UBC Theses and Dissertations

UBC Theses Logo

UBC Theses and Dissertations

Mechanisms underlying marine macroalgal invasions : understanding invasion success of Sargassum muticum White, Laura Linsey Fallaize 2010

Your browser doesn't seem to have a PDF viewer, please download the PDF to view this item.

Notice for Google Chrome users:
If you are having trouble viewing or searching the PDF with Google Chrome, please download it here instead.

Item Metadata

Download

Media
24-ubc_2010_spring_white_laura.pdf [ 1.4MB ]
Metadata
JSON: 24-1.0069842.json
JSON-LD: 24-1.0069842-ld.json
RDF/XML (Pretty): 24-1.0069842-rdf.xml
RDF/JSON: 24-1.0069842-rdf.json
Turtle: 24-1.0069842-turtle.txt
N-Triples: 24-1.0069842-rdf-ntriples.txt
Original Record: 24-1.0069842-source.json
Full Text
24-1.0069842-fulltext.txt
Citation
24-1.0069842.ris

Full Text

MECHANISMS UNDERLYING MARINE MACROALGAL INVASIONS: UNDERSTANDING INVASION SUCCESS OF SARGASSUM MUTICUM  by  LAURA LINSEY FALLAIZE WHITE B.Sc. (Hons) The University of Portsmouth, 2003  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF  DOCTOR OF PHILOSOPHY in THE FACULTY OF GRADUATE STUDIES (Zoology)  THE UNIVERSITY OF BRITISH COLUMBIA (Vancouver) May 2010 © Laura Linsey Fallaize White, 2010  ii  ABSTRACT  This thesis examines different aspects of invasion success of the brown macroalga Sargassum muticum. Chapter two investigates the relationship between native diversity and invasibility by S. muticum in experimental and unmanipulated communities of low intertidal macroalgae. I found that diversity effects on invasion varied from positive to negative with life history stage of the invader. Native diversity facilitated recruitment of S. muticum, but decreased growth and or survivorship. Phenological differences between S. muticum and native macroalgal species may contribute to the success of this invader in British Columbia. Chapter three explores the effects of S. muticum on native macroalgal diversity at different densities by manipulating exotic density in natural communities. I found that the effects of S. muticum on native macroalgal richness were both density and time dependent, and are mediated through competition for light. The reciprocal interaction between S. muticum and native macroalgal diversity has shown effects in both directions, and suggest some degree of symmetry in the interaction between non-native S. muticum and native macroalgae. Chapter four examines whether non-native S. muticum is less grazed than native macroalgae in British Columbia, and whether the concentrations of defensive polyphenolic compounds in the tissue of the exotic differ from native conspecifics. In contrast to the predictions of the Enemy-Release Hypothesis, I showed that when presented a choice, native herbivores do not discriminate between native and non-native macroalgae. The levels of polyphenolic defenses in the exotic were similar to some native  iii  macroalgae, suggesting differences in polyphenolic concentrations are not influencing herbivore choice. Reduced grazing of non-native S. muticum by native herbivores is not contributing to the success of this invasive in British Columbia. Chapter five tests two predictions of the Evolution of Increased Competitive Ability hypothesis; whether S. muticum in non-native regions attains greater sizes and lower chemical defenses than conspecifics in the native region. We compared the size (as a measure of performance) and levels of polyphenolic defenses of S. muticum from its native and invaded regions. My preliminary results suggest that in non-native regions, S. muticum attains larger sizes with lower levels of defensive polyphenolic compounds than native regions.  iv  TABLE OF CONTENTS  ABSTRACT…………………...…………………………………………………….........ii TABLE OF CONTENTS……………………………………………………………….iv LIST OF TABLES…………………………………………………………………..…..vi LIST OF FIGURES…………………………………………………………………….vii ACKNOWLEDGEMENTS………………………………………………………...…viii CO-AUTHORSHIP STATEMENT……………………………………………………ix CHAPTER ONE: General Introduction…………………..…………………...……....1 Literature Cited…………………………………………………………………...…...7 CHAPTER TWO: Diversity effects on invasion vary with life history stage in marine macroalgae……………………………………………………………………...11 Introduction…………………………………………………………………………..11 Materials and Methods……………………………………………………………….13 Results………………………………………………………………………………..19 Discussion……………………………………………………………………………24 Literature Cited………………………………………………………………………34 CHAPTER THREE: Density dependent effects of an introduced macroalga on native community diversity…….………………………………………………………37 Introduction…………………………………………………………………………..37 Materials and Methods……………………………………………………………….40 Results………………………………………………………………………………..46 Discussion……………………………………………………………………………54 Literature Cited………………………………………………………………………62 CHAPTER FOUR: Comparing herbivory and secondary compounds in invasive Sargassum muticum with native brown algae of British Columbia………………….67 Introduction…………………………………………………………………………..67 Materials and Methods……………………………………………………………….71 Results………………………………………………………………………………..78 Discussion……………………………………………………………………………83 Literature Cited………………………………………………………………………90  v  CHAPTER FIVE: Biogeographic patterns of size and chemical defense of an invasive marine macroalga…...………………………………………………………...97 Introduction…………………………………………………………………………..97 Materials and Methods……………………………………………………………...103 Results………………………………………………………………………………107 Discussion…………………………………………………………………………..109 Literature Cited……………………………………………………………………..114 CHAPTER SIX: General Discussion and Conclusions………………….………… 121 Literature Cited…………...………………………………………………………...128  vi  LIST OF TABLES Table 2.1 Ten compositional combinations (A-J) of native species over four diversity levels (0, 1, 3, 4) used for constructed plots………………………...………....17 Table 2.2 Repeated-measures ANOVA test criteria for (A) exotic invasion, (B) native recruitment, and (C) availability of primary space……………………………….24 Table 3.1 List of species sampled during study, with resource-based functional group assignment and code……………………………………………………………..……....43 Table 3.2 Repeated-measures ANOVA test criterion for the effects of Sargassum muticum on (A) native species richness, (B) native diversity, (C) native cover, and (D) canopy (E) understory, and (F) basal functional group cover……………………………49 Table 4.1 Treatment codes for macroalgal combinations used in grazing trials………...75 Table 4.2 ANOVA test criterion for (A) no-choice, and (B) choice grazing trials, and (C) polyphenolic content……………………………………………………………………..82 Table 5.1 Collection sites of Sargassum muticum individuals ………..………………103  vii  LIST OF FIGURES Figure 2.1 Exotic Sargassum muticum cover (A), and available primary space (B) plotted against native richness, for observational field data……………………………………..15 Figure 2.2 Time series of percent (A) cover of exotic Sargassum muticum germlings, (B) native cover (canopy and basal cover combined) and (C) native recruitment, per diversity level………………………………………………………………………...23 Figure 2.3 Diversity level plotted against (A) initial Sargassum muticum cover, (B) final S. muticum cover, and (C) change in cover of S. muticum from day 47 to day 77………26 Figure 2.4 Bar charts of (A, B) cover of Sargassum muticum germlings, (C, D) canopy cover, (E, F) basal cover, and (G, H) availability of primary space……………………..28 Figure 3.1 Effects of Sargassum muticum density treatments on native species richness through light and space competition……………………………………………………..48 Figure 3.2 Relationship between Sargassum muticum density and total native macroalgal cover (%) for each month (May-August) for the 2006 and 2007 growing seasons……...51 Figure 3.3 Effects of the exotic density treatments on the average cover of the 19 native species most strongly affected by Sargassum muticum, and cover of the resource-based functional groupings……………………………………………………………………..52 Figure 3.4 Relationship between Sargassum muticum density and native and exotic (S. muticum) macroalgal biomass (g dried mass)………………………………………..54 Figure 4.1 Percent macroalgae consumed (± 1SE), when presented individually (nochoice trials) to native herbivores (A-G) and average effect of all herbivores (H)……...80 Figure 4.2 Percent macroalgae consumed (± 1SE), when presented in choice trials to native herbivores…………………………………………................................................81 Figure 4.3 Average polyphenolic content (% DW) of non-native Sargassum muticum and three native macroalgae from three sites in Barkley Sound……...………………………82 Figure 5.1 Map of collection regions of Sargassum muticum………………………....104 Figure 5.2 Mean length of Sargassum muticum individuals from the native region (Japan) and invaded regions of Western Canada……………………………………….108 Figure 5.3 Mean polyphenolic content (% DW) of Sargassum muticum individuals from distinct populations within different geographic regions……………………………….109  viii  ACKNOWLEDGEMENTS I am indebted to many people who have supported and advised me through this thesis. First I thank my supervisor, Dr. Jonathan Shurin for his guidance, helping me develop research skills, and patiently reading through numerous manuscript revisions. I also thank my committee members, Dr. Rob DeWreede, Dr. Christopher Harley, Dr. Colin Levings, and Dr. Mark Vellend, who all contributed to the design and development of this thesis and whose generosity with their time is greatly appreciated. I would like to thank all of the people who have assisted me in the laboratory and field. Primarily two research assistants, Pylin Chuapetcharasopon and Mikaela Davis, who were a great help to my field research. Thanks to Roger Haslam, Jung Hyun Oak, Takeaki Hanyuda, and Eric Henry for overseas field collections. Thanks also to Sarah Nienhuis for long hours in the laboratory, and to Lyanne Burgoyne who was always willing to lend a hand. Also Stefan Dick, Amrit van der Lely, Russel Markel, and Colin Bates. I thank the Shurin lab for feedback on experimental analysis and manuscript revisions. I also thank the staff of Bamfield Marine Sciences Centre who provided logistical support. This thesis would not have been possible without the support and encouragement of Clayton McIntosh, who supported me in the most important ways, and the rest of my family, Peter, Sue, Heidi, Brenda, and Cassy. Finally, I thank Dr. William Farnham, who encouraged a young phycologist and showed me the way. Financial support was provided primarily by the States of Guernsey, and supplemented by various research grants: the John Boom Memorial Scholarship, Bamfield Marine Sciences Centre (2005, 2008), Western Canadian Universities Marine Science Research Award (2006, 2008), Phycological Society of America, Grant-In-Aid of Research (2007), and Pacific Northwest Shell Club Research Scholarship (2008).  I dedicate this thesis to my friend William J. Mather, who I will miss always.  ix  CO-AUTHORSHIP STATEMENT Chapters two and three were co-authored in published form with Jonathan Shurin. Dr. Shurin and I both contributed to the research design in these chapters. I performed all field work and data analysis. Manuscripts were prepared by myself and edited by Jonathan Shurin and members of my supervisory committee. Chapters four and five were co-authored with Sarah B. Nienhuis. Sarah and I both contributed to the research design. I performed all the field work and sample collection. We both contributed to laboratory work. I analyzed the data, and wrote the manuscript, with edits from Sarah. I acknowledge the contributions of Jonathan Shurin and Sarah Nienhuis by using the plural pronoun “we” throughout these chapters.  1  CHAPTER ONE GENERAL INTRODUCTION  The invasion of habitats by non-native species is a global phenomenon with serious consequences for ecological, economic, and social systems (Pimentel et al. 2000). Species invasions are considered one of the greatest threats to native biodiversity and resource values of the world’s oceans (Carlton 2000). In recent years, anthropogenic influences on previously existing barriers to dispersal have greatly increased the transfer of non-native marine organisms out of their natural ranges to become invasive elsewhere. Warming waters globally are further promoting the colonization of novel species in new environments. With increasing globalization and climate change contributing to the introduction of non-native species into new habitats (Galil et al. 2007), the threat of invasive species is growing. This is evidenced by the growing cost both economically and ecologically invasive species are having globally (Pimentel et al. 2005). The assessment of ecological impacts of non-native marine species has been recognized as a research priority in recent years, following the realization that large marine ecosystems are losing their biological distinctiveness (Mollo et al. 2008). Less is known about invasive species and biological invasions in the marine then in the terrestrial environment, particularly for macroalgae (Ribera Siguan 2005). Macroalga can be invaders with serious impacts, and have become a prominent component of the marine flora in many regions worldwide. The macroalgal species Caulerpa taxifolia and Undaria pinnatifida are ranked as two of the World’s 100 most invasive species  2  (www.issg.org). C. taxifolia has smothered more than 13,000 hectares of seafloor in the Mediterranean Sea (Meinesz et al. 2001), and U. pinnatifida outcompetes native macroalgae in New Zealand (Battershill et al. 1998) and Argentina (Casas et al. 2004), but causes it biggest impacts through fouling marine farms (NIMPIS 2002). Impacts of macroalgal invaders are typically expressed as community dominance through the monopolization of space, and changing competitive relationships in the native assemblage (reviewed by Schaffelke and Hewitt 2007). Because of their worldwide distribution and capacity to alter native communities, introduced macroalgae, in synergy with other anthropogenic disturbances such as climate change and coastal pollution, are potentially important agents of global ecological change (Britton-Simmons 2004). Despite this, few studies have rigorously examined the mechanisms that underpin successful macroalgal invasions (Smith et al. 2002). As a consequence many of their possible impacts remain speculative (reviewed by Schaffelke and Hewitt 2007). Information that is available is reactive, following discoveries of introductions or dominated by case studies that are strongly idiographic, focusing on high profile taxa that have, or might have, large ecological or economic impacts (Schaffelke et al. 2006, Johnson and Chapman 2007). The need for quantitative assessments of the impacts of invasive marine macroalgae has been recognized. The few studies that have demonstrated strong effects of exotic macroalgal species (Verlaque 1994b, Villele and Verlaque 1995, Walker and Kendrick 1998, Levin et al. 2002, Britton-Simmons 2004) highlight the importance of future research into the spread and impacts of introduced macroalgae. Further, understanding the mechanisms that contribute to invasion success is central to an  3  appreciation of the factors promoting marine invasions and to define control options. This is perhaps more important for macroalgae than other taxonomic groups of marine aliens as once invasive macroalgae are established, eradication is highly unlikely (i.e. Caulerpa taxifolia in the Mediterranean, Codium fragile ssp. tomentosoides in Atlantic Canada, Undaria pinnatifida in Australia and New Zealand). Scientists and policy makers increasingly recognize the introduction of non-native species as a major threat to marine biodiversity and a contributor to environmental change (Bax et al. 2003). The failure to control the threat of invasions generally has been attributed to insufficient policy, inadequate research and management funding, and gaps in scientific knowledge (Simberloff et al. 2005). Legislation pertaining to macroalgal invaders has emerged in some countries in response to prolific, local ecological and economic impacts. Australia, New Zealand, the US, Canada, Switzerland, and Germany have legislation controlling introductions for aquaculture (Johnson and Chapman 2007), but only Australia has implemented hull fouling measures, a more important introduction vector for macroalgae. In western Canada, there is an apparent paucity of legislation preventing the importation and movement of, or response to, macroalgal invasives. Nonnative macroalgal species are frequently imported by the aquaria trade as decorative plants in tanks, as cultural foods, and unintentionally attached to ship hulls, gear, and in ballast water. The marine macroalga Sargassum muticum (Yendo) Fensholt (Phaeophyceae: Fucales) is native to Japan (Yoshida 1983) where it reportedly comprises a relatively minor component of the native macroalgal flora (Norton 1977b, Critchley 1983). This  4  species has been introduced to the coastlines of several countries in the northern hemisphere in the past 60 years through the translocation of Japanese Oysters (Crassostrea gigas) for aquaculture (Scagel 1956). Sargassum muticum possesses many of the intrinsic traits of an invasive species, including very high growth rates of 2-4 cm per day (Critchley 1981, Lewey and Farnham 1981), high fecundity, monoecious individuals with a perennial life history (Norton and Deysher 