in the diet p =proportion of the food item in the environment
Preference is indicated by a positive value of E, avoidance is indicated by a negative value and no preference is indicated by a value of O.
Ivlev's electivity coefficient was selected because of the variable nature of the bivalve communities in the intertidal. Manly's a is appropriate for constant prey populations or in experimental situations where the prey is being replaced maintaining a constant supply of food (Krebs 1999). It is also not recommended
28 that 0 values calculated based on populations with different numbers of prey types (Krebs 1999).
2.3.4 Community Impacts
Density measurements and average feeding rates were used to represent the effects of E. lewisii predation on these intertidal communities. The average feeding rate was used to calculate the number of clams consumed in a month, in
6 months and in a year based on:
# consumed =(feeding rate) x (days) x (# snails)
2.4 Results
2.4.1 Prey Preference
When offered equal numbers of P. staminea, V. philippinarum and N. obsGurata, E. lewisii showed significant preference for P. staminea (0 = 0.57, P <
0.05, Figure 2.1). N. obsGurata was preferred although it was not statistically significant (Figure 2.1). A significant avoidance was observed for V. philippinarum (0 =0.07, P < 0.05, Figure 2.1).
29 0.7
0.6
...-.. 0.5 ---ij Q)>< ""0c 0.4 ~ c ------~ 0.3 ~ ~ a.. 0.2
0.1
0.0 Protothaca Venerupis Nuttallia staminea philippinarum obscurata
Species
Figure 2.1.E. lewisii (e) prey preference (± 95% Col.). The dashed line represents zero preference (0.33). Values above the dashed line indicate prey preference, values below indicate avoidance. Where the Col. does not overlap the line, preference is significant.
2.4.2 Feeding Rates
The average summer feeding rate of E. lewisii consuming a variety of prey species was found to be O.09±O.02 clams/day (± 95% C.I.), 1 clam consumed every 14 days.
When the feeding rates were analyzed for each of the three species individually, the consumption rate on P. staminea was greater than that on N. obscurata, which was greater than the rate on V. philippinarum (Figure 2.2).
30 0.14
.- 0.12 >, ro "'C (f) 0.10 -E 0 ro (,) --Q) 0.08 fl.-- 0> c: 0.06 I "'C Q) Q) u... 0.04 I 0.02 P.staminea V. phi. N. obscurata
Species
Figure 2.2.Medians and interquartile ranges of the feeding rates of E. lewisii on P. staminea, V. philippinarum and N. obscurata in clams/day/snail for each species.
The Kruskal-Wallis test showed that the feeding rates in clams per day were significantly different (H = 6.17, P < 0.05, Figure 2.2). The feeding rate on V. philippinarum was significantly different from that of P. staminea. N. obscurata was not significantly different from either species (Wilcoxon p <0.05).
2.4.3 Bivalve and E. lewisii Density and Abundance
At both sites, the total density/m2 decreased as the tide level decreased
(Figure 2.3A & B). The density of V. philippinarum was greatest in the high tide zone and decreased through the other strata to the water line. However, at both sites the density followed the same pattern of being highest in the mid-intertidal,
31 followed by the high zone and was the least dense in the low zone. Nuttallia obscurata was found in very low densities in the study areas, and was only found in the high, and to a lesser extent in the mid-tide zone (Figure 2.3A & B).
Macoma spp. was found in much higher densities at Shingle Spit and at both sites, it was at its highest densities in the mid and low strata (Figure 2.3A & B).
The other species we found at both sites were Mya arenaria, Saxidomus gigantea, Parvaleucina tenuisculpta, and Rhamphidonta retifera. Tellina carpenteri, Clinocardium nuttalli and Lyonsia californica were exclusively found at
Fillongley while Tresus nuttallii and Cryptomya californica were only found at
Shingle Spit. Macoma spp. was predominantly Macoma nasuta but at smaller sizes, it was difficult to distinguish it from Macoma obliqua so both of these species were represented in these communities.
Venerupis philippinarum was the most abundant species at both sites
(Table 2.2). Protothaca staminea was the second most abundant species at
Fillongley while Macoma spp. was the second most abundant species at Shingle
Spit (Table 2.2). From the quadrat surveys, E. lewisii was found only in the lowest stratum at Fillongley at an abundance of 2000 individuals in a 60m strip of the intertidal although data variability for these measurements was very high
(Table 2.2). E. lewisii was collected in both the mid and the low strata at Shingle
Spit and rough estimates suggested abundances of 200 and 300 snails in the mid and low stratums respectively (Table 2.2). Euspira lewisii densities were more accurately estimated using the mark-recapture techniques. At both sites, the density of E. lewisii was 0.2 snails/m2 (Table 2.1). Due to the shorter, steeper
32 intertidal area at Shingle Spit, the total population was less than that of Fillongley at 700 individuals compared to that at Fillongley at 2000 individuals.
500 ..,------.,- 500 ,------, A B
400 400
N 300 300 E _ Protothaca '! staminea ~ rz.zz;j Venerupis ·00 philippinarum c Q) 200 200 [[[[[[l] Nuttallia 0 obscurata c=::J Macoma spp. IlII:'l::Im Other
100 100
o o High Mid Low High Mid Low Tide Level Tide Level
Figure 2.3.Density of clam species in number of individuals per m2 for Fillongley (A) and Shingle Spit (8).
Table 2.1. Density of E. lewisii at Fillongley and Shingle Spit in density/m2 ± 95% C.1. and in total abundance in the survey area ± 95% C.I.
Site
Fillongley Shingle Spit
2 Density (#/m ) O.2±O.2 O.2±O.1
Total population 2000±1000 700±500
33 Table 2.2. Total clam abundance by species at Fillongley and Shingle Spit for each stratum ± 95% C.1.
Site Fillongley Shingle Spit
Tide Zone High Mid Low High Mid Low
Protothaca 190000±40000 600000±200000 360000±60000 70000±20000 39000±9000 30000±10000 staminea
Venerupis 540000±80000 600000±500000 0 290000±80000 20960±20000 2000±3000 philippinarum
Nuttallia 300±600 1000±3000 0 20000±20000 0 0 obscurata
Macoma 8000±8000 40000±20000 60000±30000 10000±10000 90000±30000 180000±30000 spp.
Other 2100±2000 20000±10000 70000±20000 20000±10000 8000±4000 6000±3000
Total 750000±80000 1200000±600000 500000±70000 410000±60000 160000±20000 220000±40000
Euspira 0 0 2000±3000 0 200±300 300±500 lewisii
34 2.4.4 Shell Assemblage Prey Preference
P. staminea was the most abundant of the drills collected at Fillongley, followed by Macoma clams (Table 2.3). These were also the most common of the drilled shells collected at Shingle Spit but the abundances were reversed,
Macoma spp. being the most abundant P. staminea being second. The highest number of drilled shells were collected from the lowest tide stratum at both sites.
Table 2.3. Raw numbers of drilled shells collected in each stratum at each site with totals.
Fillol1gley Shingle Spit
High Mid Low Total High Mid Low Total
Protothaca 16 278 623 917 12 106 157 275 staminea Venerupis 1 6 2 9 1 4 2 7 philippinarum Nuttallia 0 0 0 0 1 0 0 1 obscurata Macoma spp. 0 67 176 243 0 51 334 385
Other 1 38 103 142 3 22 26 51
Total 18 389 904 1311 17 183 519 719
At Fillongley 9 species were found with E. lewisii drill marks. The "other" group included: M. arenaria, S. gigantea, P. tenuisculpta, C. nuttallii and Nucella lamellosa. The diversity in the diet of E. lewisii was slightly lower at Shingle Spit, where 6 species were consumed. The "other" group was comprised of M. arenaria and S. gigantea.
35 When the proportions of the collected drilled shells were compared to the proportions of the species available in the community E. lewisii does not take clams in direct proportion to their availability (Figure 2.5). Even though V. philippinarum represented the species available in the highest proportion, the proportion of drilled shells collected for this species were very low.
When looking at each species individually, Ivlev's electivity coefficients showed that there was a preference for P. staminea, N. obscurata, Macoma spp.,
M. arenaria, S. gigantea, P. tenuisculpta and C. nuttallii at Fillongley (Figure
2.50). V. philippinarum, R. retifera, L. californica, and T. carpenteri were avoided.
At Shingle Spit only P. staminea, Macoma spp., and S. gigantea were preferred while all other species were avoided (Figure 2.60). Differences in this feeding pattern were noted when each stratum was analyzed individually. P. staminea was a preferred prey item at both sites in every stratum with the exception of the low zone at Fillongley where it was close to the no preference line (Figure 2.5A,
B & C). S. gigantea was present only in the mid and low zones at both sites.
Whenever it was present it was a preferred prey species for E. lewisii. According to the Ivlev electivity coefficients Macoma clams were avoided at all tide heights at Shingle Spit, even though they were the most commonly collected drilled shell at that site (Figure 2.6A, B, & C). However, E. lewisii did show a preference for them when the study area was looked at as a whole. They were preferred prey in the mid and low zones at Fillongley.
36 80
.-.. ~ ~ 60 (1) Cl !9 c ~ (1) 40 0-
20 _ Protothaca staminea r:zzzJ Venerupis philippinarum £ITIIII] Nuttallia Drills-H Clams-H Drills-M Clams-M Drills-L Clams-L Drills-T Clams-T obscurata 100 c==J Macoma spp. m±lm Other 80
~ 60 (1) Cl !9 c ~ (1) 40 0-
20
o Drills-H Clams-H Drills-M Clams-M Drills-L Clams-L Drills-T Clams-T
Figure 2.4. The proportion of drilled shells collected from Fillongley (A) and Shingle Spit (B) compared to the proportion of clams available at each site (H-high, M-mid, L-Iow, T-total).
2.4.5 Impacts of E. lewisii Predation on Intertidal Clam Communities
There were close to three million clams available in a 60 m -wide strip of beach at Fillongley within the range of E. lewisii. There was on average
2 2 228clams/m . At an overall density of 0.22 snails/m in this area, E. lewisii, feeding at a rate of 0.09 clams/day, in one month approximately 6500 clams 37 would be consumed (Figure 2.7). This is 0.26% of the clam population in the study area. If these values are then converted to 6 months and 1 year of feeding in the area, E. lewisii consumes 1.61 % and 3.22% of the clam population respectively. The year values should also be considered high estimates as E. lewisii decreases its feeding rate over the winter months (Huebner & Edwards
1981; Peitso 1993).