1989), and multiple-range dispersal mechanisms including germling settlement and drifting fertile thalli (Norton 1976). Sargassum muticum is believed to have been introduced to British Columbia, Canada, in the 1940s (possibly as early as 1902, Scagel 1956). Now well established on sheltered shores of the Pacific Northwest, S. muticum is the dominant macroalga in many low intertidal and sub tidal communities, displacing some native taxa through competition; Lithothrix aspergillum (DeWreede and Vandermeulen 1988), Laminaria spp., Cystoseira spp., Scytosiphon lomentaria, Gracilaria verrucosa, bull kelp Nereocystis luetkeana (Thom and Hallum 1990), the giant kelp Macrocystis integrifolia (Ribera and Boudouresque 1995), and Saccharina subsimplex (Britton-Simmons 2004). Sargassum muticum is an extremely successful invasive species, with negative impacts on native ecology reported from throughout its introduced range (Harris et al. 2007), but little is known of the mechanisms behind its success (but see Andrew and Viejo 1998, Britton-Simmons 2004, Britton-Simmons and Abbott 2008). This thesis investigates invasion success of S. muticum in British Columbia, Canada. Invasion of macroalgae is a complex, multi-stage process that can be arbitrarily, but usefully broken down into four distinct phases; arrival, establishment, persistence,  5  and spread (Mollison 1986). Once introduced to a new region with tolerable physical characteristics, exotic species have to establish in a new community. The first two chapters of this thesis center on the interaction between native diversity and the invasive macroalga, employing empirical field experiments and surveys of unmanipulated low intertidal macroalgal communities. First, I investigate the arrival and establishment phases of invasion by examining biotic resistance to invasion. Specifically, the role of native community diversity in governing invasion success of S. muticum (chapter two). A well established theory in ecology is that diverse communities are more resistant to invasion by non-native species (the diversity-invasion hypothesis, Elton 1958), and is based on the idea that more diverse communities show more complete or efficient use of resources, leaving less available for new species to occupy. This theory is well studied in terrestrial plants (Tilman 1997, Lyons and Schwartz 2001, Kennedy et al. 2002), but has not been investigated in macroalgae. Next I examine the persistence and spread phase of invasion through an investigation of the impacts of S. muticum on native macroalgal communities (chapter three). Whether invasions cause net increases or decreases in local or regional species biodiversity has raised much controversy, and the few studies from the marine environment have reported a range of ecological interactions (Wasson et al. 2005, Klein et al. 2005, Mineur et al. 2008). Studies identifying the mechanisms underlying the impacts of macroalgal aliens are limited. I investigate the effects of S. muticum at different densities on native macroalgal diversity, through effects of shading and space  6  competition, two common competitive mechanisms in marine macroalgae (BrittonSimmons 2004). The fourth and fifth chapters focus on the mechanisms responsible for S. muticum becoming a successful invader in British Columbia. One potential mechanism for the success of introduced species is the Enemy-Release Hypothesis (ERH, Keane and Crawley 2002), which states that the successful colonization and proliferation by an introduced species is a result of release from co-evolved enemies. I test the prediction of the ERH that in non-native regions, S. muticum will be less grazed by native herbivores than co-evolved native macroalgae. A proposed mechanism for the reduced grazing of introduced plants compared to native plants in the invaded range is differences in the levels of chemical defenses expressed. As a result, introduced plants require less energy for herbivore defense, meaning more is available for growth and reproduction, allowing for an increase in competitive ability compared to native species (Evolution of Increased Competitive Ability hypothesis (EICA) sensu Blossey and Notzold 1995). Here I combine community level studies of local species interactions, and investigations of defensive chemical ecology at both the community and biogeographic level to test these theories and investigate the traits responsible for invasion success of S. muticum in British Columbia.  7  LITERATURE CITED Andrew NL and Viejo RM (1998) Effects of wave-exposure and intraspecific density on the growth and survivorship of Sargassum muticum (Sargassaceae: Phaeophyta). European Journal of Phycology 33: 251-258. Battershill C, Miller K and Cole R (1998) The understorey of marine invasions. Seafood New Zealand 6: 31-33. Bax N, Williamson A, Aguero M, Gonzalez E, Geeves W (2003) Marine invasive alien species: a threat to global biodiversity. Marine Policy 27: 313-323. Blossey B and Nötzold R (1995) Evolution of Increased Competitive Ability in Invasive Nonindigenous Plants: A Hypothesis. J. Ecol. 83: 887-889. Britton-Simmons KH (2004) Direct and indirect effects of the introduced alga Sargassum muticum on benthic, subtidal communities of Washington State, USA. Mar. Ecol. Prog. Ser. 277: 61-78. Britton-Simmons KH and Abbott KC (2008) Short- and long-term effects of disturbance and propagule pressure on a biological invasion. Journal of Ecology 96: 68-77. Callaway RM and Ridenour WM (2004) Novel weapons: invasive success and the evolution of increased competitive ability. Front. Ecol. Environ. 2: 419-426 Carlton JT (2000) Global change and biological invasions in the Oceans. In: Invasive Species in a Changing World. Eds: Mooney HA and Hobbs RJ. Island Press, Washington DC, pp. 31-53. Casas G, Scrosati R, Piriz ML (2004) The invasive kelp Undaria pinnatifida (Phaeophyceae, Laminariales) reduces native seaweed diversity in Nuevo Gulf (Patagonia, Argentina). Biol. Inv. 6: 411-416. Critchley AT (1981) Ecological studies on Sargassum muticum (Yendo) Fensholt. Ph.D thesis. Portsmouth Polytechnic, UK. De Wreede RE and Vandermeulen H (1988) Lithothrix aspergillum (Rhodophyta): regrowth and interaction with Sargassum muticum (Phaeophyta) and Neorhodomela larix (Rhodophyta). Phycologia 27: 469-476. Galil BS, Nehring S, Panov V (2007) Waterways as Invasion Highways-Impact of Climate Change and Globalization Ecological Studies. In: Biological Invasions. Ed: Nentwig W. Springer-Verlag Berlin Heidelberg.193, pp. 59-74.  8  IUCN SSC Invasive Species Specialist Group (ISSG). Website: www.issg.org. Accessed August 2009. Johnson CR and Chapman ARO (2007) Seaweed invasions: introduction and scope. Bot. Mar. 50: 321-325. Keane RM and Crawley MJ (2002) Exotic plant invasions and the enemy release hypothesis. Trends Ecol. Evol. 17: 164-170. Kennedy TA, Naeem S, Howe KM, Knops JMH, Tilman D, Reich P (2002) Biodiversity as a barrier to ecological invasion. Nature 417: 636-638. Klein J, Ruitton S, Verlaque M, Boudouresque C (2005) Species introductions, diversity and disturbances in marine macrophyte assemblages of the northwestern Mediterranean Sea. Mar. Ecol. Prog. Ser. 290: 79-88.  Levin PS, Coyer JA, Petrik R, Good TP (2002) Community-wide effects of nonindigenous species on temperate rocky reefs. Ecology 83: 3182-3193. Lewey S and Farnham WF (1981) Observations on Sargassum muticum in Britain. Proceedings of the International Seaweed Symposium 8: 388-394. Lyons KG and Schwartz MW (2001) Rare species loss alters ecosystem functioninvasion resistance. Ecol. Lett. 4: 358-365. Meinesz A, Belsher T, Thibaut T, Antolic B, Ben Mustapha K, Boudouresque C-F, Chiaverini D, Cinelli F, Cottalorda JM, Djellouli A, El Abed A, Orestano C, Grau AM, Ivesa L, Jaklin A, Langar H, Massuti-Pascual E, Peirano A, Tunesi A, de Vaugelas J, Zavodnik N, Zuljevic A (2001) The introduced green alga Caulerpa taxifolia continues to spread in the Mediterranean. Biol. Invas. 3: 201-210. Mineur F, Johnson MP, Maggs CA (2008) Non-indigenous marine macroalgae in native communities: a case study in the British Isles. J. Mar. Biol. Ass. UK. 88: 693-698. Mollison D (1986) Modelling biological invasions: chance, explanation, prediction. Phil. Trans. R. Soc. Lond. 314: 675-694. Mollo E, Gavagnin M, Carbone M, Castelluccio F, Pozone F, Roussis V, Templado J, Ghiselin MT, Cimino G (2008) Factors promoting marine invasions: A chemoecological approach. Proc. Natl. Acad. Sci. USA. 4582-4586. Norton TA (1977b) The growth and development of Sargassum muticum (Yendo) Fensholt. J. Exp. Mar. Biol. Ecol. 26: 41-53.  9  Norton TA (1976) Why is Sargassum muticum so invasive? British Phycological Journal 11: 197-198. Norton TA and Deysher LE (1989) The reproductive ecology of Sargassum muticum at different latitudes. In: Reproduction, genetics and distributions of marine organisms Eds: Ryland JS and Tyler PA. pp. 147-152. Olsen and Olsen, Fredensberg. Pimentel D, Zuniga R, Morrison D (2005) Update on the environmental and economic costs associated with alien-invasive species in the United States. Ecol. Economics. 52: 273-288. Pimentel D, Lach L, Zuniga R, Morrison D (2000) Environmental and economic costs of non-indigenous species in the United States. Bioscience 50: 53-65. Ribera Siguan MA (2005) Pathways of Biological Invasions of Marine Plants. In: Invasive species: vectors and management strategies. Eds: Ruiz MG and Carlton JT. Island Press, pp. 183-226. Ribera MA and Boudouresque CF (1995) Introduced marine plants with special reference to macroalgae: mechanisms and impact. Progress in Phycological Research 11: 187268. Scagel RF (1956) Introduction of a Japanese alga, Sargassum muticum into the northeast Pacific. Fish. Res. Pap. Wash. Dep. Fish. 1: 49-59. Schaffelke B, Smith JE, Hewitt CL (2006) Introduced macroalgae - a growing concern. J. Applied Phycology. 18: 529-541. Schaffelke B and Hewitt CL (2007) Impacts of introduced seaweeds. Bot. Mar. 50: 397417. Simberloff D, Parker IM, Windle PN (2005) Introduced species policy, management, and future research needs. Front. Ecol. Environ. 3: 12-20. Smith JE, Hunter CL, Smith CM (2002) Distribution and reproductive characteristics of nonindigenous and invasive marine algae in the Hawaiian Islands. Pac. Sci. 56: 299-315. Thom RM, Hallum L (1990) Long-term Changes in the Areal Extent of Tidal Marshes, Eelgrass Meadows and Kelp Forests of Puget Sound. In: Final Report to Office of Puget Sound, Region 10. US Environmental Protection Agency No. EPA 910/9-91005 (FRI-UW-9008).  10  Tilman D (1997) Community invasibility, recruitment limitation, and grassland biodiversity. Ecology 78: 81-92. Verlaque M (1994b) Inventaire des plantes introduites en Me´diterranée: origines et répercussions sur l’environnement et les activités humaines. Oceanol Acta 17: 1-23. Villèle X and Verlaque M (1995) Changes and degradation in a Posidonia oceanica bed invaded by the introduced tropical alga Caulerpa taxifolia in the North Western Mediterranean. Bot. Mar. 38: 79-87. Walker DI and Kendrick GA (1998) Threats to macroalgal diversity: marine habitat destruction and fragmentation, pollution and introduced species. Bot. Mar. 41: 105112. Wasson K, Fenn K, Pearse JS (2005) Habitat differences in marine invasions of central California. Biol. Invas. 7: 935-948. Yoshida T (1983) Japanese species of Sargassum subgenus Bactrophycus (Phaeophyta, Fucales). J. Fac. Sci. Hokkaido Univ., Ser. V (Botany) 13: 99-246.  11  CHAPTER TWO DIVERSITY EFFECTS ON INVASION VARY WITH LIFE HISTORY STAGE IN MARINE MACROALGAE1  INTRODUCTION The relationship between community diversity and invasibility is a subject of ongoing debate in ecology. The diversity-invasion hypothesis proposed by Elton (1958) posits that more diverse communities should be less invasible. Based on the idea that more diverse communities show more complete or efficient use of resources, less niche space may be available for new species to occupy. The hypothesis has received much recent attention, primarily in the form of experimental studies with terrestrial plants and surveys comparing the number and abundance of exotic species with native diversity. Reviews of the literature highlight the lack of consensus on the strength and direction of the effects of diversity on community resistance to invasion (Levine and D’Antonio 1999, Levine et al. 2002). The majority of small scale experimental studies have found negative relationships between native diversity and invader success (Tilman 1997, Crawley et al. 1999, Knops et al. 1999, Stachowicz et al. 1999, 2002, Naeem et al. 2000, Symstad 2000, Lyons and Schwartz 2001, Kennedy et al. 2002). By contrast, large scale observational studies have mostly reported positive correlations between the number of native and invasive species (Planty-Tabacchi et al. 1996, Lonsdale 1999, Stohlgren et al. 1999, Levine 2000, Levine et al. 2002, Herben et al. 2004). Native and exotic diversity may be  1  A version of this chapter has been published. White LF and Shurin JB (2007) Diversity effects on invasion vary with life history stage in marine macroalgae. Oikos 116: 1193-1203.  12  positively correlated even if diversity enhances invasion resistance if native and non-native species respond similarly to environmental factors (Levine 2000, Sax 2002, Byers and Noonburg 2003). A number of studies in terrestrial plants have suggested that although higher diversity increases resistance to invasion, other environmental factors and the supply of colonists play a greater role in determining the prevalence of exotic species (Lonsdale 1999). Despite strong interest in the diversity-invasibility hypothesis, the majority of studies have been done with plants in terrestrial systems. Two experimental aquatic studies, one in marine invertebrate communities (Stachowicz et al. 1999) and the other in freshwater zooplankton (Shurin 2000), also found negative effects of native diversity on the success of invaders. However, aquatic and terrestrial communities may differ in ways that affect the relationship between diversity and invasibility. For instance, commensalisms are common in marine intertidal macroalgae that interact by modifying the physical environment, preventing stress due to desiccation and physical disturbance (Bruno et al. 2003). In addition, macroalgae may compete for light and space but to a lesser extent for mineral nutrients since water containing limiting resources is constantly refreshed. Macroalgae are unable to affect nutrients as terrestrial plants do, so diversity may or may not lead to greater intensity of competitive interactions in macroalgae. Alternatively, if diverse communities occupy more space for more time, then space or light competition may be greater with more species present. The relationship between  13  native diversity and invasibility may therefore be similar or different from that in terrestrial plants. We examined the relationship between native diversity and invasibility in experimental and unmanipulated communities of low intertidal macroalgae. All research was conducted in Barkley Sound, on the west side of Vancouver Island, British Columbia. We surveyed natural macroalgal communities around the Sound to test the relationship between diversity of native species, the availability of open space, and the abundance of the exotic macroalga Sargassum muticum. If high native richness excludes colonization by S. muticum, or if the invader excludes native taxa, then we expected to find a negative correlation between the number of native species and the prevalence of the invader. However, because environmental heterogeneity can confound the direct effect of diversity on invasibility (Levine 2000), we also experimentally tested the effects of diversity, community composition, native cover and the availability of primary space (substratum) and invasion success of S. muticum. In addition, we examined the effects of macroalgal community structure on colonization by native benthic organisms.  MATERIALS AND METHODS Life history of Sargassum muticum Sargassum muticum is a distinctive macroalga that grows on hard substrate in shallow waters. Native to Japan, S. muticum is believed to have been introduced to Vancouver, Canada in the 1940s (possibly as early as 1902) attached to imported  14  Japanese oysters (Crassostrea gigas) and is now well established throughout the northeast Pacific. Locally abundant, S. muticum forms dense canopies spanning lower intertidal to subtidal regions, most commonly on wave-sheltered shores (Druehl 2000). Thalli can reach up to 10 metres in length, but are typically one to two metres in the intertidal zone in Barkley Sound. S. muticum has one diplontic life stage, and reproductive capabilities that facilitate its spread as an invader. It can reproduce within the first year of life, and is able to self-fertilize. Fertile branches break off from the holdfast and float away, a mechanism which proves to be an effective system for global spread (Deysher and Norton 1982). Gametes are released in cycles of 13 days instead of all simultaneously, increasing the odds some of them will encounter favourable tides and conditions. Germlings are then released following the spring tide (Fletcher 1975). Sargassum muticum is considered a competitive dominant because it is extremely fast growing, with records of 4 cm day-1 for May to June (Jephson and Gray 1977). It can colonize both disturbed habitats (DeWreede 1983) and more diverse macroalgal assemblages (Sánchez and Fernández 2005). Sargassum muticum may have a competitive advantage over native species in that it reproduces prolifically, has larger and more developed zygotes, and may overshadow or simply overgrow other species. Once settled, S. muticum’s large perennial holdfast may allow it to occupy and hold space more effectively than other species.  