10 A 1.0 8 - c r- r-- c (J) (J) 05 '0 '(3 0.5 iE iE (J) (J) o 00 o n n U 00 , U n >. C '5 -05 '> ~ t5 -05 (J) '---- (J) W -10 - UJ '---~ -1.0 ~ -15
1.0 c 1.0 0 c -c ~r- (J) 0.5 .~ 0.5 '0 u iE iE (J) (J) n ~ o 0.0 nn o 00 n n n U U >. U >. .~ -0.5 .~ -0.5 '-B ~ (J) (J) w -1.0 - ~ ~ W -1.0 '-'-'-
-1.5 -1.5
Species
Figure 2.5.Electivity coefficients for E. lewisii feeding on the clam populations in the high (A), mid (B), low (C) and all three zones (0) at Fillongley. Negative values indicate avoidance, while positive values indicate preference.
38 1.0 A 1.0 B C r- --C - ill .Q2 0.5 0.5 () '0 i: iE Q) Q) 0 0.0 0 0.0 0 0 ..-» .c :> -05 -0.5 '>:;=; :s () Q) - ill iIi -1.0 - ~ - ~ iIi -1.0
-1.5 -1.5 ,
1.0 C 1.0 » 0 () C c r- Q) 0.5 Q) 0.5 'u Ti i: if: Q) Q) 0 0.0 0.0 n n I 0 0 0 2: » () ':> -0.5 c -0.5 U :s Q) Q) '0 l-'-- I W -1.0 !E -1.0 - -'-- W
-1.5 -1.5 ,
Species
Figure 2.6.Electivity coefficients for E. lewisii feeding on the clam populations in the high (A), mid (B), low (C) and all three zones (0) at Shingle Spit. Negative values indicate avoidance, while positive values indicate preference.
The impacts were similar at Shingle Spit (Figure 2.7). There were fewer
2 clams total at Shingle Spit, close to 800,000 and 228 c1ams/m . At the rate
2 previously mentioned and a density of E. lewisii of 0.22 snails/m , 2010 clams are consumed in 1 month, which is 0.25% of the total clam population.
Continuing at these feeding rates, E. lewisii consumes 1.54 and 3.08% of the clam population in 6 and 12 months.
39 4e+6 .,------,
_ Fillongley c::=::J Shingle Spit
3e+6
(/) E ~ 2e+6
:j::j:
1e+6
o Total Clams 1 month 6 months 12 months
Figure 2.7. The number of clams consumed by E. lewisii at the rate of 0.09 clams/day at a density of 0.22 snails/m2 in 1 month, 6 months and over 12 months compared to the total number of clams available at Fillongley and Shingle Spit.
2.5 Discussion
The work described here found that V. philippinarum is avoided by E. lewisii, suggested by the results of both prey preference experiments and observed shell assemblages. The only other study conducted on E. lewisii prey preferences found that only 0.4% of the drilled shells collected were V. philippinarum, indicating that this species is not favoured (Bernard 1967).
Protothaca staminea was the preferred prey of E. lewisii based on our experiments. Bernard (1967), Harbo (2001), Peitso (1980), and Reid &
Gustafson (1989) also found this preference. Beach shell assemblages also confirmed a preference for P. staminea. Despite the observed prey preference, 40 the beach shell assemblages showed a diverse diet. At specific tide heights other prey were chosen including Macoma spp., P. tenuisculpta, M. arenaria, S. gigantea, N. obscurata, and C. nuttallia. The beach shell assemblages gave important indications of the prey preferences of E. lewisii. The accuracy of this data is limited in that the drilled valves of thinner shelled prey species are not likely to persist as long in the habitat.
Prey preference is common in naticid snails. Wiltse (1980b) found that
Polinices duplicatus, an east coast naticid snail, ate 13 different species but showed preferences for M. arenaria and Gemma gemma. Euspira heros was shown to favour Macoma balthica and M. arenaria (Cornmito 1982). Spisula solidissima was preferentially consumed by E. heros (Weissberger 1999). Vignali and Galleni (1986) found that Donax trunculus was the species that was most attacked by the naticids from the Piombino, Italy.
Preferences from the cage experiments may be attributed to the stratification of the three tested species within the sediment, since burial depth is a phenomenon that can affect prey preferences (Committo 1982). Venerupis philippinarum lives very close to the sediment surface due to its short siphons
(Meyer & Byers 2005). Euspira lewisii may burrow below V. philippinarum and therefore does not encounter it as readily as it does P. staminea and N. obscurata, species found deeper within the sediment.
The distribution of clams throughout the intertidal could also result in the preferences. Venerupis philippinarum lives at the higher end of the range of E. lewisii, and therefore there is limited overlap in their distributions on the intertidal.
41 However, this does not explain the observed preferences because N. obscurata lives even higher on the intertidal than V. philippinarum and was consumed to a greater extent by E. lewisii.
Prey species is known to affect feeding rates (Moran 1985; Rodrigues et al. 1987; Thiel et al. 2001; Vignali & Galleni 1986). Our by-species feeding rates show that P. staminea was preyed upon at the highest rate by E. lewisii. Nuttallia obscurata, a newly introduced species in the area, was consumed at the second fastest feeding rate. Venerupis philippinarum was the avoided prey type with the lowest feeding rate. Bernard (1967) found that E. lewisii consumed P. staminea faster than it consumed S. gigantea and T. nuttalli, which supports the conclusions of this work. Euspira heros had higher feeding rates on soft-shell clams, its preferred prey type (Miron et al. 2005).
The feeding rate of 0.09 clams per snail per day was determined for E. lewisii consuming a variety of available species. This is within the range found in previous studies (Peitso et al. 1994). Earlier studies by Bernard (1967) found the feeding rate to be 0.25 clams per snail per day. However, in the Bernard (1967) study, snails were starved for 5 days prior to experimentation, placed in tanks with a limited amount of sediment, and all attempts and partially consumed clams were used in feeding rate calculations. Studies have shown that moon snails will not return to same drill site to continue feeding on the prey item once interrupted
(Dietl & Alexander 1997; Kingsley-Smith et al. 2003). Thus, including drill attempts could have inflated the feeding rate. Peitso et al. (1994) found that the summer feeding rate was approximately 0.07 clams per snail per day, which is
42 close to our 0.09 estimates. Our rate translates to one clam consumed every 14
days, a very slow feeding rate. Previous work on moon snail feeding rates has shown a wide range of feeding rates between snail species. Euspira heros had a maximum feeding rate of 1 clam per day (Weissberger 1999). Polinices pulchellus was found to consume 14.S7 clams per snail per month at its maximum rate (Kingsley-Smith et al. 2003). Thus, feeding rates are not comparable between species.
Predator size, prey size and temperature can all influence feeding rates.
These factors must be considered when looking at feeding ecology as they can lead to an elevated feeding rate. Smaller snails have higher consumption rates
(Seal 2006; Edwards & Huebner 1977; Huebner & Edwards 1981; Kingsley
Smith et al. 2003; Peitso et al. 1994; Wiltse 1980a). Prey size can be optimized for best grip by the moon snail that facilitates drilling and increases feeding rates
(Commito 1982; Vignali & Galleni 1986; Wiltse 1980a). Peitso (1980) found significant differences between the summer and winter feeding rate of E. lewisii, the rate being highest in the summer. The rate determined in our study is a summer feeding rate. The spring, fall and winter rates are lower due to the lower temperatures. Kingsley-Smith et al. (2003) and Weissberger (1999) found that moon snail feeding rates were dependent on temperature. Many naticid snails will actually stop feeding for 4 months in the winter, as was seen in P. duplicatus
(Huebner & Edwards 1981). This species stopped feeding completely at a temperature below SoC (Edwards & Huebner 1977). Therefore, the feeding rate
43 determined in the current work is an upper limit, which must be considered when
estimating the snails impact on the community.
Prey preferences and the resulting feeding rates can be explained using the optimal foraging theory where predators consume prey that lead to the highest energy gain for the least amount of time and energy input (Boggs et a!.
1984). Naticid gastropod prey preference follows this hypothesis (Dietl &
Alexander 1997). Savini & Occhipinti-Ambrogi (2006) found that moon snails maximize their energy intake by selecting a specific prey species that they can consume efficiently, rather than the immediately available species. Euspira lewisii followed this pattern, except where its preferred prey was not readily available.
Venerupis philippinarum is the numerically dominant species at both study sites, yet it was avoided in our experiments and the beach shell assemblages, where other species are available in lower numbers.
Feeding in E. lewisii is a large investment of energy, as they must spend quite a lot of time and energy drilling through the shell of its prey before feeding actually begins. It therefore needs to find prey that will facilitate these activities.
Rodrigues et a!. (1987) found that prey was selected based on a shell morphology that eases handling and reduces energy input. Protothaca staminea's round and inflated shell morphology facilitates drilling at the umbo
(Reid & Gustafson 1989; Vignali & Galleni 1986). Variations in shell thickness lead to variations in feeding rates and handling time. Minor changes in shell thickness can lead to dramatic changes in feeding rate. In a slow feeding organism, such as E. lewisii, fractions of millimetres can increase drilling time by
44 at least 25 hours (Dietl & Alexander 1997). It may take longer to drill P. staminea due to its relatively thick shell but it contains more calories than the other two species (Kirk 2007). Although N. obscurata has the lowest energy content, it may be selected over V. philippinarum because it has a thinner shell and takes less time to drill. In P. duplicatus, drilling alone took approximately 36 hours on its preferred prey species Mya arenaria (Boggs et al. 1984). Finding exact feeding rates in burrowing snails such as E. lewisii is complicated due to not being able to directly measure drilling times.
In eastern Canada and USA, moon snail predation on commercially valuable shellfish has been considered to be high enough to warrant the use of public funds to control their populations. Beal et a!. (2001) proposed that moon snails are responsible for 96.5% of the mortality of M. arenaria. Predation is stated to be the most important factor determining juvenile clam survival in
Maine, USA, where 77% of clam mortality is attributed to the moon snail E. heros
(Beal 2006). In B.C., the code of practice (2002) put out by the B.C. Shellfish
Growers Association listed E. lewisii as one of several species that can have significant economic impact to the V. philippinarum industry. To protect their crop, the shellfish growers are removing E. lewisii from the intertidal.