15  Field patterns We conducted surveys of the marine flora in the low intertidal zone at three sites close to Bamfield Marine Sciences Centre to quantify the relationship between native macroalgal diversity and cover of exotic S. muticum in the field. Horizontal transects 50 m long were laid along the shores through the centre of the S. muticum zone, with a total of 23 50 x 50 cm quadrats sampled. All macroalgae and sessile organisms within quadrats sampled were identified and their percent cover visually estimated.  R2 = 0.42  Figure 2.1 Exotic Sargassum muticum cover (A), and available primary space (B) plotted against native richness, for observational field data (expressed as percentages). Larger squares represent multiple data points. Analysis was by linear least squares regression, where (A) R2 = 0.42, P = 0.001, and (B) R2 = 0.17, P = 0.14, n = 23.  Experimental manipulation We constructed macroalgal diversity gradients on experimental boards to investigate the diversity-invasion relationship while controlling for extrinsic covarying  16  factors such as species identity, wave exposure, nutrient or light availability or the abundance of grazers. Boards 25 x 25 cm were constructed from HardyPlank, a fibrous cement panelling, and subdivided into 25 tiles measuring 5 x 5 cm, with the centre tile designated for attachment to rock substrate. Observational field surveys indicated that the natural range of native macroalgal diversity in plots containing S. muticum is 3-6 species, with S. muticum most frequently inhabiting plots containing four species (Fig. 2.1A). We constructed plots containing all possible macroalgal species combinations over four richness levels (0, 1, 3 and 4 species). This design gave us 10 species composition treatments (Table 2.1) and allowed us to separate the effects of diversity from those of species composition (the identity of the species present) as well as control for the sampling effect driving effects of diversity as the abundance of each native individual was initially constrained. Each treatment was replicated five times for a total of 50 boards. The four native species selected to construct diversity gradients all co-occur with S. muticum in the low intertidal zone on rocky, sheltered shores: Ulva fenestrata (Chlorophyta), Odonthalia floccosa, Neorhodomela larix and Chondracanthus exasperatus (latter three Rhodophyta). Macroalgae were collected and kept in flowthrough seawater tanks where regular emersion periods simulated tides, and water temperature and ambient light were controlled using mesh screens. Epiphytes and epifauna were removed before seeding macroalgae into experimental plots so species number could be controlled.  17  Table 2.1 Ten compositional combinations (a-j) of native species over four diversity levels (0, 1, 3, 4) used for constructed plots. The number of individual macroalgal thalli per species used for each compositional treatment are shown.  We collected whole native macroalgal thalli with holdfasts intact, which we glued directly to experimental boards using Krazy Glue. Preliminary trials were conducted to assess potential damage gluing holdfasts caused to the macroalgae. In three species, C. exasperatus, N. larix and O. floccosa, attaching holdfasts directly to the boards did not negatively affect the macroalgae. During trials, U. fenestrata became damaged when in contact with the glue. This macroalga was collected with holdfasts attached to rocks which were later chipped small enough to attach within the 5 x 5 cm tiles. Macroalgae were randomly assigned to tiles within the 5 x 5 cm grid of the plots. Each compositional treatment started with 24 individuals but differed in density and native cover due to differences in individual size and morphology among species. The boards were in seawater tanks for six days before being deployed into the low intertidal. This was necessary for all plots to be seeded before they were deployed. Replicates were assigned to five blocks along the shore, with one replicate of each treatment within each block. Blocks were evenly spaced (7.6 metres between the centre  18  of neighbouring blocks) along a horizontal transect in the lower intertidal zone. Spacing blocked replicates afforded control for local processes on a scale of 101 metres. For maximum recruitment of exotic germlings, boards were placed directly into a dense S. muticum bed which spanned the lower intertidal fringe the length of the shore. Experimental plots were deployed in the low intertidal on 3rd July 2005 on a wavesheltered, rocky shore behind Dixon Island (48˚51.141´N, 125˚06.839´W). The upper intertidal was dominated by fucoids, and the mid- and low-intertidal largely by exotic S. muticum, with U. fenestrata, C. exasperatus and Leathesia difformis. Experimental plots were monitored bimonthly, each monitoring occurring over three low tides as there was too much work to do during one low tide. We monitored until 18th September 2005, when exotic germlings had successfully settled and shown growth and native macroalgae were senescing. Percent cover of native and exotic macroalgal species, and available primary space were visually estimated on each sampling by placing a 25 x 25 cm quadrat subdivided into 25 tiles above the plots (following Dethier et al. 1993). Cover of native macroalgal species was divided into percent canopy cover and basal cover. Available primary space was taken as the percentage of unoccupied space per plot. Cover of sand was also quantified but not included as available space as it was not clear whether germlings would adhere to it. Colonization by native macroalgal taxa was taken as the percent cover of native recruits per plot, and was monitored to assess whether native species responded similarly to the treatments as exotic S. muticum. We also quantified the number of mobile grazers, such as chitons (Katharina tunicata), isopods (Idotea wosnesenskii), snails (Tegula funebralis) and limpets (Tectura scutum) in  19  the experimental plots. We quantified exotic recruitment as percent cover since individual S. muticum germlings were initially difficult to count.  Data analysis Invasibility was measured as the percent cover of S. muticum germlings recruiting to the plots. Data were found to be non-normally distributed and so were arcsine square root transformed for normality. We analyzed for effects of diversity treatment (0, 1, 3 and 4 species) and species composition (10 combinations) on exotic cover, availability of primary space and native recruitment in separate randomized block, repeated-measures ANOVAs. Linear regression was used to test associations between S. muticum abundance and native diversity in the field survey, and associations between S. muticum and diversity, community composition, native basal and canopy cover, primary space, and grazing on individual sampling dates in the experiment. Orthogonal contrasts were performed to test for differences between compositional combinations within the 1- and 3- species diversity treatments. All analyses were performed using JMP 4.0.4 (SAS Inst.).  RESULTS Observed field patterns Sargassum muticum density was higher in plots with lower native diversity in the field survey (R2 = 0.42, P = 0.001, Fig. 2.1A). Percentage of available primary space was not related to native diversity (R2 = 0.17, P = 0.14, Fig. 2.1B), indicating that diversity did not reduce availability of primary space.  20  Experimental manipulation - community diversity Cover of S. muticum germlings initially increased over time in our experimental plots (Fig. 2.2A), due to their release in cycles of 13 days, just after the spring tide. The first exotic germlings were recorded on 21st July, (day 23 of monitoring) coinciding with the spring tide. New germlings colonized experimental plots until 2nd September, (day 62) after which only growth and losses of settled recruits occurred (Fig. 2.2A). We found significant effects of diversity on cover of exotic S. muticum over the duration of the experiment (F3,42 = 2.85, P = 0.05, Table 2.2A). However greatest initial recruitment occurred in the highest diversity treatment, which contrasts with the predictions of the diversity-invasion hypothesis (Fig. 2.2A). Diversity level one, the monoculture treatments, had the lowest recruitment (Fig. 2.2A). Colonization of S. muticum was positively correlated with native diversity at day 47 (R2 = 0.26, P = 0.0001, Fig. 2.3A) indicating invasion was highest in more diverse plots. The relationship between S. muticum cover and native diversity at day 47 (time of maximum exotic germling abundance) was different from day 77 (exotic abundance at the end of the experiment, Fig. 2.3B). Despite the high initial invasion, the highest diversity level also experienced the highest loss of settled germlings (Fig. 2.3B). At day 77 diversity was negatively correlated with exotic cover, with greater invader abundance in the lowest diversity treatments (R2 = 0.16, P = 0.004, Fig. 2.3B). Our results suggest diversity facilitates early recruitment of exotic S. muticum, but this relationship is lost over time. Cover of the exotic declined in all plots from day 62 to day 77 (Fig. 2.2A), indicating that loss of settled recruits was greater than growth (Fig. 2.3C).  21  We found no relationship between abundance of benthic herbivores and cover of S. muticum germlings over the duration of the experiment (all P = 0.1). There was also no relationship between herbivores and diversity (F3,42 = 0.41, P = 0.75) or composition (F9,36 = 1.47, P = 0.20). We found significant differences in abundance of exotic germlings between blocks (P = 0.005). A number of factors such as varying supply of exotic germlings along the shore could have contributed to this block effect.  Community composition, availability of primary space and native cover Resident community composition had no effect on invasion of exotic S. muticum over the duration of the experiment (F9,36 = 1.27, P = 0.29, Table 2.2A, Fig. 2.4A,B). Repeated-measures ANOVA found no differences between the monoculture treatments over time, suggesting there was no single species effect on exotic recruitment (F3,12 = 0.75, P = 0.54). The diversity effects were therefore independent of species composition. Repeated-measures ANOVA indicated that neither community diversity (F3,42 = 2.63, P = 0.06) or composition (F9,36 = 1.66, P = 0.14, Fig. 2.4G,H) had significant effects on the availability of primary space over the duration of the experiment (Table 2.2C). We found no relationships between native canopy or basal cover and recruitment of S. muticum on any sampling date (all P = 0.1). Native cover was not greater in the more diverse systems. Orthogonal contrasts showed native cover was highest in the monoculture treatments for the first four sampling dates (all P = 0.03). There were no significant differences in native cover between the monoculture and the three species polyculture at day 47 (P = 0.18) and marginally  22  significant differences at day 77 (P = 0.06). There were significant differences in canopy cover between monocultures (day 47: P = 0.0001, day 77: P = 0.001, Fig. 2.4C) and marginal significance within 3 species polycultures (day 47: P = 0.05, day 77: P = 0.07, Fig. 2.4D). Within the monoculture treatments, the two macroalgal species with bladed morphologies, C. exasperatus and U. fenestrata had the highest canopy cover (P = 0.05, Fig. 2.4C,D).  23  Figure 2.2 Time series of percent (A) cover of exotic Sargassum muticum germlings, (B) native cover (canopy and basal cover combined) and (C) native recruitment, per diversity level. Error bars are SE.  24  Table 2.2 Repeated-measures ANOVA test criteria for (A) exotic invasion, (B) native recruitment, and (C) availability of primary space, expressed as percentages. Within subjects results are Wilks’ Lambda conservative estimates.  Recruitment of native species Recruitment of native species into the experimental plots increased over time (Fig. 2.2C) but was not affected by either community composition (F9,30 = 0.49, P = 0.87) or native diversity (F3,42 = 0.91, P = 0.45, Table 2.2B). There were also no effects of community composition within diversity level on native recruitment (diversity level 1, day 47: P = 0.18, day 77: P = 0.33; diversity level 3, day 47: P = 0.75, day 77: P = 0.91). Consistent with exotic recruitment, we found no relationship between native recruitment and abundance of benthic herbivores throughout the experiment (all P = 0.3).  DISCUSSION The relationship between native diversity and invasion by exotic Sargassum muticum depends on seasonal patterns and the stage of invasion. Unmanipulated  25  communities show that S. muticum is less abundant in areas with high diversity of natives. We found a negative relationship between native diversity and cover of exotic S. muticum (Fig. 2.1A). This pattern contrasts with a number of studies of terrestrial plants that have found positive relationships between native and exotic diversity (PlantyTabacchi et al. 1996, Lonsdale 1999, Stohlgren et al. 1999, Levine 2000, Levine et al. 2002, Herben et al. 2004). These studies differed from ours in that they examined correlations between native and exotic diversity, whereas ours measured the abundance of the only established invasive macroalga in the region. One terrestrial study by Robinson et al. (1995) examined the effects of native species richness of a California winter annual grassland on a single ‘‘invader’’, a native ruderal, California poppy. In this study the invader also established and reproduced better in sites with higher species richness, lending support for differences between marine and terrestrial systems. Positive correlations between native and exotic diversity may be interpreted as indications that the environmental factors that promote high native diversity also facilitate establishment of rich exotic communities, not that native diversity per se promotes invasion (Levine 2000). The observed negative correlation between native diversity and exotic cover has three potential mechanistic interpretations: (1) native diversity may have a direct negative effect on exotic recruitment or growth, (2) S. muticum invasion may exclude native species and thereby reduce native diversity, or (3) S. muticum’s environmental niche may incidentally coincide with areas of low native richness. Patterns in unmanipulated communities are unable to disentangle these possibilities, therefore we coupled our field surveys with experimental manipulations.  26  Figure 2.3 Diversity level plotted against (A) initial Sargassum muticum cover, (B) final S. muticum cover, and (C) change in cover of S. muticum from day 47 to day 77 (expressed as percentages). Initial cover was taken at day 47, when abundance of exotic recruitment was highest. Final cover was taken at the end of the experiment (day 77). Change in S. muticum cover was calculates as % cover at day 77- day 47. Each data point represents cover of S. muticum recruits averaged over five replicates, per compositional treatment. Analysis was by linear least squares regression, where (A) R2 = 0.26, P < 0.0001, (B) R2 = 0.16, P = 0.004, and (C) R2 = 0.38, P < 0.0001, n = 50.  27  Experimental manipulation Our experimental manipulations of diversity found greater settlement of new recruits in more diverse plots. However, losses of settled recruits were also highest from these high diversity treatments. The slope of the relationship between native diversity and exotic cover reversed over the duration of the experiment (Fig. 2.3). High native diversity may enhance settlement but reduce growth or survival of new recruits (Fig. 2.3). Despite initial opposite relationships between field patterns and experimental manipulations (Fig. 2.1A, 2.3A), results converged as the experiment progressed, suggesting that experimental results depend on the temporal scale of the study. Our results contrast with recent findings of Arenas et al. (2006) that functional diversity was more important than species diversity in determining invasibility of marine macroalgal assemblages. However, our treatments were selected to represent a gradient of diversity of the most common native taxa, not a wide range of functional morphology.  