Recent work has shown that feeding rates and impacts may be exaggerated. Clam deaths by crabs and other predators have been attributed to moon snails in some studies, implying that moon snail predation was over emphasized (Beal et a!. 2001). Green (1968) estimated annual mortality rates of
28.2% from skate predation and other shell destroying causes, 14.3% from
45 crowding related causes and only 4% from naticid predation and this was by two
different species. Predation by P. duplicatus was found to be only a minor source
of mortality for G. gemma, one of its preferred prey species (Wiltse 1980a). Miron et a!. (1985) found that the naticid E. heros, was the predator that had the lowest feeding rate on all clam species tested compared to two sea star predators in eastern North America. Feeding rates in P. duplicatus were found to be less than
previously believed (Huebner & Edwards 1981). Our work and the work performed by Peitso (1980) and Peitso et a!. (1994), demonstrated that the feeding rates of E. lewisii are much lower than Bernard (1967) originally found.
Our findings as with Peitso (1994) suggest that over a year about 3% of clam population mortality is due to E. lewisii predation. This study stresses the importance of understanding the feeding ecology of a predator before suggesting anti-predation measures.
E. lewisii's avoidance of V. philippinarum, low feeding rate and low impacts to the bivalve community can be applied to sustainable shellfish aquaculture practices. The results demonstrate that there is no longer a need to remove E. lewisii from intertidal lease areas, saving the time and energy of shellfish growers. The impact to the intertidal ecosystem by aquaculture activities is thereby reduced and E. lewisii can be left in place to fulfil its ecological function.
Our study and the results of recent studies can lead to the general conclusion that moon snails have very low impacts on natural clam populations through predation activities. Biases on the amounts the moon snail prey on could
46 stem from the incriminating artefacts that are left behind, the bored shell, which
numbers will accumulate over time given a false impression of the numbers of
clams actually preyed upon in a given time period. Studies prior to 1990 have
also been conducted under artificial conditions over short time periods, which
lead to predation overestimates.
Acknowledgments
Many thanks go out to Tracey L'Esperance for all her assistance in the field. Thanks also to Carolyn Allen, Chris Kowalchuk, Bruno L'Esperance,
Jonathan Whiteley, and Wayne Kowalchuk for their support and assistance on various aspects of this research. Appreciation also goes out to Jenna Thomson,
Mike White and Charlotte Voss for taking such a keen interest in the project and helping out with data collection. Mike Hart provided constructive and helpful comments throughout this research. Funding for this work was provided by an
NSERC strategic grant to L. Bendell-Young.
2.6 Literature Cited
BCSGA. 2002. British Columbia Shellfish Farming Industry - Environment Management System Code of Practice.
Beal BF. 2006. Biotic and abiotic factors influencing growth and survival of wild and cultured individuals of the softshell clam (Mya arenaria L.) in eastern Maine. Journal of Shellfish Research 25(2):461-74.
Beal BF, Parker MR, Veneile KW. 2001. Seasonal effects of intraspecific density and predator exclusion along a shore-level gradient on survival and growth
47 of juveniles of the soft-shell clam, Mya arenaria L., in Maine, USA. Journal of Experimental Marine Biology and Ecology 264(2):133-69.
Bernard FR. 1967. Studies on the biology of the naticid clam drill Polinices lewisii (Gould) (Gastropoda Prosobranchia). Fisheries Research Board of Canada Technical Report 42:1-41.
Boggs CH, Rice JA, Kitchell JA, Kitchell JF. 1984. Predation at a snail's pace: What's time to a gastropod? Oecologia (Berlin) 62(1 ):13-7.
Commito JA. 1982. Effects of Lunatia heros predation on the population dynamics of Mya arenaria and Macoma balthica in Maine, USA. Marine Biology 69(2):187-93.
DeGraaf JD and Tyrrell MC. 2004. Comparison of the feeding rates of two introduced crab species, Carcinus maenas and Hemigrapsus sanguineus, on the blue mussel, Mytilus edulis. Northeastern Naturalist 11 (2):163-6.
Dietl GP and Alexander RR. 1997. Predator-prey interactions between the naticids Euspira heros Say and Neverita duplicata Say and the Atlantic surfclam Spisula solidissima Dillwyn from Long Island to Delaware. Journal of Shellfish Research 16(2):413-22.
Dudas SE, McGaw IJ, Dower JF. 2005. Selective crab predation on native and introduced bivalves in British Columbia. Journal of Experimental Marine Biology and Ecology 325(1 ):8-17.
Edwards DC and Huebner JD. 1977. Feeding and growth rates of Polinices duplicatus preying on Mya arenaria at Barnstable Harbor, Massachusetts. Ecology 58(6):1218-36.
Green RH. 1968. Mortality and stability in a low diversity subtropical intertidal community. Ecology 49(5):848-54.
Harbo RM. 2001. Shells and shellfish of the Pacific Northwest. Madeira Park: Harbour Publishing.
Huebner JD and Edwards DC. 1981. Energy budget of the predatory marine gastropod Polinices duplicatus. Marine Biology (Berlin) 61 (2-3):221-6.
48 Ivlev VS. 1961. Experimental ecology of the feeding fishes. Scott 0, translator; New Haven: Yale University Press. 302 p.
Kingsley-Smith PR, Richardson CA, Seed R. 2003. Stereotypic and size selective predation in Polinices pulchellus (Gastropoda: Naticidae) Risso 1826. Journal of Experimental Marine Biology and Ecology 295(2):173-90.
Kirk M. 2007. Movement and foraging behaviours of surf scoters wintering in habitats modified by shellfish aquaculture. MSs Thesis, Simon Fraser University, Burnaby, B.C.
Krebs CJ. 1999. Manly's a. In: Ecological methodology. 2nd ed. Menlo Park, California: Addison-Wesley Educational Publishers, Inc. 483-486 p.
Manly BFJ. 1974. Model for certain types of selection experiments. Biometrics 30(2):281-94.
Manly BFJ, Miller P, Cook LM. 1972. Analysis of a selective predation experiment. American Naturalist 106(952):719-36.
Meyer JJ and Byers JE. 2005. As good as dead? Sublethal predation facilitates lethal predation on an intertidal clam. Ecology Letters 8(2):160-166.
Miron G, Audet 0, Landry T, Moriyasu M. 2005. Predation potential of the invasive green crab (Garcinus maenas) and other common predators on commercial bivalve species found on Prince Edward Island. Journal of Shellfish Research 24(2):579-86.
Moran MJ. 1985. Effects of prey density, prey size and predator size on rates of feeding by an intertidal predatory gastropod Morula marginalba Blainville (Muricidae), on several species of prey. Journal of Experimental Marine Biology and Ecology 90(2):97-105.
Peitso E. 1980. Predation by the moon snail, Polinices lewisii (Gould), on the littleneck clam, Protothaca staminea (Conrad). MSs Thesis, Simon Fraser University, Burnaby, B.C.
Peitso E, Hui E, Hartwick B, Bourne N. 1994. Predation by the naticid gastropod Polinices lewisii (Gould) on littleneck clams Protothaca staminea (Conrad) in British Columbia. Canadian Journal of Zoology 72(2):319-25.
49 Reid RGB and Gustafson BD. 1989. Update on feeding and digestion in the moon snail Polinices lewisii (Gould, 1847). Veliger 32(3):327.
Rodrigues Cl, Nojima S, Kikuchi T. 1987. Mechanics of prey size preference in the gastropod Neverita didyma preying on the bivalve Ruditapes philippinarum. Marine Ecology Progress Series 40( 1-2):87-93.
Savini D and Occhipinti-Ambrogi A. 2006. Consumption rates and prey preference of the invasive gastropod Rapana venosa in the northern Adriatic Sea. Helgoland Marine Research 60(2): 153-9.
Schwarz, C.J. 2005. Stat 403/Stat 650 - Intermediate sampling and experimental design and analysis - Course notes. Simon Fraser University, Burnaby, B.C.
Thiel M, Ullrich N, Vasquez N. 2001. Predation rates of nemertean predators: The case of a rocky shore hoplonemertean feeding on arnphipods. Hydrobiologia 456:45-57.
Vignali Rand Galleni L. 1986. Naticid predation on soft bottom bivalves: A study on a beach shell assemblage. Oebalia 13:157-77.
Weissberger EJ. 1999. Additive interactions between the moon snail Euspira heros and the sea star Asterias forbesi, two predators of the surfclam Spisula solidissima. Oecologia 119(3):461-6.
Wiltse WI. 1980a. Predation by juvenile Polinices duplicatus (Say) on Gemma gemma (Totten). Journal of Experimental Marine Biology and Ecology 42(2):187-99.
Wiltse WI. 1980b. Effects of Polinices duplicatus (Gastropoda: Naticidae) on infaunal community structure at Barnstable Harbor, Massachusetts, USA. Marine Biology (Berlin) 56(4):301-10.
50 CHAPTER 3 EFFECTS OF BIOTURBATION BY LEWIS'S MOON SNAIL (EUSPIRA LEWISII) ON SEDIMENT PROPERTIES AND BIOLOGICAL COMMUNITIES IN BRITISH COLUMBIA2
2 The following chapter has been submitted to Journal of Experimental Marine Biology and Ecology under the co-authorship of Leah Bendell-Young. 51 3.1 Abstract
Lewis's moon snail, Euspira lewisii, is being manually removed from intertidal ecosystems in western British Columbia (B.C.) due to its reputation as an economically detrimental species to the shellfish aquaculture industry. Little is known about the ecological role of E. lewisii and it is hypothesized that due to its burrowing activities, E. lewisii has large impacts on the physical, chemical and biological properties of the sediments. To determine the ecological role of E. lewisii an exclusion experiment was carried out. The sediment became significantly less permeable in exclusion cages. There were no significant differences in terms of sediment grain size profiles. Nutrients accumulated in exclusion areas but these trends were not statistically significant. The biological communities in exclusion cages at different tide heights became more homogenous and tide zones with more diverse communities became very similar to tide zones with lower diversity. This study stresses the importance of understanding the function of all the organisms in a community before control measures are carried out. We recommend that further studies be conducted to accurately determine E. lewisifs role in nutrient exchanges.
3.2 Introduction
Bioturbation is recognized as an important contributor to ecosystem processes including sediment modification and nutrient cycling (Lohrer et al.
2004; Thrush & Dayton 2002). Bioturbation is the dominant mode of transport in the upper centimetres of oceanic sediments. It also affects the composition of
52 marine sediments and influences their role as geochemical sources and sinks
(Thrush & Dayton 2002).