Influences of phenology Invasion success of S. muticum in this region may be due to phenological differences from native taxa. Maximum abundance of exotic germlings (day 47) coincided with a sharp decrease in native cover as mature macroalgae senesced, both in our plots and naturally on the shore (Fig. 2.2A,B). Recruitment by native species was lower than that of S. muticum at the peak of exotic germling abundance (day 47, Fig. 2.2A,C), after which native recruitment increased. Exotic S. muticum cover declined late  28  Figure 2.4 Bar charts of (A, B) cover of Sargassum muticum germlings, (C, D) canopy cover, (E, F) basal cover, and (G, H) availability of primary space. Data are expressed as percentages, plotted against compositional treatment (Table 2.1) within diversity level (0, 1, 3, 4). The highest abundance of exotic recruitment occurred on day 47, day 77 is  29  Figure 2.4 (Continued) recruitment at the end of the experiment. Data are averaged cover five replicates, per compositional treatment. Statistical analysis was by univariate ANOVA, where (A) P = 0.03, (B) P = 0.28, (C) P = 0.007, (D) P = 0.02, (E) P < 0.0001, (F) P < 0.0001, (G) P = 0.65, and (H) P = 0.42. Controls were omitted from analysis of percent canopy cover (C and D) as they violated the assumption of homogenous variance. Omitting controls from the remaining analysis: (A) P = 0.03, (B) P = 0.14, (E) P = 0.002, (F) P = 0.11, (G) P = 0.76, and (H) P = 0.34. Error bars are SE.  in the experiment as settled recruits either grew or were lost. This suggests that S. muticum occupies a distinct temporal niche from the native taxa. Stachowicz et al. (2002) suggested that diverse intertidal communities repel invasion because phenological differences between taxa lead to greater space utilization when more species are present. Seasonal fluctuations in the abundance and recruitment patterns of individual species are often out of phase. Successful invaders may be those that have seasonal recruitment periods that coincide with temporal minima in the recruitment of natives (Stachowicz and Tilman 2005). However, our results differed from Stachowicz in that diverse native communities did not have higher native cover or show greater space utilization over time, indicating that space competition was not driving competitive effects on invasion in this system. Native cover was highest in the low diversity systems for the first four sampling dates (Fig. 2.2B). The reason for the high initial cover in the monoculture treatments is unknown, but may be responsible for the low initial colonization of S. muticum at low diversity, thereby driving the ‘facilitative’ effect of diversity on invader recruitment.  30  Increased diversity may also affect ecosystem functioning through hydrodynamic effects on water flow that can influence the physical forces exerted on new recruits of the invader (Cardinale et al. 2002). Numbers of mobile herbivores were unaffected by the diversity manipulations, indicating that the response was not due to covariation between native macroalgal diversity and grazing pressure. Early recruitment of exotic S. muticum was highest in higher diversity communities, suggesting the increased structural complexity of high diversity macroalgal communities may alter topographic complexity caused by variation in the morphology of individuals, thereby providing shelter for S. muticum recruits to settle to the substrate.  Limiting resources related to life history stages The influence of native diversity on invasibility in this system varies over life history stages of the invader, with different determinants of invasibility becoming important at different times. Our results suggest that high native diversity creates conditions that facilitate early recruitment of exotic S. muticum (Fig. 2.3A) but decreases survivorship (Fig. 2.3C). Stachowicz et al. (2002) reported marine invertebrate assemblages with higher native diversity had lower available primary space, providing a mechanism for higher invasion success in lower diversity plots. Contrary to this, in our experiment repeated-measures ANOVA found no overall relationship between native diversity and available primary space (Table 2.2C). The mechanisms driving diversity effects on invasibility once exotic germlings are settled is unknown but might include factors such as temperature, desiccation or ambient light. While one factor may exert the  31  strongest influence at a given life history stage, it is the cumulative influence of many factors, in this case community diversity, cover, and the phenologies of native species that comprise the community that determine its invasibility. Diversity may have positive effects on some stages of the invader, however these may be overwhelmed by negative effects at other stages, leading to negative associations between native diversity and exotic cover, as we observed in our field survey.  Native recruitment vs exotic invasion In this system, the ecosystem determinants of resistance to exotic S. muticum invasion differ from those that affect establishment of native species (Table 2.2). Diversity and composition of native macroalgae had no effect on recruitment of native species. These results suggest differences between the effects of diversity on the invasion and recruitment processes of exotic and native species. There were no block effects on native recruitment (all P > 0.2, Table 2.2B), unlike exotic recruitment. These differences could be attributed to different phenologies of colonizing native species.  Marine vs terrestrial systems Our results indicate that the relationship between native diversity and the success of invaders shows both similarities and contrasts between marine and terrestrial ecosystems. The negative correlation between exotic cover and native diversity in our field observations deviates from the positive relationship generally reported from terrestrial studies (Robinson et al. 1995, Stohlgren et al. 1999, Levine et al. 2002), suggesting  32  support for structural differences between systems. However, positive correlations have most often been observed at much larger geographic scales than our study (Lonsdale 1999, Stohlgren et al. 1999). We found no differences in native cover between monoculture and three species polycultures (Fig. 2.4C-F) which is also inconsistent with terrestrial systems where more diverse polycultures generally have higher native cover (Tilman 1997, Hooper et al. 2005). The overall negative effect of native diversity on invader success in our experiment is consistent with empirical terrestrial studies and may be more general than previously thought. A competitive trait in terrestrial plants is the ability to acquire limiting resources such as nutrients. Diversity may not affect competition for nutrients in macroalgae as they are constantly refreshed with seawater flow, suggesting macroalgae may be less able to reduce the availability of limiting resources than terrestrial flora. The identity of limiting resources (i.e. space, light, nutrients) may differ across ecosystems (Stachowicz and Byrnes 2006), which suggests a variety of mechanisms may produce the same result. Further work in intertidal, macroalgal communities is warranted to determine the mechanisms driving diversity effects on invasibility once invaders are established, whether these results are species or location specific, and whether they hold across macroalgal taxa and biogeographic region. Terrestrial studies investigating how temporal scale is related to the diversity-invasion relationship would provide insight into whether these results are limited to intertidal systems. Although the mechanisms linking native diversity to invasibility appear to differ between marine and terrestrial systems, the  33  negative effect of diversity on invasibility appears to be more common than previously thought and may be subject to generalization across the two systems.  34  LITERATURE CITED Arenas F, Sánchez I, Hawkins SJ, Jenkins SR (2006) The invasibility of marine algal assemblages: roles of functional diversity and identity. Ecology 87: 2851-2861. Bruno JF, Stachowicz JJ, Bertness MD (2003) Inclusion of facilitation into ecological theory. Trends Ecol. Evol. 18: 119-125. Byers JE and Noonburg EG (2003) Scale dependent effects of biotic resistance to biological invasion. Ecology 84: 1428-1433. Cardinale BJ, Palmer MA, Collins SL (2002) Species diversity enhances ecosystem functioning through interspecific facilitation. Nature 415: 426-429. Crawley MJ, Brown SL, Heard MS, Edwards GR (1999) Invasion-resistance in experimental grassland communities: species richness or species identity? Ecol. Lett. 2: 140-148. DeWreede RE (1983) Sargassum muticum (Fucales, Phaeophyta): regrowth and interaction with Rhodomela larix (Ceramiales, Rhodophyta). Phycologia 22: 153160. Dethier MN, Graham ES, Cohen S, Tear LM (1993) Visual versus random-point percent cover estimations: ‘‘objective’’ is not always better. Mar. Ecol. Prog. Ser. 96: 93100. Deysher LE and Norton TA (1982) Dispersal and colonization in Sargassum muticum (Yendo) Fensholt. J. Exp. Mar. Biol. Ecol. 56: 179-195. Druehl LD (2000) Pacific seaweeds. Harbour publishing. Elton CS (1958) The ecology of invasions by plants and animals. Methuen & Co. Fletcher RL (1975) Studies on the recently introduced brown alga Sargassum muticum (Yendo) Fensholt I. Ecology and reproduction. Bot. Mar. 18: 149-156. Herben T, Mandak B, Bimova K, Munzbergova Z (2004) Invasibility and species richness of a community: a neutral model and a survey of published data. Ecology 85: 3223-3233. Hooper, DU, Chapin III FS, Ewel JJ, Hector A, Inchausti P, Lavorel S, Lawton JH, Lodge DM, Loreau M (2005) Effects of biodiversity on ecosystem functioning: a consensus of current knowledge. Ecol. Monogr. 75: 3-35.  35  Jephson NA and Gray PWG (1977) Aspects of the ecology of Sargassum muticum in the Solent region of the British Isles. I. The growth cycle and epiphytes. In: Biology of benthic organisms. Eds: Keegan, B. F. et al. Proc. 11th Eur. Mar. Biol. Symp. Pergamon Press, pp. 367-375. Kennedy TA, Naeem S, Howe KM, Knops JMH, Tilman D, Reich P (2002) Biodiversity as a barrier to ecological invasion. Nature 417: 636-638. Knops JMH, Tilman D, Naeem S, Mitchell CE, Haarstad MJ, Ritchie ME, Howe KM, Reich PB, Siemass E, Groth J (1999) Effects of plant species richness on invasion dynamics, disease outbreaks, insect abundances and diversity. Ecol. Lett. 2: 286293. Levine JM (2000) Species diversity and biological invasions: relating local process to community pattern. Science 288: 852-854. Levine JM and D’Antonio CM (1999) Elton revisited: a review of evidence linking diversity and invasibility. Oikos 87: 15-26. Levine JM, Kennedy T, Naeem S (2002) Neighborhood scale effects of species diversity on biological invasions and their relationship to community patterns. In: Loreau, M. et al. (eds), Biodiversity and ecosystem functioning. Oxford Univ. Press, pp. 114124. Lonsdale WM (1999) Global patterns of plant invasions and the concept of invasibility. Ecology 80: 1522-1536. Lyons KG and Schwartz MW (2001) Rare species loss alters ecosystem function: invasion resistance. Ecol. Lett. 4: 358-365. Naeem S, Knops JMH, Tilman D, Howe KM, Kennedy T, Gale S (2000) Plant diversity increases resistance to invasion in the absence of covarying factors. Oikos 91: 97108. Planty-Tabacchi A, Tabacchi E, Naiman RJ, Deferrari C, DéCamps H (1996) Invasibility of a species-rich community in Riparian zone. Conserv. Biol. 10: 598. Robinson GR, Quinn JF, Stanton ML (1995) Invasibility of experimental habitat islands in a California winter annual grassland. Ecology 76: 786-794. Sánchez I and Fernández C (2005) Impacts of the invasive seaweed Sargassum muticum (Phaeophyta) on an intertidal macroalgal assemblage. J. Phycol. 41: 923-930.  36  Sax DF (2002) Native and naturalized plant diversity are positively correlated in scrub communities of California and Chile. Div. Distr. 8: 193-210. Shurin JB (2000) Dispersal limitation, invasion resistance, and the structure of pond zooplankton communities. Ecology 81: 3074-3086. Stachowicz JJ and Tilman D (2005) Species invasions and the relationships between species diversity, community saturation, and ecosystem functioning. In: Sax, D. F. et al. (eds), Species invasions: insights into ecology, evolution and biogeography. Sinauer, pp. 42-64. Stachowicz JJ and Byrnes JE (2006) Species diversity, invasion success and ecosystem functioning: combining experimental and observational approaches to assess the roles of resource competition, facilitation and extrinsic factors. Mar. Ecol. Prog. Ser. 311: 251-262. Stachowicz JJ, Whitlatch RB, Osman RW (1999) Species diversity and invasion resistance in a marine ecosystem. Science 286: 1577-1579. Stachowicz JJ, Fried H, Osman RW, Whitlatch RB (2002) Biodiversity, invasion resistance, and ecosystem function: reconciling pattern and process. Ecology 83: 2575-2590. Stohlgren TJ, Binkley D, Chong GW, Kalkhan MA, Schell LD, Bull KA, Otsuki Y, Newman G, Bashkin M, Son Y (1999) Exotic plant species invade hot spots of native plant diversity. Ecol. Monogr. 69: 28-46. Symstad AJ (2000) A test of the effects of function group richness and composition on grassland invasibility. Ecology 81: 99-109. Tilman D (1997) Community invasibility, recruitment limitation, and grassland biodiversity. Ecology 78: 81-92.  37  CHAPTER THREE DENSITY DEPENDENT EFFECTS OF AN EXOTIC MARINE MACROALGA ON NATIVE COMMUNITY DIVERSITY INTRODUCTION The introduction and spread of exotic species have caused dramatic changes in marine macroalgal assemblages around the world (e.g. Trowbridge 1995, Stæhr et al. 2000, Boudouresque and Verlaque 2002, Williams and Smith 2007). The question of whether exotics cause net increases in local or regional species biodiversity, or exclude natives has raised much controversy. The effects on community richness can be positive (Dunstan and Johnson 2004, Mineur et al. 2008), negative (Stachowicz et al. 2002, Wasson et al. 2005), or neutral (Klein et al. 2005). Further, impacts of exotics can vary in response to disturbance (Rejmánek 1989), phenology (Thomsen et al. 2006), speciesspecific traits of the invader (i.e. life history stage, White and Shurin 2007), the structure of the native assemblage, or density of the exotic (Reush and Williams 1998). As a result few generalities about the interaction between exotic and native species have emerged, making predictions about the impacts of exotics on native species difficult. Exotic macroalgae impact native macroalgal community structure and diversity (Nyberg 2007) by altering the physical, chemical, and biotic characteristics of the habitat. Some macroalgal invasives have devastating impacts to whole ecosystems, i.e. Caulerpa taxifolia in the Mediterranean, Codium fragile ssp. tomentosoides in the western Atlantic. 1  A version of this chapter will be submitted for publication. White LF and Shurin JB (2010) Density dependent effects of an exotic marine macroalga on native community diversity.  38  However, mechanisms underlying the impacts of macroalgal invaders are uncertain and inferences about patterns are hampered because impact studies are available for only a few species, cover only a fraction of their introduced distribution, and are generally conducted over short time scales (Schaffelke and Hewitt 2007). The macroalga Sargassum muticum (Yendo) Fensholt (Phaeophyceae: Fucales) is one of the most aggressive marine invaders (Norton 1976, Boudouresque and Verlaque 2002). Native to South East Asia, S. muticum was introduced to the Pacific Northwest around the 1940’s as a hitchhiker species with Japanese oysters (Crassostrea gigas) imported for aquaculture (Scagel 1956). Sargassum muticum occupies lower intertidal and subtidal habitats, and forms a conspicuous component of the intertidal macroalgae owing to its large size and radial morphology. Since its introduction, competitive effects by S. muticum on a number of native macroalgal species have been shown, including Lithothrix aspergillum (DeWreede and Vandermeulen 1988), Laminaria spp., Cystoseira spp., Scytosiphon lomentaria, Gracilaria verrucosa, bull kelp Nereocystis luetkeana (Thom and Hallum 1990), the giant kelp Macrocystis pyrifera (Ribera and Boudouresque 1995), and Saccharina subsimplex (Britton-Simmons 2004). These studies indicate that S. muticum is a strong competitor relative to native taxa; however, the resources which it co-opts and the mechanisms of these effects are largely unknown. Limiting resources that are likely to give rise to competition between native and exotic macroalgae include nutrients, light and space (Arenas et al. 2006). Competition for mineral nutrients may be less important than for space and light since seawater containing limiting resources can be constantly refreshed by waves, currents, and upwelling. Some  39  studies have found that macroalgae in the Northeast Pacific are not nutrient limited (Wootton 1991, Pfister and Van Alstyne 2003). In nearby Washington State, BrittonSimmons (2004) compared nutrient levels in dense stands of S. muticum with plots where it was absent, and found no effect of S. muticum on nutrients. Alternatively, S. muticum is likely to influence other species through space and light competition. Staehr et al. (2000) reported significant changes in native macroalgal community structure in response to an increased abundance of S. muticum, indicating competitive displacement through competition for hard substrate and light. Sargassum muticum has perennial basal axes which can pre-empt space from annual native species that senesce each fall, preventing them from re-establishing (Britton-Simmons 2006). Britton-Simmons (2004) demonstrated negative effects of light competition by S. muticum in displacing underlying native macroalgae through shading. The tall stature and perennial basal axes and holdfast of S. muticum may contribute to its apparent competitive dominance in its invasive range. Native species may also exert competitive effects on S. muticum, providing biotic resistance in some resident macroalgal communities. We previously investigated the effects of native macroalgal community diversity on recruitment of exotic S. muticum, and showed negative and positive effects of native macroalgae at different life history stages of the invader (White and Shurin 2007). Diversity facilitated recruitment of S. muticum, but decreased growth and/or survivorship. We found higher cover of S. muticum in unmanipulated plots with lower native diversity, suggesting a negative relationship between diversity and invasibility at the neighbourhood scale. The negative  40  association between S. muticum and native diversity has three potential, non-mutually exclusive explanations: (1) S. muticum excludes natives through interspecific competition for limiting resources, (2) diverse native communities repel invasion by S. muticum through more intense competition, and (3) S. muticum responds negatively to environmental factors that produce higher native diversity. Our previous study tested (2), and the current study tests (1). If the negative correlation between S. muticum cover and native diversity observed in the field is the result of exclusion by S. muticum, then experimental reductions of S. muticum should lead to increased native diversity. We reduced S. muticum to a range of densities (0, 20, 40 and 60% cover) to ask whether the effects on the resident community depended on the density of the invader. We also experimentally separated the roles of space versus light competition by manipulating the overstory of S. muticum while either removing or leaving intact the basal holdfast.  