Bioturbation influences a wide range of physical, chemical and biological variables within the sediment. Grain-size distributions, shear strength, stability, sediment resuspension, sediment softness, and permeability are all physical parameters influenced by burrowing activities (Biles et al. 2002; Katrak & Bird
2003; Palomo et al. 2003). Increased permeability allows organic matter, water, and oxygen to penetrate deeper into the sediment (Biles et al. 2002; Coleman &
Williams 2002; Palomo et al. 2003; Snelgrove 1999; Thrush & Dayton 2002).
Larger animals, such as predators, playa particularly important role in sediment reworking rates resulting in increased permeability (Thrush & Dayton 2002).
Organisms that burrow and create mounds or tubes generate structure in the habitat and increase the surface area of the sediment that is in contact with the water column which helps in nutrient recycling and increases water and oxygen availability in the sediments (Coleman & Williams 2002; Katrak & Bird 2003;
Lohrer et al. 2004; Snelgrove 1999; Thrush & Dayton 2002). Increased nutrient fluxes can contribute to increased ecosystem functions such as primary production and can influence the biological community (Lohrer et al. 2004).
Bioturbation affects infaunal communities through direct disturbance and through its influences on the physical and chemical nature of the sediments.
Ploughing and moving through the sediment can smother or bury larvae or adult infauna within the sediment (Ambrose 1991; Gee et al. 1985). Bioturbators can have negative effects on infaunal densities and clam growth rates (Beal et al.
53 2001). Snelgrove (1999), however, found that disturbance through bioturbation increased infaunal diversity. The increased surface area created by burrowing activity and burrow construction provides favourable conditions for microbial activity and microphytobenthos productivity (Biles et al. 2002; Lohrer et al. 2004).
The modification of the physical and chemical properties through bioturbation can increase the three-dimensional nature of the sediment allowing more organisms to live in these areas (Katrak & Bird 2003; Palomo et al. 2003; Thrush & Dayton
2002). Small-scale disturbances that occur through burrowing and predatory activities create patches in the habitat, which increases the heterogeneity and diversity and play an important role in structuring communities (Biles et al. 2002;
Escapa et al. 2004; Raghkumar & Anil 2003).
The structure of soft-sediment habitats, including biodiversity, is tightly linked to the functioning of those ecosystems (Raghkumar & Anil 2003). Several studies have shown that decreases in biodiversity lead to loss of ecosystem function (Chalcroft & Resetarits 2003; Duarte 2000; Lohrer et al. 2004). Losing one species, especially if the species is a large, bioturbating predator can have severe impacts on ecosystem function and can influence benthic diversity
(Coleman & Williams 2002; Lohrer et al. 2004). Losing the species that influence the cycling of nutrients can have significant consequences on many ecosystem processes. To effectively manage our coastal and offshore waters, it is essential to understand the relationship between biodiversity and ecosystem functioning
(Raghkumar & Anil 2003). Understanding the role of all species in a community
54 has become more important than a simple biodiversity inventory (Raghkumar &
AniI2003).
Euspira lewisii is a large, infaunal predator of the family Naticidae found in intertidal to subtidal habitats on the west coast of North America from Mexico to southern Alaska. This species burrows through the sand at depths of 10-20cm searching for and consuming clams. Feeding rates for this species were originally thought to be high at 0.25 clams/snail/day, however, recent studies have shown that this rate may be much lower at 0.07-0.09 clams/snail/day
(Bernard 1967; Cook 2008; Peitso 1994). Shellfish managers consider E. lewisii a pest species to the shellfish aquaculture industry and for this reason it is being removed from intertidal lease areas (Bernard 1967).
Little is known as to the effects of moon snails as bioturbators. Work on this species has focused on feeding ecology or development. No work has been conducted on the effects of moon snails on the physical or chemical properties of the sediment. Wiltse (1980) showed that Polinices duplicatus, a moon snail species from the east coast of North America, decreased diversity. Species richness, evenness and heterogeneity all decreased with increasing moon snail density. Wiltse's (1980) study, as with several others, focuses only on the effects of moon snails as predators.
Here, our objective is to determine the role of E. lewisii as bioturbators of intertidal sediments. The influences on the physical and chemical properties of the sediment and on the biological community are examined. The focus is on the penetrability, water content, grain size distributions, and the ammonium, carbon
55 and phosphate concentrations of the sediment. To determine these properties an exclusion experiment was conducted to mimic the impact of the removal of E. lewisii from intertidal shellfish leases. It is expected that the exclusion cages will show decreased penetrability of the sediment, decreased water content due to the decrease in permeability, an accumulation of fine and silt sediment particles, an accumulation of ammonium, organic matter and phosphate within the sediment, and shifts in the biological community driven by the altered physical and chemical state of the sediment.
3.3 Methods
3.3.1 Study Areas
Field research was conducted in southern B.C. (Figure 3.1) at Fillongley
Provincial Park, on Denman Island (49°31'59"N, 124°49'0"W) and Shingle Spit, on Hornby Island (49°31'0"N, 124°37'59"W). Both sites are home to a known population of E. lewisii at a density of approximately 0.2snails/m2 (Cook 2008).
Venerupis philippinarum, the commercially valuable Manila clam and Protothaca staminea, the native Pacific littleneck clam dominate the bivalve community at these sites. Nuttallia obscurata, the varnish clam, a recent introduction to southern B.C., as well as several other clam species, are also found at these sites.
At each site, a 60 m wide strip of the intertidal was selected based on preliminary surveys that showed the area was representative of the intertidal area at each site. A tide height of 2.3 m above chart datum was the top of the
56 strip as this was towards the high end of the moon snail intertidal range. The strip was stratified into a high, mid and low zone in order to reduce the variability across the intertidal. Table 3.1 shows the tide heights and length of each stratum at each site.
Table 3.1. Length of the three tide strata at each site.
Zone Length (meters)
Tide Zone Tide Height Fillongley Shingle Spit (m above chart datum)
High 2.3-1.7 30 25
Mid 1.7-1.3 67 12
Low 1.3 - 0.7 80 20
Within each stratum, coordinates were selected at random for the locations of the exclusion cages, control cages, and control areas. Four exclusion cages, four control areas and two control cages were placed in 3 strata at 2 sites for a total of 24 exclusion cages, 24 control areas and 12 control cages.
3.3.2 Cage Design
Exclusion cages, 1x1 xO.3 m, enclosing an area of 1m2 with mesh having a
3cm aperture, were constructed to determine the role of E. lewisii as a bioturbator. The cages were designed so that infaunal organisms were free to enter and exit the cages while excluding E. lewisii. Cages were dug into the sediment to a depth of 0.2m, leaving 0.1 m exposed at the surface. This design mimics the impact of anti-predator netting used in aquaculture practices as large predators are excluded. Due to its infaunal nature, its large size, and its 57 British Columbia
Vancouver ' Island
Scale
IwoooI IwoooI
Figure 3.1.Map showing the location of the study sites on Denman and Hornby Islands (Based on http://atlas.nrcan.gc.ca/site/english/maps/reference/outlinecanada/canada01, http://atlas.nrcan.gc.ca/site/english/maps/reference/outlineprov terr/bc outline) distribution to depths of 20cm, E. lewisii is the species that would have the strongest effects to the infaunal communities and properties at the study sites at interest. It can therefore be assumed that any significant findings within the sediment can be attributed to the exclusion of moon snails.
Sediment was sieved back into the cage through 6 mm mesh and all macroinfauna was removed. Whiteley (2005) found that only 10% of species and 58 25% of species count data were lost using 6mm versus 1mm sieve mesh. The
larger aperture mesh also allowed for increased sampling as field researchers
were not limited by the lengthy sieving time through 1mm mesh. Control cages
used the same frame as the exclusion cages but only 3 sides were covered with
mesh to test for alterations to water and sediment flow and shading due to the
cage structure. Control areas were marked with rope that was held in place in the
four corners using rebar sunk into the sediment. Control cages, control areas and
exclusion cages were prepared in the same way differing only in the cage type or
lack of cage used. The cages were dug in May and June 2005 and sealed until the summer of 2006. In 2006, the cages were opened and dug up and data were collected on the physical, chemical and biological properties.
3.3.3 Sediment Characteristics
A Durham S-170 pocket penetrometer was used to collect sediment penetrability measurements for each cage or control area. Three measurements were taken for each replicate to account for variability within the cage itself.
Three 3.8cm diameter bulk sediment cores were taken to a depth of 10cm.
The samples were immediately put on ice and frozen for determination of water content, grain size and chemical concentrations in the lab in September 2006.
One hundred grams of each sample was weighed out, dried in a drying oven for at least 48 hours and weighed to determine percent water content. The dried sediment was separated into 4 grain size fractions through wet sieving using 3 sieves: gravel (>2mm), coarse sand (>0.5mm), fine sand (>0.0625mm).
59 Each size fraction was dried for 24 hours and weighed. The silt fraction was
calculated from the total dry weight less the weight of the three larger fractions.
Using the total dry weight the percent of each fraction was calculated.
3.3.4 Sediment Chemistry
The concentration of organic matter was determined through loss on
ignition. 0.5g of sample were weighed and dried for 24 hours in a drying oven.
The dry sample was weighed and ashed in a muffler furnace at 400°C for 1 hour.
The samples were cooled and weighed.
Ammonium concentrations within the sediments were determined using the indophenol blue method of Page (1982), a method deemed acceptable for
intertidal sediments which are between soils and marine sediments. Ten grams of sample were mixed with 2M potassium chloride. After the sediments had settled EDTA, phenol-nitroprusside solution and a buffering solution were combined and heated with 5mL of the sample. After heating for 30 minutes at
40°C, the sample absorbance was read in a spectrophotometer at 636nm.
Weights were then calculated based on the slope of the calibration curve determined before sample analysis.
The sulfuric acid - nitric acid digestion technique and
Vanadomolybdophosphoric Acid colorimetric method were used to extract phosphate from the intertidal sediments based on Greenberg (1992). Five grams of sediment was placed in a Teflon tube and 0.1 002N sulfuric acid and concentrated nitric acid were added. The samples were placed in a CEM MDS-
60 2000 Microwave for 18 minutes at 200°C. Each sample was filtered and diluted
to 100ml using distilled water. 17.5mL of the sample was mixed with Vanadate
Molybdate reagent and distilled water. The sample absorbances were read at
470nm on a spectrophotometer. The phosphate concentrations were determined
using a calibration curve.