MATERIALS AND METHODS Experimental field manipulations We used field experiments to experimentally investigate the impact of invasion by Sargassum muticum on native macroalgal communities in Barkley Sound, British Columbia, Canada. To determine the effects of S. muticum on the richness, diversity, and cover of native macroalgae, we established permanent plots spanning a sheltered, rocky shore behind Dixon Island (48˚51.15′N, 125˚07.01′W) in April 2006. S. muticum cover was manipulated by reducing its density in natural communities, as artificial seeding may restrict the opportunities for species to respond to interspecific interactions and  41  environmental heterogeneity, potentially limiting the expression of niche differences among species (Stachowicz et al. 2008). We randomly selected 1 x 1 metre plots within the low intertidal zone which contained at least 60% S. muticum cover, to match fieldobserved densities at the 1 x 1 metre scale. We used a two-factor block design with four treatment levels of S. muticum density. Control plots contained 60% S. muticum cover. The remaining three treatment densities were reduced to 40%, 20% and 0% cover of S. muticum. All plots were subjected to treatments as the control plots usually contained more than 60% exotic cover. Treatments were weeded to maintain these densities throughout the growing seasons. Each plot had a 0.25 metre buffer zone where S. muticum cover was weeded to match that inside the plots to minimize edge effects. The centre of each plot was marked with a drop-in masonry anchor which we relocated using a metal detector.  Light and space competition by Sargassum muticum The density of S. muticum was reduced in two ways that allowed us to separate the effects of light and space competition by S. muticum on native macroalgae. In the first treatment (the light/space competition treatment), we randomly removed whole S. muticum individuals, including the holdfast, from 1 x 1 metre plots. This created both primary space and increased light penetration to the substrate. Primary space on the substrate varied between S. muticum density treatments due to differences in the size and morphology of individuals. To test the effect of light competition alone (the light competition treatment), we removed only the annual frond, leaving the perennial basal  42  axes attached. This was achieved by snipping the annual frond at the base, leaving the holdfast attached to the substratum so that no primary space was created, but effects of shading through light competition were removed with the annual frond. The four density treatments, crossed with light/space competition treatments, were replicated six times (4 S. muticum density treatments x 2 light/space treatments x 6 replicates, n = 48 experimental plots). Natural substrate heterogeneity existed within and between manipulated plots. Substrate heterogeneity, like heterogeneity of soils for plants in terrestrial systems, can influence macroalgal species composition and performance (e.g. Lubchenco 1983, Airoldi 2000). To test if substrate type affected native community composition, we separated plots into three substrate categories (blocks) along the shore based on topographical features, i.e. noticeable substrate differences (sandy, rocky outcrop, and rock/sand combined) with 16 manipulated plots in each block. Macroalgae differ in morphology and the ways they utilize resources. To understand how community structure is influenced by S. muticum, species were assigned to three functional form groups based on similarities in morphology that influence functional use of light and space (following Britton-Simmons 2006). Canopy species occupy little substrate (small holdfasts), but have strong effects on light through large blades shading smaller species (i.e. Costaria costata); understory species also occupy little substrate, but have weaker effects on light due to their shorter statures than canopy species (i.e. Cryptosiphonia woodii), and basal species occupy the most substrate but have little impact on light as they grow close to the substrate (i.e. Hildenbrandia rubra).  43  A list of macroalgal species sampled during this study and their functional groupings are presented in Table 3.1.  Table 3.1 List of species sampled during study, with resource-based functional group assignment and code. See Gabrielson et al. (2000) for species authorities.  Species Ceramium sp. Chondracanthus exasperatus Corallina sp. Costaria costata Cryptosiphonia woodii Egregia menziesii Fucus distichus Gastroclonium subarticulatum Antithamnionella sp. Gracilaria sp. Halosaccion glandiforme Hildenbrandia rubra Leathesia difformis Macrocystis pyrifera Mastocarpus papillatus Mazzaella splendens Microcladia coulteri Neorhodomela larix Odonthalia floccosa Osmundea spectabilis ‘Petrocelis’ stage of M. papillatus Phyllospadix sp. Plocamium oregonum Prionitis lanceolata Pseudolithophyllum Sphacelaria sp. Scytosiphon lomentaria Ulva fenestrata Ulva intestinalis  Functional Group  FG code  understory understory basal canopy understory canopy understory understory understory understory understory basal understory canopy understory understory understory understory understory understory basal canopy understory understory basal understory understory understory  U U B C U C U U U U U B U C U U U U U U B C U U B U U U  basal  B  44  Competitive interactions may vary with time as resource availability changes between seasons and years (Thomsen et al. 2006). We predicted negative effects of both space and light competition by S. muticum on native macroalgae, but at different times in its life history. Effects of space should be important in early spring when macroalgae are recruiting and substrate for attachment is limiting, with effects of light becoming important when S. muticum’s annual frond is fully developed. To determine if the competitive mechanisms of S. muticum (light/space) vary through time, we ran the experiment over two consecutive growing seasons (2006, 2007), sampling monthly (May-August).  Data collection This experiment was based at the Bamfield Marine Sciences Centre (herein BMSC). Macroalgal community composition was sampled monthly between May and August of 2006 and 2007, over three consecutive low tides on each sampling event. The richness, abundance, and percent cover of native macroalgae in each plot were recorded. Percent cover for each macroalgal species was estimated on each sampling using a 1 x 1 metre quadrat sub-divided into 20 x 20 cm sections (following Dethier et al. 1993). Macroalgal communities can be structurally complex as different species comprise different layers, which overgrow each other. We measured the percent canopy and substrate cover of native species in each plot. All plots received similar levels of disturbance when the treatments were imposed because most initially constrained >60% exotic cover. Some amount of S. muticum was therefore removed from even the highest cover treatment.  45  Sargassum muticum density treatments were maintained after each monthly sampling. One of each of the light/space treatment plots were excluded from sampling in 2007 following disturbance by a bear. To investigate the effects of S. muticum on macroalgal community productivity, we quantified the total biomass of macroalgae for the different treatments. To avoid destructive sampling that could confound the effect of our treatments, biomass was taken only after the last sampling (August 2007). All macroalgae from each plot were removed, sorted into native species and S. muticum, thoroughly cleaned to remove all sediment, stones, and epifauna, and separately bagged. Macroalgae were spun in a salad spinner to remove excess moisture, dried at 60˚C for 48 hours, and the dry mass recorded. These values of community biomass are likely lower than the rest of the experiment, as many native annual species (i.e. Mazzaella splendens, Halosaccion glandiforme, Leathesia difformis, and Acrosiphonia arcta) senesce at the end of August.  Data analysis We tested effects of the S. muticum density and space/light competition treatments, the effect of time (month), and substrate type on the richness, cover, and diversity of native species using repeated-measures analysis of variance (RM ANOVA). We analyzed diversity, calculated as the Shannon-Weiner Index, to account for variation in abundance and the presence of rare species. Effects of the S. muticum density treatments on native species richness varied with time. We therefore regressed native richness on S. muticum  46  cover for each sampling date separately, including second order terms to test for nonlinear patterns. Because we found unimodal relationships between richness and S. muticum density, we tested if low densities of S. muticum promoted native richness by analysing differences between the 0 and 20% density treatments (excluding the 40 and 60% cover treatments) using RM ANOVA. The impact of the experimental treatments on the aggregate cover of the three functional form groups (canopy, understory, basal) was explored using separate RM ANOVAs. Finally, to test the response of individual species to different densities of S. muticum, we regressed the time averaged cover of the 20 native species most strongly affected by S. muticum against the density treatments. The relationship between the S. muticum density treatments and biomass of native macroalgae and exotic S. muticum was analyzed with univariate ANOVA, and between treatment differences explored with Tukey’s HSD. The assumption of normality was tested using a Shapiro-Wilk test. Richness, diversity, and biomass data were square root transformed, and native cover (individual species and functional groups) were arcsine square root transformed to conform to the assumptions of ANOVA. All analyses were performed in JMP 4.0.4 (SAS Inst.).  RESULTS Native species richness We found non-linear, density-dependent effects of S. muticum on native macroalgal richness (F3,23 = 17.01, P <0.0001, Fig. 3.1). High cover of S. muticum reduced native  47  richness, which declined between 20 and 60% exotic cover (Fig. 3.1). In contrast low cover of the exotic promoted higher native richness (RM ANOVA with only the 0 and 20% S. muticum cover treatments, F1,21 = 15.28, P = 0.002). These treatment effects varied with time (RM ANOVA density treatment x time, F21,49 = 1.75, P = 0.05, Fig. 3.1, Table 3.2A). We found no difference between the effects of the light/space competition treatments on native species richness (F1, 23 = 2.19, P = 0.15), which suggests that effects of S. muticum through light monopolization are more important than space utilization, although space utilization became important at the end of the experiment (Fig. 3.1). The light/space competition treatment effects varied with time (RM ANOVA light/space treatment x time, F7,17 = 2.32, P = 0.05, Table 3.2A). We found no effect of substrate type on native species richness (F2, 23 = 1.09, P = 0.35).  48  Light/space competition 2006 16 14  2007 16  2  R = 0.26  May  Light competition 2006  R = 0.48  14  2007 16  16  2  14  14  12  12  12  10  10  10  8  8  8  6  6  6  12 10 8 6 4  0  Native species richness (S)  16  20  40  60  2  June  R = 0.51  0  20  16  40  0  60  2  R = 0.60  20  40  60  16  16  14  14  12  12  12  12  10  10  10  10  8  8  8  8  6  B  14  6 0  16 14  20  40  60  R2 = 0.38  July  14  20  16  40  60  R2 = 0.34  14  0  20  40  60  16  16  14  14  12  12  12  12  10  10  10  10  8  8  8  8 6  16 14  6  6 0  20  40  60  0  20  16  Aug  40  2  R = 0.48  14  20  40  16  16  14  14  12  12  12  10  10  10  10  8  8  8  6  6  R2 = 0.57 0  20  40  60  0  20  40  60  60  0  20  40  60  R2 = 0.31  0  60  12  6  40  6  0  60  20  6  6  0  0  20  40  60  2  R = 0.25  8  R2 = 0.26 0  20  6 40  60  0  20  40  60  Exotic Sargassum muticum (%)  Figure 3.1 Effects of Sargassum muticum density treatments on native species richness through light/space competition(holdfast and frond removal) and light competition (frond removal, holdfast intact) treatments, for each month (May-August) for the 2006 and 2007 growing seasons. For the light/space treatments, 2006 plots n=24; 2007 plots n=23.  49  Table 3.2 Repeated-measures ANOVA test criterion for the effects of Sargassum muticum on (A) native species richness, (B) native diversity, (C) native cover, and (D) canopy (E) understory, and (F) basal functional group cover. Within subjects results are Wilks' Lambda conservative estimates.  (A) Native richness  (B) Native diversity  (C) Native cover  Between subjects Substrate Density Light/Space Substrate x density Substrate x Light/Space Density x Light/Space Substrate x Density x Light/Space Within subjects Time Time x Substrate Time x Density Time x Light/Space Time x Substrate x Density Time x Substrate x Light/Space Time x Density x Light/Space Time x Substrate x Density x Light/Space  DF 2 3 1 6 2 3  F 1.09 17.01 2.19 0.92 0.36 0.97  P 0.35 <0.0001 0.15 0.50 0.70 0.42  DF 2 3 1 6 2 3  F 0.09 5.55 0.13 0.75 0.12 0.81  P 0.92 0.005 0.73 0.62 0.89 0.5  DF 2 3 1 6 2 3  F 0.99 114.4 1.40 0.52 1.28 0.73  P 0.39 <0.0001 0.25 0.79 0.3 0.55  6  0.40  0.87  6  0.43  0.85  6  1.12  0.38  7 14 21 7 42  113.05 1.65 1.75 2.32 1.53  <0.0001 0.11 0.05 0.05 0.05  7 14 21 7 42  16.15 1.36 0.8 11.38 0.88  <0.0001 0.23 0.71 <0.0001 0.67  7 14 21 7 42  12.52 1.35 2.63 1.22 0.99  <0.0001 0.23 0.003 0.34 0.49  14 21  0.48 1.26  0.93 0.25  14 21  1.08 1.19  0.41 0.3  14 21  2.15 0.99  0.03 0.49  42  1.08  0.38  42  0.76  0.84  42  1.04  0.44  Between subjects Substrate Density Light/Space Substrate x Density Substrate x Light/Space Density x Light/Space Substrate x Density x Light/Space Within subjects Time Time x Substrate Time x Density Time x Light/Space Time x Substrate x Density Time x Substrate x Light/Space Time x Density x Light/Space Time x Substrate x Density x Light/Space  DF 2 3 1 6 2 3  F 0.44 0.17 0.59 0.39 0.26 0.12  P 0.65 0.91 0.45 0.88 0.77 0.95  DF 2 3 1 6 2 3  F 0.22 27.36 3.93 2.05 0.53 1.43  P 0.8 <0.0001 0.06 0.1 0.6 0.26  DF 2 3 1 6 2 3  F 1.88 2.75 1.94 1.6 2.89 1.21  P 0.18 0.07 0.18 0.19 0.08 0.33  6  0.59  0.73  6  0.7  0.65  6  5.72  0.001  7 14 21 7 42  1.92 1.4 1.29 0.86 0.7  0.13 0.21 0.23 0.56 0.89  7 14 21 7 42  15.5 1.04 1.49 3.92 0.99  <0.0001 0.44 0.13 0.01 0.5  7 14 21 7 42  0.79 1.96 0.6 1.04 1.99  0.59 0.06 0.89 0.44 0.004  14 21  1.09 0.45  0.4 0.98  14 21  2.91 1.36  0.01 0.18  14 21  3.21 1.22  0.003 0.28  42  0.65  0.94  42  1.2  0.24  42  1.69  0.02  (D) Canopy cover  (E) Understory cover  (F) Basal cover  50  Community diversity Similarly to native richness, the S. muticum density treatments had strong effects on native diversity (F3,23 = 5.55, P = 0.005), but these treatment effects did not differ with time (RM ANOVA time x density treatments, F21,49 = 0.80, P = 0.71, Table 3.2B). Again, we found no main effects of the light/space treatments (F1,23 = 0.13, P = 0.73), but the treatment effects differed with time (F7,17 = 11.38, P <0.0001). We found no main or interactive effects of substrate type on native diversity (F2,23 = 0.09, P = 0.92, Table 3.2B).  Native community cover We found negative effects of the S. muticum density treatments on total native macroalgal cover (F3,23 = 114.4, P < 0.0001, Fig. 3.2). These treatment effects varied with time (RM ANOVA density treatment x time, F21,49 = 2.63, P = 0.03, Table 3.2B). We found no difference between the S. muticum light/space treatments (F1,23 = 1.40, P = 0.25), or of substrate type on native cover (RM ANOVA, F2,23 = 0.99, P = 0.39, Table 3.2C). Effects of S. muticum on aggregate cover of the three morphological functional groups varied. S. muticum had no detectable effects on cover of basal (F3 = 2.75, P = 0.07) or canopy (F3 = 0.17, P = 0.91) functional groups (Table 3.2D, F). There were weak negative effects of the S. muticum density treatments on cover of the understory  51  functional group (F3 = 27.36, P < 0.0001, Table 3.2E), suggesting competitive effects of S. muticum are strongest on these species.  2006 100  2007 R2 = 0.73  100  80  80  60  60  40  40  20  R2 = 0.67  20  May  Total native macroalgal cover (%)  0  100  0 0  20  40  60  100  0  20  40  2  R = 0.89  R = 0.72  80  80  60  60  40  40  20  60  2  20  June 0 100  0 0  20  40  60  100  0  20  40  R2 = 0.95 80  80  60  60  40  40  20  60  R2 = 0.82  20  July 0 100  0 0  20  40  60  100  0  20  2  60  2  R = 0.89  R = 0.85  80  80  60  60  40  40  20  40  20  Aug 0  0 0  20  40  60  0  20  40  60  Exotic Sargassum muticum (%) Figure 3.2 Relationship between Sargassum muticum density and total native macroalgal cover (%) for each month (May-August) for the 2006 and 2007 growing seasons. Data points represent the time averaged cover for each replicate plot (n=48). The light/space treatments were pooled as they were not significantly different (Table 3.2).  52  Antithamnionella sp.  Ceramium sp.  6  Chondracanthus exasperatus  R2 = 0.48  9  16  R2 = 0.10  Costaria costata  2.5  R2 = 0.16  2  12 4  1.5  6  8 2  1  3  0.5  4 0  0  0 0  20  40  Fucus distichus  6  0 0  60  20  40  60  0  20  0  60  Gracilaria sp.  Gastroclonium subarticulum  R2 = 0.22  40  6  R2 = 0.11  12  20  40  60  Halosaccion glandiforme 4  3  4  4 8  2 2  2 4  0  1 0  0  0  20  40  60  0  Macroalgal cover (%)  Hildenbrandia rubra  R2 = 0.27  2  20  40  0 0  60  20  Leathesia difformis 10  0  60  Mazzaella splendens  R2 = 0.38  8  40  6  R2 = 0.15  20  40  60  Neorhodomela larix 16  R2 = 0.24  12 4  1.5  6 8  1  4  0.5  2  0  0 0  20  40  2  4 0  0  60  Odonthalia floccosa  20  0  60  0  20  R2 = 0.24  40  0  60  Phyllospadix sp.  Osmundea spectabilis 8  6  40  16  6  12  4  8  2  4  20  40  60  Prionitis lanceolata 10  R2 = 0.39  R2 = 0.16  8  4  6 4  2  0  0 0  20  40  60  6  R2 = 0.08  4  0  0  0  Pseudolithophyllum sp.  2  20  40  0  60  20  Scytosiphon lomentaria  R2 = 0.15  12  40  60  0  Ulva fenestrata  20  40  60  Ulva intestinalis 6  R2 = 0.10  15  4  8  10 2  2  4  0  5  0 0  20  40  60  0  0 0  20  40  60  0  20  40  60  0  20  40  60  Exotic Sargassum muticum (%) Figure 3.3 Effects of the exotic density treatments on the average cover of the 20 native species most strongly affected by Sargassum muticum. Data points represent the time averaged cover for each replicate plot (n=48). The light/space treatments were pooled as they were not significantly different (Table 3.2). Analysis was by linear least squares regression.  53  Only one species showed increased cover in response to increased exotic density, the marine angiosperm Phyllospadix scouleri (R2 = 0.39, P < 0.0001, Fig. 3.3) suggesting it does better in the presence of the canopy forming exotic. Increased S. muticum density corresponded with decreased cover of most native macroalgae. Fourteen of the 20 natives most affected by S. muticum showed negative relationships with the exotic density treatments (Fig. 3.3). These results suggest the effect of S. muticum on individual native macroalgae are primarily negative, but vary with species identity.  Biomass Sargassum muticum suppressed total native community biomass; native biomass increased as S. muticum density decreased (Fig. 3.4). No differences in native biomass were found between the S. muticum light/space treatments (ANOVA, F1,44 = 0.01, P = 0.93). The biomass produced by S. muticum failed to compensate for the loss of native biomass due to competition with the exotic, suggesting a negative effect of S. muticum on community productivity (Fig. 3.4).  54  Light/space competition  Dry biomass (g)  250  a  200  Invasive ab  Native  bc  150  c  250 200  ac  c  150  100  100  50  50  0  Light competition a ab  0 0  20  40  60  0  20  40  60  Exotic Sargassum muticum (%)  Figure 3.4 Relationship between Sargassum muticum density and native and exotic (S. muticum) macroalgal biomass (g dried mass) for the light/space (holdfast and frond removal) and light (frond removal, holdfast intact) competition treatments. Data are the average of 46 manipulated plots, from August 2007. Error bars are SE, plus for exotic biomass, negative for native biomass.  DISCUSSION Density dependent effects of Sargassum muticum Along with habitat modification, invasion by exotic species is one of the main threats to the conservation of biodiversity (Wilcove et al. 1998). Despite a growing number of well-documented invasions in marine systems (Cohen and Carlton 1998), little experimental work has been done to separate the mechanisms underlying the success and failure of marine invasions (but see Byers 2000, Britton-Simmons 2004, 2006). Studies from the marine environment that examine impacts of exotics on natives have reported both positive (Wonham et al. 2005), and negative (Stachowicz et al. 2002) effects,  55  resulting in no generalizations emerging about the impacts of exotics on natives. In a recent review of the impacts of macroalgal invasions, Schaffelke and Hewitt (2007) reviewed 69 studies that presented data on the impacts of exotic macroalgae. None of the studies examined reported positive effects of the exotic macroalga on the recipient community. Similarly, Williams and Smith (2007) reviewed 407 global seaweed introduction events, of which the ecological effects of introduced seaweeds were studied in only 6% of the species, with mostly negative effects or changes to the native biota reported. Our study demonstrates that at high cover, S. muticum negatively effects native species richness through light competition mediated through the shading of smaller, underlying macroalgae by large fronds. Positive effects of S. muticum at low cover appear to be overwhelmed by negative effects at higher cover, leading to negative associations between native diversity and exotic cover, a pattern reported from unmanipulated low intertidal plots (White and Shurin 2007). It is possible that the impacts of invaders across systems are not generalizable, perhaps due to important differences between systems or organisms, but that extractable trends of impacts within systems, such as intertidal macroalgae, may exist.  Mechanisms of competition and facilitation Sargassum muticum’s effect on native taxa arose mainly through light competition, although space utilization became important at the end of the experiment. The S. muticum exerts shading effects on smaller macroalgae by forming dense stands (Deysher and  56  Norton 1982). In dense stands, S. muticum individuals grow taller and thinner (Arenas et al. 2002). This plastic response of S. muticum at high densities enables the establishment of dense populations with high persistence. This may explain why S. muticum has strong impacts on native macroalgae through light competition at high densities, but not at low densities. The importance of shading by S. muticum in interspecific competition with native macroalgae has been demonstrated previously. Ambrose and Nelson (1982) showed that S. muticum negatively affected kelp recruitment though shading effects, and Britton-Simmons (2004) demonstrated shading by S. muticum slows kelp growth rates. Effects of the light/space treatments varied through time (Fig. 3.1, Table 3.2A, B). Britton-Simmons (2004) showed the influence of S. muticum on light intensity varies over time as the fronds grow, mature, and senesce. The light/space treatment effects varied with time, suggesting that competition for space may play an important role at certain times. The perennial holdfast may be important for preventing the reestablishment of annual macroalgal species and maintaining population persistence of S. muticum. The importance of space preemption has been demonstrated in both marine invertebrate (Stachowicz et al. 2002, Dunstan and Johnson 2004) and marine macroalgal communities (Britton-Simmons 2006). However, we found that competition for light was more important than for space in that the removal of the annual frond produced results similar to those of removing the entire macroalga. The mechanism for the positive effects of S. muticum on native richness at low densities is unclear. The response of individual macroalgal species to the S. muticum density treatments varied qualitatively, from neutral (no effect) to negative effects (Fig.  57  3.3). This suggests some native species benefited from a modest amount of S. muticum cover (perhaps due to reduced water motion from settling near the holdfast, or maybe S. muticum facilitates obligate understory species that survive poorly without canopy species), while other species suffered competitive effects. While habitat modification by S. muticum to the benefit of native species is possible, Britton-Simmons (2004) found no effect of S. muticum on water flow, sedimentation, or nutrient availability in nearby Washington State. One possibility is that S. muticum affords protection to native macroalgae through the “associational plant refuge” (Pfister and Hay 1988), which suggests that the presence of unpalatable macroalgae alters the feeding behaviour of herbivores on palatable macroalgae when both types are present together. Further experimentation is required to determine the mechanisms by which S. muticum promotes native macroalgal richness at low densities. Interactions between exotic and native species have been shown to change in response to a variety of factors. Reush and Williams (1998) reported neutral, negative, and positive effects of the exotic marine mussel Musculista senhousia on the native eelgrass Zostera marina at different times as the abundance of the invader changed. The effects of S. muticum in the present study are both density and time dependent (Fig. 3.1). It is possible that mechanisms other than light and space competition are influential in this system. Evidence suggests that synergistic interactions with other stressors play an important role in the establishment and spread of marine aliens species, and, hence, for any negative impacts (Occhipinti-Ambrogi and Savini 2003). Examples of other potentially influential mechanisms in this system may include reduction in hydrodynamic  58  flow, disturbance through scouring of smaller macroalgae by frond sweeping of larger macroalgae, and sloughing of epithelial layers by crusts (Johnson and Mann 1986, Keats et al. 1997).  Community impacts Functional form groups have been widely invoked in studies of macroalgal community dynamics based on the concept that similar morphologies confer functional equivalency (Littler and Littler 1980, Steneck and Watling 1982, Steneck and Dethier 1994). Based on this idea, the effects of S. muticum should be similar on species of similar morphology. The three functional form groups we identified were differentially affected by S. muticum. Understory functional group species (smaller fleshy macroalgae) were negatively affected by S. muticum (Table 3.2E). These species occupy little substrate, but may be less shade tolerant than encrusting and calcified basal macroalgal forms (Table 3.1), so more impacted by shading effects. Canopy species are functionally similar to S. muticum and therefore have similar effects on light and space, which may lead us to expect them to be most inhibited by competition with S. muticum. While there appears to be no effect of S. muticum on canopy species (Table 3.2D), Ambrose and Nelson (1982) showed recruitment of the canopy forming giant kelp (Macrocystis pyrifera) was inhibited by S. muticum through shading effects, suggesting effects of the exotic on native species may vary with life history stage of the native. Variable impacts of S. muticum on different functional groups was also reported by Olabarria et al. (2009), who found cover of understory filamentous and foliose macroalgae were negatively  59  affected by cover of the exotic, but not the entire macroalgal assemblage. However, there was substantial variation in the effects of S. muticum on macroalgae within functional form groups (Fig. 3.3), suggesting that our definition of functional form groups indicates little about the performance of the constituent species or their ecological interactions, a finding recently reported by Bates (2009). Models of community organization predict that removing a competitively dominant species from a system often results in an increase in species richness or diversity, and changes in the relative abundance of understory species (Dayton 1975, Lubchenco 1978, Clark et al. 2004). Our results suggest that S. muticum is a competitive dominant in this system. Native biomass was higher in plots without the exotic (>40% average difference in exotic biomass between 0 and 60% treatments, Fig. 3.4), suggesting S. muticum displaces native biomass. This suggests that in addition to reducing the diversity and cover of native macroalgal assemblages, S. muticum also reduces the total biomass. We measured community biomass from a one time sampling in August 2007. The effects of S. muticum on native biomass are likely strongest at this time as individuals have the highest biomass in July-August (Wernberg et al. 2001).  Implications of invasion by Sargassum muticum Community studies show that invasion by S. muticum has primarily negative effects on native macroalgal biodiversity once it becomes abundant in its exotic range (Kraan 2007). S. muticum invaded Denmark in 1984 and shifted macroalgal community structure toward lower diversity and evenness (Stæhr et al. 2000, Wernberg et al. 2001). In  60  Portugal S. muticum dominates rock pools and often excludes all other macroalgal species. Halidrys siliquosa, a previously common species is now conspicuously absent from rock pools altogether (Engelen et al. 2003). Neutral effects of S. muticum on native diversity have been reported from Spain (Sánchez and Fernández 2005) and California (Wilson 2001). Our study is the first to demonstrate positive effects of S. muticum on benthic macroalgal richness. Positive effects of S. muticum have been reported from two other studies. S. muticum increased richness of filamentous macroalgae in Denmark by providing habitat for filamentous epiphytes (Thomsen et al. 2006), and increased richness of mobile marine invertebrates such as amphipods and littorine snails by increasing available habitat through its distinct, complex structure (Giver 1999). These results highlight the complexity of the interactions between S. muticum and the recipient community. While S. muticum can facilitate native macroalgal richness at low cover, it largely excludes other macroalgae through competition, but can enhance diversity and abundance of epiphytes and mobile invertebrates that colonize its fronds. The reciprocal interaction between exotic S. muticum and native macroalgal diversity has shown both competitive and facilitative effects operating in both directions. We showed native diversity effects on invasion varied from positive to negative with life history stage of the invader (White and Shurin 2007). More diverse native communities favoured greater settlement of S. muticum germlings, but subsequent growth was slower than when fewer native species were present. The current study demonstrates the effects of S. muticum on native taxa are both density and time dependent. Positive effects became apparent at low density of the invader, however at densities seen in our sites, the  61  effects on native richness are largely negative. The competitive suppression came about largely through competition for light. Effects of space only became important at the end of the experiment. Our results indicate that the negative association between S. muticum cover and native richness results both from invasion resistance in diverse native communities, and by suppression of native taxa by competition with the invader.  62  LITERATURE CITED Airoldi L (2000) Effects of disturbance, life-histories and overgrowth on coexistence of algal crusts and turf. Ecology 81: 798-814. Ambrose RF and Nelsson BV (1982) Inhibition of giant kelp recruitment by an introduced brown algae. Bot. Mar. 25: 265-267. Arenas F, Sánchez I, Hawkins SJ and Jenkins SR (2006) The invasibility of marine algal assemblages: role of functional diversity and identity. Ecology 87: 2851-2861. Arenas F, Viejo RM, Fernández C (2002) Density-dependent regulation in an invasive seaweed: Responses at plant and modular levels. J. Ecol. 90: 820-829. Bates CR (2009) Host taxonomic relatedness and functional-group affiliation as predictors of seaweed-invertebrate epifaunal associations. Mar. Ecol. Prog. Ser. 387: 125-136. Boudouresque CF, Verlaque M (2002) Biological pollution in the Mediterranean Sea: invasive versus introduced macrophytes. Mar. Pollut. Bull. 44: 32-38. Britton-Simmons KH (2006) Functional group diversity, resource preemption and the genesis of invasion resistance in a community of marine algae. Oikos 113: 395-401. Britton-Simmons KH (2004) Direct and indirect effects of the introduced alga Sargassum muticum on benthic, subtidal communities of Washington State, USA. Mar. Ecol. Prog. Ser. 277: 61-78. Byers JE (2000) Competition between two estuarine snails: implications for invasions of exotic species. Ecology 81: 1225-1239. Cohen AN and Carlton JT (1998) Accelerating invasion rate in a highly invaded estuary. Science 279: 555-557. Clark RP, Edwards MS, Foster MS (2004) Effects of shade from multiple kelp canopies on an understory algal assemblage. Mar. Ecol. Prog. Ser. 267: 107-119. Dayton PK (1975) Experimental evaluation of ecological dominance in a rocky intertidal community. Ecol. Monographs 45: 137-159. Dethier MN, Graham ES, Cohen S and Tear LM (1993) Visual versus random-point percent cover estimations: ‘‘objective’’ is not always better. Mar. Ecol. Prog. Ser. 96: 93-100.  