3.3.5 Biological Community
When the cages were extracted from the intertidal, all the sediment in the
cage was sifted through a 6mm mesh. All of the macroinfauna in the cage was
identified and counted. All bivalves were measured using vernier calipers to the
nearest 0.1 mm. Measurements of species richness, evenness and the Shannon
Weiner diversity index were calculated.
3.3.6 Analyses
In order to present general trends for this experiment, the data for each treatment was pooled across all tide heights and the results presented represent the total intertidal area used in the study.
As the data were not normally distributed and could not be transformed, all statistical analyses were carried out using non-parametric Kruskal-Wallis tests.
Bonferroni corrections were applied to all the analyses which reduced the significant p-value to 0.017.
Similarities within the biological communities at the study sites were compared using a Bray-Curtis similarity index. Values close to one indicate a
61 high degree of similarity between communities while values close to zero indicate
dissimilarity. The results are displayed in a tree-diagram.
3.4 Results
3.4.1 Physical Characteristics of the Sediment
The unconfined compressive strength of the sediments at both study sites
was found to be significantly higher in exclusion cages (Figure 3.2, Table 3.2).
No significant differences were found in terms of sediment water content (Figure
3.3).
1.8 .------,
1.6 _ Exclusion [:==J Control Area 1.4 Control Cage
.r. 1.2 +-' OJ C ~ 1.0 +-' (f) Q) 0.8 > 'i/.i Cf) 0.6 ~ 0- E 0.4 o o 0.2
0.0 ...L.- _ Fillongley Shingle Spit Study Site
Figure 3.2.Compressive strength of the sediments at each study site under each treatment (Medians, error bars represent interquartile range).
62 18 .,------,
_ Exclusion c::=:J Control Area Control Cage ...... c Q) ...... c o 12 o I- Q) 10 ...... ~ 8 Q) 0) ...... ro 6 c Q) e 4 Q) 0.... 2 o --'------Fillongley Shingle Spit Study Site
Figure 3.3. Water content of the sediments at each study site under each treatment (Medians, error bars represent interquartile range).
Table 3.2. Summary of the non-parametric Kruskal-Wallis analyses on the physical properties of the sediments between treatments at both sites. * indicates a significant result and ** indicates a marginally significant result.
Chi- Study Physical Property squared p-value Significant? Site Value Compressive Strength 11.01 0.0041 * Fillongley Water Content 2.12 0.347
Shingle Compressive Strength 18.25 0.0001 * Spit Water Content 0.22 0.8948
3.4.2 Grain Size Analyses
Although we expected to see an accumulation of fine sand and silt in the exclusion cases this was not the case. We did not detect any significant trends in 63 terms of the grain size profiles of the sediment at either site (Figure 3.4, Table
3.3).
100 ...... 60 ~ 0 ~ "D---- 0 80 c 50 -, C\l ----Qj if) > C\l (j) 40 ..... 60 .....(/J (9 C\l (j) 0 30 OJ (.) C\l 40 I (j) c , OJ 20 --(j) co I U ..... 20 c (j) --(j) 10 0.. u..... (j) 0 0.. 0 Fillongley Shingle Spit Fillongley Shingle Spit Study Site Study Site 50 3.5 -,------, ~ 0 3.0 ----"0 40 ~ c o C\l 2.5 if) ----=: (j) 30 U) c (j) 2.0 u::: OJ co (j) 1.5 O'l 20 c C\l , -- c ~ 10 -- (j) (j) 10 .....u 0.. (j) 0.5 0.. 0 II Fillongley Shingle Spit Fillongley Shingle Spit Study Site Study Site _ Exclusion c:::=J Control Area Control Cage
Figure 3.4.Percentages of gravel, coarse sand, fine sand, and silt at each site under each treatment (Medians, error bars represent interquartile range).
64 Table 3.3. Summary of the non-parametric Kruskal-Wallis analyses on the grain size analyses between treatments at both study sites. * indicates a significant result and ** indicates a marginally significant result.
Study Physical Property Chi- p-value Significant? Site squared Value Percentage of Gravel 5.14 0.0765
Percentage of Coarse Sand 5.16 0.0756 Fillongley Percentage of Fine Sand 2.99 0.2236
Percentage of Silt 0.22 0.8942
Percentage of Gravel 1.77 0.4132
Shingle Percentage of Coarse Sand 4.9 0.0865 Spit Percentage of Fine Sand 0.09 0.9536
Percentage of Silt 0.02 0.9919
3.4.3 Chemical Properties of the Sediment
There were no statistically significant trends in terms of nutrient concentrations at the Fillongley field site (Figure 3.5, Table 3.4). There are indications of slight accumulations of ammonium and carbon in exclusion cages.
At Shingle Spit no significant accumulations of nutrients were detected but, as was seen at Fillongley, there is a possibility of carbon and ammonium accumulations in exclusion cages (Figure 3.5, Table 3.4).
65 ---. OJ 0.012 OJ 0.18 .------, OJ --E --....- --E 0.16 c 0.010 0 -- 0.14 :;:::; c C\l 0.008 ...... 2 0.12 c ~ Q) C 0.10 0 0.006 T c ~ 0 0.08 U c 0.004 E 8 006 ::J § 0.04 'c 0.002 0 ~ E 0.02 E 0.000 U 0.00 -'---.....,r'-L'------"a...y..J.L-----.J « Fillongley Shingle Spit Fillongley Shingle Spit
---. Study Site Study Site OJ 0.16 --OJ E 0.14 c 0 :;:::; 0.12 C\l .-.... 0.10 _ Exclusion c Q) T c::::=:::::J Control Area 0 0.08 T c Control Cage 0 ;:r:; U 0.06 Q) .- 0.04 .cC\l Q.en 0.02 0 .c 0.00 0... Fillongley Shingle Spit Study Site Figure 3.5.Nutrient concentrations of ammonium, carbon and phosphorous for each treatment at each study site (Medians, error bars represent interquartile range).
Table 3.4. Summary of the non-parametric Kruskal-Wallis analyses on the sediment nutrient characteristics between treatments at both sites. * indicates a significant result and ** indicates a marginally significant result.
Study Site Nutrient Chi- p- Significant? squared value Value Ammonium 2.58 0.2754
Fillongley Carbon 2.4 0.3017
Phosphorous 6.23 0.0444
Ammonium 3.26 0.1955 Shingle Carbon 3.97 0.1372 Spit Phosphorous 5.94 0.0512
66 3.4.4 Biological Community
At Fillongley, the high and mid exclusion areas were closely similar
(Figure 3.7). Included in this grouping yet less similar were the control area and control cage for the high zone. Excluding E. lewisii makes the community more similar to that found in the high zone, a community of lower diversity (Figure 3.6).
Based on baseline density measurements these communities show marginal significant differences (Wilcoxon, i =5.05, P =0.02). The removal of moon snails from the low zone had the lowest impact. The communities in the exclusion areas in the low zone were similar to communities of the control cages and control areas of the low and mid zones.
12 _ High * 10 c:::==J Mid Low
(/) (/) Q) 8 c .c () ex: 6 C/) Q) 'u Q) Q. 4 (f)
2
o -L..- _ Fillongley Shingle Spit Study Site
Figure 3.G.Total invertebrate species richness for each tide height at both sites. * indicates a significant result (Medians, error bars represent interquartile range).
67 0.0
0.2
...c Q) 0.4 "0 !E Q) 8 ~ ·C co "E 0.6 Ci5
I I 0.8 I I
1.0 HE ME HCC HCA LE LCC LCA MCA MCC Treatment Figure 3.7.Tree diagram illustrating the Bray-Curtis similarities for the Fillongley community at all tide heights under each treatment. H =high, M =mid, L =low. E =Exclusion, CA =Control area, CC =Control cage.
At Shingle Spit, there were three groupings amongst the communities
(Figure 3.8). The exclusion communities in the low and the mid zones were similar to the control cages and control areas of the low zone. The similarities amongst the exclusion areas indicates the homogenization of these communities.
The low zone at Shingle Spit had lower species richness than the mid zone
68 (Figure 3.6). There was a grouping of the mid and high control cages with the
mid control areas. The high exclusion community was similar to that of the high
control areas.
0.0
0.2
..... c .~ 0.4 IE Q) 8 .?:' 'C ro 'E 0.6 U5 I I I 0.8 I
1.0 LCC LE ME LCA MCA MCC HCC HE HCA Treatment Figure 3.8.Tree diagram illustrating the Bray-Curtis similarities for the Shingle Spit community at all tide heights under each treatment. H = high, M = mid, L = low. E = Exclusion, CA = Control area, CC = Control cage.
69 3.4.5 Control Cage Impacts
Due to the low number of replicates used to test the effects of the cage structure on the parameters tested in this experiment, it was hard to determine whether the cages had a significant effect. Cage structures had impacts to grain size profiles which could be due to the way the cages were prepared for exclusion (Table 3.3). There also was a marginal impact of the cage structure on phosphate concentrations (Table 3.4). In most cases it appears that the control cages showed similar results to the control area implying that the cage structure did not have a large impact on the chemical, physical and biological measurements taken throughout the course of this research.
3.5 Discussion
The goal of the work presented was to determine the ecological role of E. lewisii in terms of how it influences the functioning of the intertidal ecosystem.
This role is especially important to determine in light of the fact that E. lewisii is being treated as a pest species and manually removed from intertidal shellfish lease areas, a management strategy that does not take into account any function this species may have in the ecosystem. More studies like this one are needed to understand the ecological function of the species that we are busy eradicating.
The exclusion of E. lewisii had a significant impact on the penetrability of the sediment. The activities of bioturbators break up the surface of and displace sediment, creating interstitial spaces which are available for water and make sediment less compact (Katrak & Bird 2003; Lohrer et al 2004; Volkenborn et a!.
2007). Larger organisms, such as E. lewisii, are particularly important for their 70 role in the redistribution of sediments (Snelgrove 1999). The burrowing crab,
Chasmagnathus granulatus (Escapa et al. 2004; Palomo et al. 2003), the
lugworm, Arenicola marina (Volkenborn et al. 2007), the ghost shrimp Trypaea
australiensis, and the semaphore crab Heloecious cordiformis (Katrak & Bird
2003) have all been shown to decrease sediment softness and increase the
water content of the sediment through their bioturbatory activities.
Sediment penetrability and water content have significant consequences
to infaunal organisms. Larger infaunal bioturbators and those that build deeper
burrows within the sediment, extend the available habitat for other infaunal
organisms by creating interstitial spaces thereby increasing the depth to which
water, nutrients and oxygen penetrate the sediments (Escapa et al. 2004; Katrak
& Bird 2003; Snelgrove 1999). This would reduce competition for space, oxygen
and nutrients in areas of high infaunal density (Widdicombe & Austen 1998). Low
densities of T. australiensis created oxidizing conditions within the sediments
(Katrak & Bird 2003). Oxygen within the sediment decreases sulphide
concentrations and can benefit infaunal organisms with low sulphide tolerances
(Morrisey et al. 1999; Volkenborn & Reise 2007). Maintaining permeable
sediments is essential in locations, such as Fillongley, where the sediment is on
average finer grained and interstitial spaces are smaller in fine-grained sediment
and more susceptible to clogging (Volkenborn et al. 2007).
Permeability is directly related to grain size (Katrak & Bird 2003;
Volkenborn et al. 2007). Sediments become less permeable in areas where there
are many fine-grained and silt particles as these clog the interstitial spaces
71 (Volkenborn et al. 2007). We did not see the predicted significant trends in the grain size data due to the heterogeneous nature of the sediments at the study sites. This prediction was based on previous work that showed that the burrowing activities of the amphipod Corophium vo/utator (Biles et al. 2002), T. australiensis
(Katrak & Bird 2003), and A. marina (Volkenborn et al. 2007) caused the resuspension of fine grained sediments, which altered sediment grain profiles.
For future work, smaller and more homogeneous areas of the intertidal should be selected and a larger sampling size should be used to better determine the impacts of E. lewisii on grain size. The preparation of the study areas could also have led to the lack of significance through the action of sieving the sediments facilitating the removal of finer grained sediment through tidal action.
Increased permeability increases pore-water nutrient exchanges (Lohrer et a!. 2004; Volkenborn et al. 2007). Bioturbation is important in intertidal regions as it increases the depth to which chemicals and nutrients penetrate the substratum (Volkenborn et al. 2007). When lugworms are excluded, organic matter accumulated at the surface of the sediment (Volkenborn et al. 2007).
Bartoli et al.'s (2001) study in Italy demonstrates that accumulated organic carbon under shellfish aquaculture netting led to anoxic conditions and bivalve mortality. Volkenborn et al. (2007) found the decrease in pore-water spaces resulting from the exclusion of the bioturbator A. marina led to accumulations of several nutrients in the sediment. The burrowing urchin Echinocardium and
Austrovenus stutchburyi, a bivalve that actively ploughs across the surface of the sediment, both had significant effects on the release of NH4-N from the
72 sediments (Lohrer et al. 2004; Thrush et al. 2006). The experimental design used in this study may have prevented the detection of significant effects of E. lewisii on nutrient cycling therefore further work in this area is recommended.
The exclusion of large organisms including E. lewisii led to the homogenization of intertidal biological communities. The mid and low zone communities were similar at Shingle Spit. The communities in the high exclusions at both sites and the Fillongley mid exclusion were also found to be very similar.
This is problematic because the high communities are those with the lowest species richness and diversity. Through its influences on the physical properties of the sediment E. lewisii may have a positive impact of other infaunal species.
Burrowing to a depth of approximately 20 cm, E. lewisii would supply organisms living at this depth with water containing nutrients and oxygen (Bernard 1967).
Bioturbators create favourable conditions for other organisms through increased oxygenation, increasing the available habitat, which supports higher infaunal densities and allows them to live deeper in the sediment which leads to competitive release and protection from predators (Escapa et a1.2004; Palomo et al. 2003; Widdicombe & Austen 1998). Through these and other processes, bioturbation enhances diversity (Snelgrove 1999).
This study stresses the importance of understanding the role of an organism before management strategies are carried out. Infaunal organisms are very important as they are responsible for their habitat's structure and have crucial roles in many population, community and ecosystem processes (Thrush &
Dayton 2002). The preliminary findings here on sediment compaction and the
73 biological communities imply that this species may be an ecosystem engineer because it modifies the physical properties of the sediment and facilitates the survival of other organisms (Coleman & Williams 2002). Removing ecosystem engineers can be especially detrimental as they are responsible for ecosystem function and biological diversity (Coleman & Williams 2002; Volkenborn et al.
2007). In some marine systems, key species have been linked to a single role in terms of ecosystem function so losing it can have devastating effects (Lohrer et al. 2004). Increasing numbers of studies are showing the importance of each species in a community and the link between species richness and ecosystem function (Duarte 2000). It is important that it not be assumed that functionally similar organisms such as predators all have identical functions in the community as each species can have an individual function (Chalcraft & Resetarits 2003).
With further work on E. lewisii it is very possible that this species will be linked to nutrient cycling and ecosystem functioning. Not enough is known to determine which species are critical so we should consider all species important (Snelgrove
1999). Species loss and even simply density changes, through activities such as
E. lewisii removal, can lead to losses in biodiversity, resilience or provision of ecosystem services (Thrush & Dayton 2002).
Acknowledgments
We are extremely grateful to 1. L'Esperance for all her assistance in the field and support in the lab. Much appreciation goes to J. Whiteley and M. Hart for their constructive criticisms and guidance on various aspects of this research.
Thanks also to C. Allen, C. Kowalchuk, B. L'Esperance, and W. Kowalchuk for
74 their support and assistance. Funding for this work was provided by an NSERC strategic grant to L. Bendell-Young.
3.6 Literature Cited
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Bartoli M, Nizzoli D, Viaroli P, Turolla E, Castaldelli G, Fano EA, Rossi R. 2001. Impact of Tapes philippinarum farming on nutrient dynamics and benthic respiration in the Sacca di Goro. Hydrobiologia 455:203-212.
Beal BF, Parker MR, Veneile KW. 2001. Seasonal effects of intraspecific density and predator exclusion along a shore-level gradient on survival and growth of juveniles of the soft-shell clam, Mya arenaria L., in Maine, USA. Journal of Experimental Marine Biology and Ecology 264(2): 133-69.
Bernard FR. 1967. Studies on the biology of the naticid clam drill Polinices lewisii (Gould) (Gastropoda Prosobranchia). Fisheries Research Board of Canada Technical Report 42:1-41.
Biles Cl, Paterson DM, Ford RB, Solan M, Raffaelli DG. 2002. Bioturbation, ecosystem functioning and community structure. Hydrology and Earth System Sciences 6(6):999-1005.
Chalcraft DR and Resetarits WJ,Jr. 2003. Predator identity and ecological impacts: Functional redundancy or functional diversity? Ecology 84(9):2407 18.
Coleman FC and Williams SL. 2002. Overexploiting marine ecosystem engineers: Potential consequences for biodiversity. Trends in Ecology & Evolution 17(1 ):40-4.
Cook N. 2008. Feeding ecology and bioturbation: determining the ecological role of Euspira lewisii. MSc Thesis, Simon Fraser University, Burnaby, B.C.
Duarte CM. 2000. Marine biodiversity and ecosystem services: An elusive link. Journal of Experimental Marine Biology and Ecology 250(1-2): 117-31.
75 Escapa M, Iribarne 0, Navarro D. 2004. Effects of the intertidal burrowing crab Chasmagnathus granulatus on infaunal zonation patterns, tidal behavior, and risk of mortality. Estuaries 27(1 ):120-31.
Gee JM, Warwick RM, Davey ..IT, George CL. 1985. Field experiments on the role of epibent~lic predators in determining prey densities in an estuarine mudflat. Estuarine, Coastal and Shelf Science 21 (3):429-48.
Greenberg AE. 1992. Vanadomolybdophosphoric acid colorimetric method. In: Standard methods for the examination of water and wastewater. 18th ed. Washington, D.C.: The American Public Health Association. 4-112-4-113 p.
Katrak G and Bird FL. 2003. Comparative effects of the large bioturbators, Trypaea australiensis and Heloecius cordiformis, on intertidal sediments of Western Port, Victoria, Australia. Marine and Freshwater Research 54(6):701-8.
Lohrer AM, Thrush SF, Gibbs MM. 2004. Bioturbators enhance ecosystem function through complex biogeochemical interactions. Nature (London) 431 (7012): 1092-5.
Morrisey OJ, DeWitt TH, Roper OS, Williamson RB. 1999. Variation in the depth and morphology of burrows of the mud crab Helice crassa among different types of intertidal sediment in New Zealand. Marine Ecology Progress Series 182:231-242.
Page AL, editor. 1982. Ammonium by colormetric methods. In: Methods of Soil Analysis Part 2, 2nd ed. Madison, Wisconsin: American Society of Agronomy: Soil Science Society of America. 672-677 p.
Palomo G, Botto F, Navarro 0, Escapa M, Iribarne O. 2003. Does the presence of the SW Atlantic burrowing crab Chasmagnathus granulatus Dana affect predator-prey interactions between shorebirds and polychaetes? Journal of Experimental Marine Biology and Ecology 290(2):211-28.
Raghukumar Sand Anil AC. 2003. Marine biodiversity and ecosystem functioning: A perspective. Current Science 84(7):884-92.
Snelgrove PVR. 1999. Getting to the bottom of marine biodiversity: Sedimentary habitats - ocean bottoms are the most widespread habitat on earth and
76 support high biodiversity and key ecosystem services. Bioscience 49(2):129 38.
Thrush SF and Dayton PK. 2002. Disturbance to marine benthic habitats by trawling and dredging: Implications for marine biodiversity. Annual Review of Ecology and Systematics:449-73.
Thrush SF, Hewitt .JE, Gibbs M, Lundquist C, Norkko A. 2006. Functional role of large organisms in intertidal communities: Community effects and ecosystem function. Ecosystems 9(6): 1029-1 040.
Volkenborn N and Reise K. 2007. Effects of Arenicola marina on polychaete functional diversity revealed by large-scale experimentallugworm exclusion. Journal of Sea Research 57(1 ):78-88.
Volkenborn N, Hedtkamp SIC, Beusekom ...lEE, Reise K. 2007. Effects of bioturbation and bioirrigation by lugworms (Arenicola marina) on physical and chemical sediment properties and implications for intertidal habitat succession. Estuarine, Coastal and Shelf Science 74(1-2):331-343.
Whiteley ...1.2005. Macroinvertebrate community responses to clam aquaculture practices in British Columbia, Canada. MSc Thesis, Simon Fraser University, Burnaby, B.C.