63  DeWreede RE and Vandermeulen H (1988) Lithothrix aspergillum (Rhodophyta): regrowth and interaction with Sargassum muticum (Phaeophyta) and Neorhodomela larix (Rhodophyta). Phycologia 27: 469-476. Deysher L and Norton TA (1982) Dispersal and colonization in Sargassum muticum (Yendo) Fensholt. J. Exp. Mar. Biol. Ecol. 56: 179-1960. Dunstan PK and Johnson CR (2004) Invasion rates increase with species richness in a marine epibenthic community by two mechanisms. Oecologia 138: 285-292. Engelen A, Santos R, Alves C (2003) Demography of the alien species Sargassum muticum at its southern European distribution limit. Third European Phycological Congress, Queens University, Belfast, 21-26 July 2003. Program and Book of abstracts, M8-2, p 48. Gabrielson PW, Widdowson TB, Lindstrom SC (2006) Keys to the seaweeds and seagrasses of southeast Alaska, British Columbia, Washington, and Oregon. Phycological Contribution No. 7. University of British Columbia, Vancouver, BC. Giver KJ (1999) Effects of the invasive seaweed Sargassum muticum on native marine communities in northern Puget Sound, Washington. Master of Science Thesis, Western Washington University, Bellingham, Washington. 93 pp. Washington State Dept. Ecology (No. 00-06-010), Padilla Bay National Estuarine Research Reserve Reprint Series No. 30. Johnson CR and Mann KH (1986) The crustose coralline alga, Phymatolithon foslie, inhibits the overgrowth of seaweeds without relying on herbivores. J. Exp. Mar. Biol. Ecol. 96: 127-146. Keats DW, Knight MA and Pueschel CM (1997) Antifouling effects of epithelial shedding in three crustose coralline algae (Rhodophyta: Coralinales) on a coral reef. J. Exp. Mar. Biol. Ecol. 213: 281-293. Klein J, Ruitton S, Verlaque M, Boudouresque CF (2005) Species introductions, diversity and disturbances in marine macrophyte assemblages of the northwestern Mediterranean Sea. Mar. Ecol. Prog. Ser. 290: 79-88. Kraan S (2007) Sargassum muticum (Yendo) Fensholt in Ireland: an invasive species on the move. J. Appl. Phycol. 20: 375-382. Levine JM (2000) Species diversity and biological invasions: relating local process to community pattern. Science 288: 852-854.  64  Littler MM and Littler DS (1980) The evolution of thallus form and survival strategies in benthic marine macroalgae: field and laboratory tests of a functional form model. Am. Nat. 116: 25-44. Lubchenco J (1983) Littorina and Fucus: effects of herbivores, substratum, heterogeneity, and plant escapes during succession. Ecology 64: 1116-1123. Lubchenco J (1978) Plant species diversity in a marine intertidal community: importance of herbivore food preference and algal competitive ability. Am. Nat. 112: 23-39. Mineur K, Johnson MP, Maggs CA (2008) Non-indigenous marine macroalgae in native communities: a case study in the British Isles. J. Mar. Biol. Ass. UK. 88: 693-698. Norton TA (1976) Why is Sargassum muticum so invasive? Br. Phycol. J. 11: 197. Nyberg C (2007) Introduced macroalgae and habitat modifiers - their ecological role and significant attributes. Ph.D Thesis, Dept. Marine Ecology, Goteborg University, Sweden. Occhipinti Ambrogi A, Savini D (2003) Biological invasions as a component of global change in stressed marine ecosystems. Mar. Pollut. Bull. 46: 542-551. Olabarria C, Rodil IF, Incera M, Troncoso JS (2009) Limited impact of Sargassum muticum on native algal assemblages from rocky intertidal shores. Mar. Env. Res. 67: 153-158. Pfister CA, Hay ME (1988) Associational plant refuges: convergent patterns in marine and terrestrial communities result from differing mechanisms. Oecologia 77: 118129. Pfister CA and Van Alstyne KL (2003) An experimental assessment of the effects of nutrient enhancement on the intertidal kelp Hedophyllum sessile (Laminariales, Phaeophyceae). J. Phyc. 39: 285-290. Rejmanek M (1989) Invasibility of plant communities. In: Biological Invasions: a global perspective. Eds: Drake JA, Mooney HA, di Castri F, Groves RH, Kruger FJ, Rejmanek, Williamson M. Wiley, Chichester, England. Ribera MA and Boudouresque CF (1995) Introduced marine plants, with special reference to macroalgae: mechanisms and impact. Prog. Phycol. Res. 11: 187-268. Reush TBH and Williams SL (1998) Variable responses of native eelgrass Zostera marina to a non-indigenous bivalve Musculista senhousia. Oecologia 113: 428-441.  65  Sánchez I and Fernández C (2005) Impact of the invasive seaweed Sargassum muticum (Phaeophyta) on an intertidal macroalgal assemblage. J. Phycol. 41: 923-930. Scagel RF (1956) Introduction of a Japanese alga, Sargassum muticum, into the northeast Pacific. Fish. Res. Pap. Wash. Dept. Fish 1: 49-59. Schaffelke B and Hewitt CL (2007) Impacts of introduced seaweeds. Bot. Mar. 50: 397417. Stachowicz JJ, Fried H, Osman RW, Whitlatch RB (2002). Biodiversity, invasion resistance, and marine ecosystem function: reconciling pattern and process. Ecology 83: 2575-2590. Stachowicz JJ, Graham M, Bracken MES and Szoboszlai AI (2008) Diversity enhances cover and stability of seaweed assemblages; the role of heterogeneity and time. Ecology 89: 3008-3019. Stæhr PA, Pederson MF, Thomsen MS, Wernberg T, and Krause-Jensen D (2000) Invasion of Sargassum muticum in Limfjorden (Denmark) and its possible impact on the indigenous macroalgal community. Marine Ecology Progress Series 207: 7988. Steneck RS and Dethier MN (1994) A functional group approach to the structure of algaldominated communities. Oikos 69: 476-498. Steneck RS and Watling L (1982) Feeding capabilities and limitations of herbivorous molluscs: a functional group approach. Marine Biology 68: 299-319. Thom RM and Hallum L (1990) Long-term changes in the areal extent of tidal marshes, eelgrass meadows and kelp forests of Puget Sound. Final report to Office of Puget Sound, Region 10, U.S. EPA. University of Washington Fisheries Research Institute, Seattle, WA. Thomsen MS, Wernberg T, Stæhr PA, Pedersen MF (2006) Spatio-temporal distribution patterns of the invasive macroalga Sargassum muticum within a Danish Sargassumbed. Helgol. Mar. Res. 60: 50-58. Trowbridge CD (1995) Establishment of the green alga Codium fragile ssp. tomentosoides on New Zealand rocky shores: current distribution and invertebrate grazers. J. Ecol. 83: 949-965.  66  Wernberg T, Thomsen MS, Stæhr PA and Pedersen MF (2001) Comparative phenology of Sargassum muticum and Halidrys siliquosa (Phaeophyceae: Fucales) in Limfjorden, Denmark. Bot. Mar. 44: 31-39. Wasson K, Fenn K, Pearse JS (2005) Habitat differences in marine invasions of Central California. Biol. Invas. 7: 935-948. White LF and Shurin JB (2007) Diversity effects on invasion vary with life history stage in marine macroalgae. Oikos 116: 1193-1203. Wilcove DS, Rothstein D, Dubow J, Phillips A and Losos E (1998) Quantifying threats to imperiled species in the United States. Bioscience 48: 607-615. Williams SL and Smith JE (2007) A global review of the distribution, taxonomy & ecological impacts of introduced seaweeds. Annual Review of Ecology Evolution and Systematics 38: 327-359. Wilson CM (2001) Is Sargassum muticum a benign invader of tidepools on the Pacific coast of North America? MSc Thesis, California State University, Monterey Bay, California. Wonham MJ, O’Connor M, Harley CD (2005) Positive effects of a dominant invader on introduced and native mudflat species. Mar. Ecol. Prog. Ser. 289: 109-116. Wooton JT (1991) Direct and indirect effects of nutrients in intertidal community structure: variable consequences of seabird guano. J. Exp. Mar. Biol. Ecol. 151: 139-153.  67  CHAPTER FOUR COMPARING HERBIVORY AND SECONDARY COMPOUNDS IN INVASIVE SARGASSUM MUTICUM WITH NATIVE BROWN ALGAE OF BRITISH COLUMBIA  INTRODUCTION  Invasive species can competitively displace native species, especially when natural predators, disease, or other suppressants that might regulate the exotic population are absent (Byers 1999). Understanding the mechanisms that promote successful invasions is central to our understanding of their interactions with recipient communities. Reports that large ecosystems such as the Mediterranean Sea are losing their biological distinctiveness puts more urgency on understanding what factors promote invaders and their impacts on native species (Mollo et al. 2008). One potential mechanism for the success of introduced species is the EnemyRelease Hypothesis (ERH, Keane and Crawley 2002). The ERH states that the successful colonization and proliferation by an introduced species is a result of release from population regulation by their co-evolved natural enemies that occur in their native range. No specialist grazers of the introduced plant are present in the invaded range, and generalist grazers native to the new range will prefer native over introduced individuals (Prider et al. 2008). A potential mechanism for the reduced grazing of introduced plants 1  A version of this chapter will be submitted for publication. White LF and Nienhuis SB (2010) Comparing herbivory and secondary compounds in invasive Sargasum muticum with native brown algae of British Columbia.  68  compared to native plants in the invaded range is differences in the levels of chemical defenses expressed. As a result, the introduced plants require less energy for herbivore defense, meaning more is available for growth and reproduction, allowing for an increase in competitive ability compared to native species (Evolution of Increased Competitive Ability hypothesis (EICA) sensu Blossey and Notzold 1995). Biogeographical studies of enemy-release compare native and introduced populations of the same species. Community studies examine native and non-native species co-occurring within the same community, and indicate the relative impact of enemies on coexisting native and nonnative species (Colautti et al. 2004). Biogeographical studies examining the presence or effects of parasites, pathogens or predators on native and introduced populations of a single host species show enemy impact was higher on native than non-native populations, supporting the predictions of the ERH across systems; aquatic invertebrates (Torchin et al. 2001), agricultural plants (Wolfe 2002), salmonid fishes (Poulin and Mouillot 2003), and terrestrial shrubs (DeWalt et al. 2004). Results from community studies are mixed. Some find support for the ERH demonstrating greater enemy impact on species native to the invaded range than introduced species (Schierenbeck et al. 1994, Lesica and Miles 1999, Goergen and Daehler 2001, Siemann and Rogers 2003b, Prider et al. 2008), some find no support (Radho-Toly et al. 2001, Novotny et al. 2003), and others find results opposite to the predictions of the ERH in that exotics are more impacted than natives in the invaded region (Clay 1995, Agrawal and Kotanen 2003).  69  The functional role of polyphenolic compounds in marine macroalgae Among macroalgal divisions, the brown macroalgae (class Phaeophyceae) frequently contain the highest levels of polyphenolic compounds as a single structural class, the phlorotannins, i.e. polymers of phloroglucinol (1,3,5-trihydroxybenzene, Connan et al. 2006). Polyphenolic compounds are ubiquitous metabolic constituents of marine brown macroalgae (Steinberg 1989), which can possess primary metabolic functions as cell wall components (Schoenwaelder and Clayton 1999), and whose metabolites often display a putative secondary function in macroalgal defense. The mechanisms by which polyphenolics achieve defense remain unclear, but include preventing bacterial infections (Conover and Sieburth 1964, Sieburth 1968), endophytes (Geiselman 1974), and fouling (Schmitt et al. 1995), protection from UV damage (Swanson and Druehl 2002), serving as allelopathic agents (McLachlan and Craigie 1966), and have been strongly correlated with deterrence of herbivores, which may be manifested by their ability to bind to digestive proteins, reducing assimilation efficiency (Rosenthal and Janzen 1979). The functioning of macroalgal metabolites as herbivore deterrents is reviewed by Pavia and Toth (2008). Marine herbivores deploy a range of behavioural and physiological responses to polyphenolic compounds in brown macroalgae that varies over biogeographic scales (see Steinberg 1989 for review). Temperate Australian herbivores are not typically deterred by polyphenolic rich macroalgae (Steinberg and van Altena 1992), and results from the tropical Indo Pacific are mixed (Van Alstyne and Paul 1990, Steinberg et al. 1991). However, a number of north temperate marine herbivores have been shown to be  70  consistently deterred by polyphenolic rich tissue (Geiselman and McConnell 1981, Steinberg 1984, 1985, 1988, Steinberg and van Altena 1992) which can also affect herbivore assimilation efficiencies at ecologically relevant concentrations (Irelan and Horn 1991, Steinberg and van Altena 1992). In experimental tests, macroalgal polyphenolics incorporated into agar disks deterred grazing by gastropod and echinoid herbivores (Geiselman and McConnell 1981, Steinberg 1988). Phloroglucinol deterred grazing by the echinoid Strongylocentrotus purpuratus across a range of concentrations commonly observed in macroalgal tissue (Steinberg 1988). Steinberg (1984, 1985) reported consistent preference by 17 northeastern Pacific herbivores for polyphenolic poor brown macroalgae when compared with polyphenolic rich tissue (see also Barker and Chapman 1990, Denton et al. 1990, Boettcher and Targett 1993). Polyphenolic-rich macroalgal species have also been shown to have higher survival in a number of geographic regions (Estes and Steinberg 1988, Lubchenco 1983, Barker and Chapman 1990), indicating that polyphenolics can function as chemical defences under natural conditions (Steinberg 1992). We used grazing experiments and polyphenolic bioassays to address three comparisons between the introduced marine macroalga Sargassum muticum and three native brown macroalgae (Alaria marginata, Egregia menziesii, and Fucus distichus) in British Columbia, Canada. First we asked whether a suite of seven generalist invertebrate herbivores consume S. muticum more or less than native taxa in a no-choice experiment. Second, we used paired-choice experiments to ask if the same native herbivores exhibit selective grazing on the native or invasive macroalgae. If the ERH explains invasion  71  success of S. muticum in British Columbia, we expect the invader to be less grazed by native herbivores than native macroalgae. Third, we measured the concentrations of polyphenolic compounds in the tissue of the four macroalgal species to determine whether the native and invasive species express different levels of chemical defenses.  MATERIALS AND METHODS Field collections Macroalgae were collected from three sites within Barkley Sound (BS) on the west coast of Vancouver Island, British Columbia (site 1: 48' 51.01" N, 125' 07.07"W; site 2: 48' 52.18" N, 125' 08.11" W; site 3: 48" 51.43' N, 125"06.49'W). Observational surveys of the environmental conditions at each site were first conducted as a suite of factors have been shown to influence polyphenolic content in macroalgae. We attempted to identify collection sites that were similar in aspect, wave exposure and tidal height. Polyphenolics have been shown to vary geographically as a result of differing environmental influences. For example, polyphenolics are photo reactive and therefore affected by the quality and intensity of light, especially by exposure to UV-radiation (Pavia et al. 1997). We selected west-facing sites with no obvious shade (i.e. tree line). To control for effects of emersion time (Martínez 1996), all individuals were collected from the same tidal height, with no rock pool individuals selected. Seasonal variation in polyphenolics has also been reported (Rönnberg and Ruokolahti 1986). To standardize for effects of seasonal or temporal variation, all macroalgae were collected within a three day period over consecutive low tides in June 2007, with macroalgae from each site collected on the same tide. Damaged  72  individuals were not collected as correlations between damage and polyphenolic content have been reported (Van Alstyne 1988, Lüder and Clayton 2004). Salinity (Pedersen 1984), site exposure, and nutrient concentrations (Yates and Peckol 1993, Hemmi et al. 2004) are other potentially influential factors. While we did not measure these factors, we carefully selected collection sites similar in physical and biological characteristics, and tested for differences in polyphenolic content by statistically blocking by site. Macroalgae were subjected to the same collection, handling, and storage conditions at Bamfield Marine Sciences Centre (herein BMSC) in an attempt to control for macroalgal condition, which may influence herbivore preference. We compared the three native brown macroalgal species Alaria marginata (Laminariales), Egregia menziesii (Laminariales), and Fucus distichus (Fucales), with non-native S. muticum (Fucales). All four species co-occurred in abundance at the same tidal height at all three sites. Macroalgal species will hereafter be referred to by genus. Whole individuals for both grazing trials and polyphenolic bioassays were randomly selected along a 75 metre transect within the low intertidal zone at each site. Mobile generalist herbivore species were selected based on field observations of their occurrence on the brown macroalgal species utilized in this study, and collected from sites where they occurred in abundance. These included the kelp crab (Pugettia producta), purple urchin (Strongylocentrotus purpuratus), red turban snail (Astraea gibberosa), kelp isopod (Idotea wosnesenskii), blue-banded hermit crab (Pagurus samuelis), shield limpet (Lottia pelta), and beach-hopper amphipod (Gammarid  73  amphipod). Herbivore species were kept in separate flow-through seawater tables at BMSC with a mixed macroalgal food source for one week prior to use.  