Widdicombe S and Austen MC. 1998. Experimental evidence for the role of Brissopsis Iyrifera (Forbes, 1841) as a critical species in the maintenance of benthic diversity and the modification of sediment chemistry. Journal of Experimental Marine Biology and Ecology 228(2):241-255.
Wiltse WI. 1980. Effects of Polinices duplicatus (Gastropoda: Naticidae) on infaunal community structure at Barnstable Harbor, Massachusetts, USA. Marine Biology (Berlin) 56(4):301-10.
77 CHAPTER 4 CONCLUSIONS AND RECOMMENDATIONS
78 4.1 Conclusions
Euspira lewisii is a species of naticid snail native to the west coast of
North America. Due to its predatory nature and the way in which it attacks its
prey, this species has been presented as a hindrance to clam aquaculture in
British Columbia. In response to its reputation, E. lewisii is manually removed
from the intertidal by shellfish growers. The goals of this study were to fill the
gaps in knowledge about this species and gain understanding as to its feeding
ecology and functional role in intertidal ecosystems by mimicking the manual
removal through an exclusion experiment.
This work demonstrates that E. lewisii has distinct prey preferences on the
native littleneck clam, Protothaca staminea and avoids the commercially valuable
Manila clam, Venerupis philippinarum. This feeding pattern was observed in both
experiments and collected drilled shells and could be attributed to the
stratification of different bivalve species within the sediment. The feeding rate on
a variety of prey species was found to be only O.09c1ams/day or 1 c1am/14 days.
The feeding rate was species dependent and was slower for non-preferred prey types. This would be a maximal rate for E. lewisii as it was determined in the summer when feeding rates are highest. At the determined feeding rate, E. lewisii would have very low impacts on the bivalve communities at the study sites. Due to the high densities of clams in these areas and the slow feeding rate, only 3% of the bivalve community is consumed in one year and that is feeding at a constant, maximal feeding rate for 12 months. Realistically the impact would be even lower.
79 The exclusion experiment revealed that E. lewisii does playa role in the intertidal ecosystem. Sediment permeability decreased in the absence of this bioturbator as was expected. The effects on grain size and sediment chemistry were not statistically significant yet showed trends towards accumulations of several nutrients in exclusion areas. These alterations to the physical and chemical properties of the intertidal community impact the biological properties.
Several of the exclusion communities at various tide heights became very similar indicating a homogenization of the intertidal towards less diverse communities.
The results of the exclusion experiment demonstrate the importance of E. lewisii in this community and stress the importance of a full understanding of the role of each species in an ecosystem prior to carrying out control measures.
The results of the work presented here provide evidence of the limited impact of E. lewisii on bivalve communities as well as the importance of the ecological function of this species in the intertidal. These results are conclusive enough to advise the shellfish aquaculture industry that control measures of E. lewisii are not necessary. Even though the evidence is strong more work is needed to get a full understanding of the role of this species in the ecosystem.
4.2 Future Work
I recommend that more work be done on the bioturbation activities of E. lewisii. A better design for this experiment would be to compare enclosures to exclusion cages. If this design were to be used, larger cages or a fencing technique would be recommended allowing more mobility for the enclosed moon
80 snails. This would allow for better generalizations about the impacts this species
has on the physical, chemical and biological properties of the intertidal.
To understand the extent to which this species influences sediment grain
size profiles, I recommend that smaller, more homogeneous sections of the
intertidal be used to reduce the variability that comes from the patchy nature of
the sediments at the study sites. I would also recommend that a larger sample
size be used.
The trends towards nutrient accumulations under E. lewisii exclusion
conditions imply that with further work significant trends may be detected. It is
therefore recommended that bulk cores be taken at regular intervals to track the
changes of the nutrient concentrations over time. The time scale for the current
study may have too long and the nutrient concentrations would be more
influenced by daily tidal fluctuations than the presence or absence of E. lewisii.
Porewater peepers should also be used to determine the depth to which E.
lewisii influences nutrient fluxes from the sediment. An alternative to field studies would be to carry out measurements in lab mesocosms, enabling control over as many variables as possible and to use more accurate measurement techniques.
4.3 Recommendations
The results from this study can be used to advise the clam aquaculture industry. The feeding experiments demonstrate that E. lewisii is not detrimental to the industry in that they avoid Venerupis philippinarum, they feed at a low rate, and natural and aquaculture tenure densities the numbers of clams consumed
81 are extremely low. For these reasons it is no longer necessary to recommend or continue practicing the manual removal of moon snails from intertidal areas. It can also be stated that E. lewisii alters sediment properties and further work might show that this species is a benefit to infaunal organisms including those inhabiting shellfish leases.
82 APPENDICES
Appendix A: Exclusion Experiment By-Tide-Height Results
1.6 1.6 A B 1.4 1.4 N N E E u 1.2 u 1.2 OJ OJ --~ C ~ ..c 1.0 ..c...... 1.0 _ Exclusion OJ OJ C C c::::::::::J Control Area Q) Q) ...... Control Cage ti5 0.8 ti5 0.8 Q) Q) > ·w 0.6 ·w> 0.6 (/) (/) ....Q) ....Q) 0- 0- E 0.4 E 0.4 0 0 U U 02 0.2
0.0 0.0 High Mid Low High Mid Low Tide Height Tide Height Compressive strength of the sediment at Fillongley (A) and Shingle Spit (8) in each of the treatments at each tide height (Medians, error bars represent interquartile range).
83 20 20 A B ~ ~ ~ 0 '-' ~ C 15 C 15 Q) ...... Q) c c 0 0 0 0 ...... Exclusion ...... Q) ...... Q) c=::J Control Area 10 cu 10 - Control Cage ~ S Q) Q) OJ OJ ...... cu ...... cu c c Q) 5 Q) 5 u.... u.... Q) Q) n... n...
0 0 High Mid Low High Mid Low Tide Height Tide Height
Percentage of water in the sediment at Fillongley (A) and Shingle Spit (8) in each of the treatment conditions at each tide height (Medians, error bars represent interquartile range).
84 Summary of the non-parametric Kruskal-Wallis analyses on the physical properties of the sediments between treatments at each site and at each tide height. * indicates a significant result and ** indicates a marginally significant result.
Study Site Physical Tide Chi- p- Significant? Property Height squared value Value High 8.22 0.0164 * Compressive Mid 1.5 0.4724 Strength Low 4.05 0.1321 Fillongley High 4.32 0.115 Water Mid 1.95 0.3779 Content Low 9.44 0.0089 *
High 12.46 0.002 * Compressive Mid 10.74 0.0047 * Strength Shingle Low 2.54 0.2804 Spit High 0.91 0.6342 Water Mid 2.6 0.273 Content Low 1.77 0.4125
85 100 100 A B
80 80 ~ ~ 0 ~ ~ ill ill > 60 > 60 ....co ....co _ Exclusion l') l') (J) (J) c::=::J Control Area OJco OJco Control Cage C 40 c 40 (J) -(J) u.... u.... (J) (J) n... n... 20 20
0 0 High Mid Low High Mid Low Tide Height Tide Height
Percentage of gravel in the sediment at Fillongley (A) and Shingle Spit (B) in each of the treatment conditions at each tide height (Medians, error bars represent interquartile range).
60 60 A B ~ 50 ~ 50 -:!2.0 '-' ~ u u c c co co (f) 40 (fJ 40 (J) (J) _ Exclusion en en ...... c:::::=J Control Area CO CO 0 30 0 30 Control Cage 0 0 (J) (J) OJ OJ ...... co 20 ...... co 20 c c (J) (J) ....u ....u (J) (J) D... 10 D... 10
0 0 High Mid Low High Mid Low Tide Height Tide Height
Percentage of coarse sand in the sediment at Fillongley (A) and Shingle Spit (B) in each of the treatment conditions at each tide height (Medians, error bars represent interquartile range).
86 50 50 A B
---. ~ 40 ~ 40 ...... 0 0...... "0 "0 C C ro ro (J) 30 (J) 30 Q) Q) c c _ Exclusion u:: u:: c==:J Control Area Q) Q) 0) 0) Control Cage ro 20 ro 20 C c Q) -Q) u u '- '- Q) Q) 0- 10 0- 10 I~ a a High Mid Low High Mid Low
Tide Height Tide Height
Percentage of fine sand in the sediment at Fillongley (A) and Shingle Spit (B) in each of the treatment conditions at each tide height (Medians, error bars represent interquartile range).
3.0 3.0 A B 2.5 2.5
---. ~ ~ ~ 2.0 ...... 0 2.0 ~ ~ U5 U5 _ Exclusion Q) Q) c=:::J Control Area 0) 1.5 0) 1.5 ro ro Control Cage c c -Q) -Q) u u '- '- Q) 1.0 Q) 1.0 0- 0-
0.5 0.5
0.0 0.0 High Mid Low High Mid Low
Tide Height Tide Height
Percentage of silt in the sediment at Fillongley (A) and Shingle Spit (B) in each of the treatment conditions at each tide height (Medians, error bars represent interquartile range).
87 Summary of the non-parametric Kruskal-Wallis analyses on the grain size analyses between treatments at each site and at each tide height. * indicates a significant result and ** indicates a marginally significant result.
Study Site Physical Tide Chi- p- Significant? Property Height squared value Value High 3.64 0.1619 Percentage Mid 0.84 0.6564 of Gravel Low 5.83 0.0541
High 7.5 0.0235 Percentage of Coarse Mid 0.37 0.8294 Sand Low 4.34 0.1142 Fillongley High 1.31 0.5195 Percentage Mid 1.19 0.5515 of Fine Sand Low 8.26 0.0161 *
High 15.74 0.0004 * Percentage Mid 2.47 0.2904 of Silt Low 7.25 0.0267
High 4.66 0.0972 Percentage Mid 2.37 0.3061 of Gravel Low 1.25 0.534
High 6.48 0.0392 Percentage of Coarse Mid 1.36 0.5071 Sand Shingle Low 1.19 0.5527 Spit High 2.64 0.2667 Percentage Mid 5.98 0.0502 of Fine Sand Low 1.26 0.5317
High 1.46 0.4806 Percentage Mid 1.44 0.4855 of Silt Low 4.3 0.1163
88 0.012 0.012 A B ~ OJ OJ --OJ 0.010 --OJ 0.010 E -S c c a a 0008 ~ 0.008 .~ ...... c C _ Exclusion Ammonium concentrations at Fillongley (A) and Single Spit (8) in each treatment at each tide height (Medians, error bars represent interquartile range). 0.18 018 A B 0.16 016 OJ OJ 0> 014 --OJ 0.14 -S E c 0.12 -c 0.12 a a _ Exclusion ~ .~ ..... 0.10 ..... 0.10 c:==J Control Area C C Control Cage 0.00 0.00 High Mid Low High Mid Low Tide Height Tide Height Carbon concentrations at Fillongley (A) and Shingle Spit (8) in each treatment at each tide height (Medians, error bars represent interquartile range). 89 016 0.16 A B ~ ~ OJ 0.14 OJ 0.14 OJ OJ E E 0.12 012 -c -c 0 0 'iU 0.10 ~ 0.10 L.. L.. _ Exclusion C C Q) Q) c=:J Control Area u 0.08 u 0.08 c c Control Cage 0 0 0 0 0.06 0.06 Q) roQ) ro J:: J:: 0- 0.04 0- 0.04 (/J (/J 0 0 J:: J:: 0... 0.02 0... 0.02 0.00 0.00 High Mid Low High Mid Low Tide Height Tide Height Phosphorous concentrations at Fillongley (A) and Shingle Spit (8) in each treatment at each tide height (Medians, error bars represent interquartile range). 90 Summary of the non-parametric Kruskal-Wallis analyses on the sediment nutrient characteristics between treatments at each site and at each tide height. * indicates a significant result and ** indicates a marginally significant result. Study Site Nutrient Tide Chi- p- Significant? Height squared value Value High 7.49 0.0236 ** Ammonium Mid 3.45 0.1778 Low 0.26 0.8768 High 6.58 0.0371 Fillongley Carbon Mid 2.26 0.3223 Low 0.73 0.6932 High 6.73 0.0346 Phosphorous Mid 3.97 0.1369 Low 0.71 0.702 High 6.39 0.0409 Ammonium Mid 4.44 0.1084 Low 1.36 0.5056 High 2.2 0.3321 Shingle Spit Carbon Mid 0.64 0.724 Low 3.05 0.2176 High 7.18 0.0275 ** Phosphorous Mid 2.63 0.2683 Low 1.98 0.3705 91 0.0 0.2 +oJ C ·uQ) 0.4 li= Q) 8 ·e.z- co ·E 0.6 U5 I 0.8 t- 1 ~ ~ 1 1.0 SLC SLE SME SLA SMA SMC SHC SHA FHA FHC SHE FHE FME FLE FLC FLA FMA FMC Treatment Tree diagram illustrating the Bray-Curtis similarities for the communities under all treatment at all tide heights at both sites. F =Fillongley, S =Shingle Spit. H = high, M =mid, L =low. E =Exclusion, A =Control area, C =Control cage. 92 Appendix B: Exclusion Experiment Supplementary Data Total invertebrate abundance means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion 756±354 185±116 Overall Control Area 342±110 172±57 Control Cage 478±302 99±47 Exclusion 651±518 372±358 High Control Area 281±128 259±84 Control Cage 368±1340 114±616 Exclusion 692±755 103±54 Mid Control Area 450±136 175±117 Control Cage 414±2255 106±540 Exclusion 924±1378 81±33 Low Control Area 296±429 82±25 Control Cage 823±O 77±25 93 Bivalve abundance means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion 453±260 180±117 Overall Control Area 153±114 169±57 Control Cage 203±218 95±49 Exclusion 610±519 370±358 High Control Area 278±126 258±84 Control Cage 354±1455 112±642 Exclusion 664±770 94±44 Mid Control Area 153±421 171±118 Control Cage 140±1010 104±546 Exclusion 85±79 76±35 Low Control Area 28±17 80±25 Control Cage 27±O 70±6 94 Non-prey abundance means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion 302±376 5±3 Overall Control Area 190±140 3±1 Control Cage 275±456 4±4 Exclusion 40±56 2±1 High Control Area 3±4 2±2 Control Cage 14±114 2±25 Exclusion 28±32 9±12 Mid Control Area 298±300 4±4 Control Cage 275±3265 2±6 Exclusion 838±1341 6±4 Low Control Area 268±426 3±2 Control Cage 796±O 8±19 95 Total invertebrate species richness means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion 8.3±1.8 6.1±1.3 Overall Control Area 7.0±1.4 5.8±1.2 Control Cage 7.6±3.0 5.5±1.4 Exclusion 5.8±2.0 4.8±2.0 High Control Area 5.0±1.8 5.0±2.2 Control Cage 7.0±25.4 4.5±6.4 Exclusion 8.2±4.6 7.2±3.5 Mid Control Area 7.0±1.3 7.0±4.1 Control Cage 7.5±31.8 5.0±O.O Exclusion 11.0±1.8 6.2±3.3 Low Control Area 9.0±3.4 5.2±2.0 Control Cage 9.0±O.O 7.0±12.7 96 Bivalve species richness means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion 4.7±1.1 3.4±O.7 Overall Control Area 3.9±O.9 4.1±O.8 Control Cage 3.6±1.4 3.5±O.9 Exclusion 3.0±1.3 3.5±1.6 High Control Area 3.5±1.6 4.0±1.3 Control Cage 3.5±6.4 3.0±12.7 Exclusion 4.8±2.7 3.8±1.5 Mid Control Area 3.5±1.6 5.0±2.2 Control Cage 3.5±19.0 3.5±6.4 Exclusion 6.2±O.8 3.0±2.2 Low Control Area 4.8±3.0 3.2±2.0 Control Cage 4.0±O.O 4.0±O.O 97 Non-prey species richness means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion 3.7±O.8 2.7±1.2 Overall Control Area 3.1±O.9 1.7±O.6 Control Cage 4.0±1.8 2.0±1.5 Exclusion 2.8±O.8 1.2±O.8 High Control Area 1.5±1.6 1.0±1.3 Control Cage 3.5±19.0 1.5±19.0 Exclusion 3.5±2.0 3.5±4.6 Mid Control Area 3.5±O.9 2.0±2.2 Control Cage 4.0±12.7 1.5±6.4 Exclusion 4.8±1.5 3.2±2.0 Low Control Area 4.2±O.8 2.0±O.O Control Cage 5.0±O.O 3.0±12.7 98 Total invertebrate species evenness means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion OA5±O.13 O.39±O.10 Overall Control Area O.34±O.10 OA8±O.13 Control Cage O.33±O.22 O.58±O.16 Exclusion O.58±O.25 O.25±O.11 High Control Area OA4±O.19 O.51±O.31 Control Cage OA8±1A5 O.62±O.O4 Exclusion OA3±O.25 O.51±O.23 Mid Control Area O.24±O.23 O.57±O.26 Control Cage O.28±O.17 O.71±O.O4 Exclusion O.33±OAO OA2±O.26 Low Control Area O.35±O.28 O.34±O.22 Control Cage O.12±O.OO OAO±O.64 99 Bivalve species evenness means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site I Shingle Tide Heights Treatment Fillongley Spit Exclusion 0.47±0.15 0.43±0.10 Overall Control Area 0.36±0.13 0.52±0.23 Control Cage 0.38±0.28 0.64±0.27 Exclusion 0.69±0.17 0.28±0.14 High Control Area 0.53±0.24 0.57±0.32 Control Cage 0.60±0.96 0.74±1.29 Exclusion 0.45±0.22 0.54±0.28 Mid Control Area 0.26±0.002 0.62±0.11 Control Cage 0.28±0.003 0.85±0.44 Exclusion 0.26±0.32 0.46±0.22 Low Control Area 0.28±0.00 0.37±0.07 Control Cage 0.11±0.00 0.32±0.28 100 Non-prey species evenness means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion O.20±O.O9 O.19±O.O7 Overall Control Area O.15±O.O6 O.12±O.18 Control Cage O.13±O.11 O.26±O.22 Exclusion O.24±O.26 O.12±O.10 High Control Area O.10±O.O3 O.OO±O.91 Control Cage O.14±1.29 O.29±3.65 Exclusion O.16±O.18 O.23±O.17 Mid Control Area O.11±O.O6 O.16±O.13 Control Cage O.16±O.24 O.20±O.52 Exclusion O.21±O.26 O.22±O.25 Low Control Area O.23±O.OO O.19±O.O1 Control Cage O.O3±O.OO O.30±O.O5 101 Total invertebrate Shannon-Wiener index means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site S~lingle Tide Heights Treatment Fillongley Spit Exclusion O.90±O.25 O.72±O.22 Overall Control Area O.64±O.19 O.84±O.22 Control Cage O.65±O.40 O.95±O.19 Exclusion 1.01±O.43 O.38±O.17 High Control Area O.71±O.30 O.83±O.50 Control Cage O.93±2.82 O.93±O.O6 Exclusion O.91±O.52 1.00±O.45 Mid Control Area O.46±O.45 1.11±O.50 Control Cage O.57±O.35 1.14±O.O7 Exclusion O.79±O.95 O.76±O.48 Low Control Area O.76±O.63 O.57±O.37 Control Cage O.26±O.OO O.77±1.25 102 Bivalve Shannon-Wiener index means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion O.65±O.17 O.52±O.14 Overall Control Area OA8±O.16 O.74±O.22 Control Cage OA7±O.34 O.78±O.30 Exclusion O.76±O.19 O.36±O.17 High Control Area O.67±O.30 O.8O±OA9 Control Cage O.75±1.21 O.82±1A2 Exclusion O.71±O.34 O.71±O.36 Mid Control Area O.33±OA4 1.00±OAO Control Cage O.36±O.O1 1.06±O.55 Exclusion OA7±O.59 O.50±O.24 Low Control Area OA3±O.36 OA3±O.34 Control Cage O.15±O.OO OA5±O.39 103 Non-prey Shannon-Wiener index means/m2±95% confidence interval for each treatment over the entire study area and each tide height at both sites. Study Site Shingle Tide Heights Treatment Fillongley Spit Exclusion O.26±O.12 O.19±O.11 Overall Control Area O.17±O.10 O.O9±O.O4 Control Cage O.18±O.13 O.18±O.16 Exclusion O.25±O.26 O.O3±O.O2 High Control Area O.O4±O.O1 O.O3±O.O4 Control Cage O.18±1.62 O.12±1A8 Exclusion O.20±O.22 O.29±O.22 Mid Control Area O.13±O.O6 O.11±O.11 Control Cage O.22±O.34 O.O8±O.62 Exclusion O.32±OAO O.26±O.27 Low Control Area O.33±O.28 O.13±O.O7 Control Cage O.12±O.OO O.32±O.86 104