Grazing and preference assays We conducted grazing trials to test whether native herbivores preferentially graze native, sympatric macroalgae over non-native S. muticum. Modified from Liszka and Underwood (1990), we first examined grazing by native herbivores on each of the four macroalgal species with no available alternative macroalgal food source (the no-choice experiment). This experiment was followed by preference trials to determine whether herbivores discriminate between macroalgal species in mixture (the choice experiment). Upon collection, mixed macroalgae were supplied to herbivores for a seven day acclimation period. Herbivores were then subjected to no food conditions for four to six days to control for any differences in recent consumption. The no-choice experiment measures the physiological capacity of the herbivore to consume macroalgal tissue, not the behavioural preference for a particular macroalgal species. For the no-choice grazing trials, 24 individuals of each of the seven herbivore species were housed separately in flow-through containers in seawater tables with a single piece of macroalgae. Six replicate individuals of each herbivore species were tested with each of the four macroalgal species (Table 4.1, Treatments 1-4). We ran grazing trials over four time periods to increase replication as we were limited by seawater table space. A new herbivore individual and piece of macroalgae were used at the start of each grazing trial to avoid non-independence. Each grazing trial lasted six  74  days, and the blotted mass of each herbivore and piece of macroalgae were recorded at the beginning and end of each. This allowed us to determine the amount of macroalgae consumed (grams) per herbivore mass as grazers differed considerably in size. We used six day trials to prevent macroalgal quality degrading from tissue necropsy, and to ensure that any decrease in mass was due directly to herbivory and not tissue loss independent from, or subsequent to, herbivore damage. An equal number of macroalgae-only controls were used to assess autogenic changes over the grazing trials, and were randomly paired with treatment containers in the seawater tables. We simulated low tides in the tanks daily to coincide with field conditions. No-choice trails ran June 18th to July 13th 2007. In the absence of choice, herbivores may consume one food type out of necessity, but avoid it in the presence of alternatives (Underwood et al. 2004). To measure relative grazer preferences for the four macroalgal species (i.e discrimination among types when confronted by choice, sensu Liszka and Underwood 1990) herbivore individuals were next presented with paired combinations of two macroalgal species. Eighteen individuals of each of the seven herbivore species were housed separately in flow-through containers in seawater tables with two pieces of macroalgae of known mass. The three possible macroalgal combinations (Table 4.1, Treatments 5-7) were tested for each of the seven herbivore species, with each herbivore-macroalgal combination (and randomly paired macroalgae only controls) replicated six times. Macroalgal pieces were of similar starting mass as electivity can be influenced by availability (Underwood et al. 2004). Choice grazing trials were replicated over time, consistent with the no-choice trials. New organisms were again used in each trial. At the end of each time step, the blotted mass of  75  the macroalgae and herbivore were taken to determine mass changes during the trials. Choice grazing trials ran August 19th to September 13th 2007.  Table 4.1 Treatment codes for macroalgal combinations used in grazing trials.  Treatment  Macroalgal Species  1 2 3 4 5 6 7  Sargassum Fucus Egregia Alaria Sargassum + Fucus Sargassum + Egregia Sargassum + Alaria  Species Code S F E A SF SE SA  Polyphenolic assays We extracted and quantified polyphenolic compounds from the four brown macroalgal species used in the grazing trials to test whether they differed in polyphenolic content. Fifteen replicate adult individuals of each macroalgal species were randomly collected from each of the three sites, cleaned of epiphytes, and individually sealed in labelled freezer bags. These samples were transported on ice to BMSC, where they were immediately stored in the dark at -80ºC for approximately two weeks prior to analysis. We selected non-reproductive, apical macroalgal tissue to standardize our assays and minimize the intra-individual variation in polyphenolic content. We used a modified Folin-Ciocalteu assay for quantifying total polyphenolic content of macroalgae, following the method described by Van Alstyne (1995). This colorimetric assay has been widely  76  used in studies of macroalgal polyphenolic content (Connan et al. 2006, Van Alstyne et al. 1999, Pavia et al. 1999, Van Alstyne et al. 2001, Stiger et al. 2004, Koivikko et al. 2005, Jormalainen and Ramsay 2009). 0.5 g blotted mass of tissue was excised from each individual, homogenized in liquid nitrogen with a mortar and pestle, and extracted in 80% aqueous methanol in darkness at 4ºC for 12 hours. After extraction, samples were agitated and filtered through a Whatman no. 1 filter. Folin-Ciocalteu assays were performed on the filtrate. A second 0.5 g sample of apical tissue from each individual was weighed (blotted mass) and dried at 60ºC for 48 hours to determine the dry:blotted mass ratio of the macroalgae. We measured the blotted mass instead of the frozen mass as the macroalgal species trapped water differently, contributing to variation in mass due to ice accumulation. Absorbance of each solution was read spectrophotometrically at 765 nm (Biochrom Ultrospec 2100 pro) using phloroglucinol (Fisher Scientific) as standard. Total polyphenolic content was expressed as a percent with respect to the dry weight (DW) in grams. Assays were run in triplicate for each extract, such that for each individual the % DW was reported as a mean ± standard error.  Statistical analyses Consumption of each macroalgal species was calculated as [Ti(Cf/Ci)] –Tf, where Ti and Tf are the initial and final macroalgal masses of the treatments and Ci and Cf are the initial and final macroalgal masses of the randomly paired controls (Peterson and Renaud 1989). Grazing trials were analyzed as independent replicates as new herbivore and macroalgal individuals were used for each trial.  77  In the no-choice grazing trials, we tested the null hypothesis that all four macroalgal species were consumed at the same rate. To view consumption between the different herbivore species, we calculated the grams consumed of each macroalga for each herbivore species individually, and analyzed for differences using separate, univariate ANOVAs of macroalgal species followed by Tukey’s HSD. To view the mean effect across herbivores, we normalizied consumption rates among herbivores having different biomasses by divided the consumed macroalgal biomass by herbivore biomass to express macroalgal consumption per gram herbivore mass. We analyzed for differences in the amount of macroalgae consumed using a mixed model ANOVA with herbivore (seven levels) and macroalgal species (four levels) as fixed factors, and grazing trial (four levels) as a random factor. Significant main effects of macroalgal species were assessed using Tukey’s HSD. In the choice grazing trials, we tested the null hypothesis that both macroalgal species (the native and S. muticum) offered together would be consumed equally. To analyze individual preferences between the paired macroalgal choices (Table 4.1, Treatments 5-7), we plotted the grams of each macroalga consumed (g + 1SE), for each native herbivore species separately to view species-specific preferences. To test for differences in the amount of native versus exotic macroalgae consumed across all herbivores (the mean effect), we calculated the grams macroalgae consumed per gram of grazer mass, and analyzed for differences using a mixed model ANOVA with herbivore (seven levels) and alternate macroalgal species (three levels) as fixed factors, and grazing trial (four levels) as a random factor. Preference for one macroalgal species was  78  demonstrated if it was consumed significantly more than the alternative when offered in choice, determined using paired t-tests. Grazing trials (time) were pooled for the analysis of preference as no main effects of time were found in the ANOVA. A Bonferroni adjusted significance level (0.05/3, P = 0.017) was used for performing multiple tests. We tested for differences in the concentrations of polyphenolic compounds between macroalgal species in Barkley Sound using a mixed model ANOVA, with macroalgal species (four levels) as a fixed factor, and site (three levels) as a random factor. Significant effects of macroalgal species and site on polyphenolic content were explored using Tukey’s HSD. Polyphenolic data (% DW) were arcsine square root transformed to meet assumptions of normality. All analyses were performed in JMP 4.0.4 (SAS Inst.).  RESULTS Grazing and preference assays In the no-choice grazing trials, non-native Sargassum was consumed significantly less than native macroalgae by native herbivores (F3 = 179.43, P < 0.0001, Fig. 4.1H). Native herbivores differed in their consumption of native algae (ANOVA herbivore x macroalgae, F18 = 10.76, P < 0.0001, Table 4.2A, Fig. 4.1A-G). There results did not vary with grazing trial, although these results were marginally insignificant (F3 = 4.51, P = 0.06). Individually, six of the seven herbivore species consumed less Sargassum than any of the native macroalgal species (all P < 0.0001), with the exception of P. samuelis, which did not differentiate between non-native Sargassum and native Fucus (Fig. 4.1F).  79  In the choice grazing trial, there were significant differences in the amount of each macroalgal species consumed (F2 = 3.29, P = 0.02), but no difference in the total amount consumed by each herbivore species (F6 = 0.63, P = 0.70). Patterns of discrimination between the macroalgal treatment combinations varied with herbivore identity (ANOVA herbivore x macroalgae, F18 = 4.66, P < 0.0001, Fig. 4.2). Analyzed separately, each herbivore species preferentially consumed Sargassum over one native macroalga, but no clear patterns of preference emerged (Fig. 4.2A-G). The mean effect across herbivores revealed no preference between native and non-native algae (Fig. 4.2H). Consistent with the no-choice experiment, these results did not vary with grazing trial (F3 = 1.74, P = 0.18).  0  c  Alaria  Egregia  0  Fucus  1 0.8  1 0.8  b  0.6 0.5 0.3  0.6  0.4  0.4  0.2  0.2  0.1  0  0  0.0  c  c  Average effect H  b  0.4  0.6  0.2  Alaria  b  Gammarid amphipod a G b  Egregia  a  1.2  Fucus  b  1.2  Alaria  E  0  Pagurus samuelis a F b b a  Egregia  Lottia pelta  d  a  a  c  Macroalgal species  Figure 4.1 Macroalgae consumed (g + 1SE), when presented individually (no-choice trials) to native herbivores (A-G), and average effect across herbivores (H, g + 1SE per gram of herbivore mass). Non-native Sargassum is shown with closed bars, native species with open bars. Data are the mean of 24 replicates for each herbivore species. Significant differences in consumption between macroalgal species was determined from separate univariate ANOVAs for each herbivore species. Letters indicate significant differences between macroalgae (Tukey’s HSD).  b  Alaria  0.5  a  Egregia  2  c  Idotea wosnesenskii a D  Fucus  1  b  1.4 1.2 1 0.8 0.6 0.4 0.2 0  Sargassum  4  Astraea gibberosa a C  Sargassum  c  1 0.5  b  Strongylocentrotus purpuratus 10 2.5 a B a 8 2 b 6 1.5 c  Fucus  1.5  Pugettia producta a A ab  Sargassum  7 6 5 4 3 2 1 0  Sargassum  Macroalgae consumed (g)  80  81  Pugettia producta 14 12  A  Strongylocentrotus purpuratus 18  *  16  B  *  Macroalgae consumed (g)  14 10  12  8  10  6  8  2.5  *  2  0  0  *  *  1.0 0.5 0.0  Lottia pelta E  *  0.8  *  1.0 0.8  *  F  *  *  *  0.4  0.2  SA  SF  1.0  G  *  *  *  *  *  Average effect 2.5 2.0  H  *  1.5  0.0  1.0  0.4  0.5  0.2  0.2  SE  *  0.6 0.4  0.6  1.2  0.8  0.6  D  Gammarid amphipod  Pagurus samuelis 1.0  1.2  0.0  *  1.5  4 2  1.4  C  2.0  Idotea wosnesenskii 2.0 1.8 1.6 1.4 1.2 1.0 0.8 0.6 0.4 0.2 0.0  6  4  1.6  Astraea gibberosa  SE  SA  SF  0.0  0.0  SE  SA  SF  SE  SA  Paired macroalgal species  Figure 4.2 Macroalgae consumed (g + 1SE) when presented in choice trials to native herbivores individually (A-G), and average effect across herbivores (H, g + 1SE per gram of herbivore mass). Closed bars represent non-native Sargassum, open bars the native macroalgal alternate. Data are the mean of 24 replicates for each herbivore species. SE = Sargassum + Egregia, SA = Sargassum + Alaria, SF = Sargassum + Fucus. * significant differences in macroalgae consumed between paired macroalgal species from paired t-tests at the Bonferroni adjusted value of P < 0.017. Species codes are given in Table 4.1.  SF  82  Polyphenolic DW (%)  7  a  6 5  c  c  b  4 3 2 1 0  Sargassum muticum  Fucus distichus  Egregia menziesii  Alaria marginata  Macroalgal species Figure 4.3 Average polyphenolic content (% DW) of non-native Sargassum muticum (closed bar) and the three native macroalgae (open bars) from three sites in Barkley Sound (left to right, sites 1 to 3). Each bar represents the average of 15 replicates, each analyzed in triplicate. Error bars are SE. Significant differences in polyphenolic content between macroalgal species is denoted by letters (Tukey’s HSD, P < 0.05). The different macroalgal morphologies are shown above the bars. Table 4.2 ANOVA test criterion for the amount of macroalgae consumed in (A) nochoice, and (B) choice grazing trials, and (C) concentration of polyphenolic compounds.  Grazing trials Herbivore species Macroalgae Herbivore species x macroalgae Trial (C) Polyphenolic content Macroalgae Site Macroalgae x site  (A) No-choice (B) Choice DF F P DF F P 6 34.44 <0.0001 6 0.63 0.70 3 179.43 <0.0001 2 3.29 0.02 18 10.76 <0.0001 18 4.66 <0.0001 3 4.51 0.06 3 1.74 0.18 3 2 6  363.8 16.44 7.14  <0.0001 <0.0001 <0.0001  83  Polyphenolic assays Polyphenolic content varied significantly among macroalgal species (ANOVA, F3 = 363.8, P < 0.0001, Table 4.2C). Levels were highest in Fucus, followed by Alaria, both of which were significantly higher than non-native Sargassum, which did not differ from native Egregia (Tukey’s HSD, Fig. 4.3). Polyphenolic concentrations varied with site for Alaria, Egregia, and Sargassum, but not for Fucus (ANOVA F2 = 16.44, P < 0.0001, Table 4.2C).  DISCUSSION Competitive release In contrast to the predictions of the ERH, Sargassum is not consumed less by native grazers relative to native macroalgae despite the fact that it was consumed the least in no choice trials. Our study shows that in the absence of choice, native herbivores consistently have a higher grazing rate on native macroalgae than non-native Sargassum (Fig. 4.1). When presented a choice (which is presumably representative of nature), preferential consumption of the macroalgal species varied with herbivore identity, with no clear pattern of preference between native and non-native macroalgae apparent (Fig. 4.2A-G). Why in the absence of choice would native herbivores consume less non-native Sargassum than native macroalgae, but consume it in similar proportions to the native alternative when presented in choice? One possibility is that Sargassum has a higher nutritional value (lower C:N) than the native macroalgae tested. Assuming native herbivores consumed macroalgae until they obtain enough N, then they would consume  84  more native macroalgae than Sargassum in no choice trials. In the choice trials, it would be beneficial to preferentially consume Sargassum due to the higher nutritional value. Without further testing we are unable to disentangle the mechanism responsible for Sargassum being consumed in similar proportions to the native alternative when presented in choice, but consumed less when offered alone. As the choice grazing experiment is presumably more representative of natural conditions, we conclude that these results are likely more representative of patterns of consumption in natural conditions.  Factors determining macroalgal preferences of marine herbivores Numerous studies have demonstrated the ability of secondary polyphenolic metabolites (phlorotannins) in brown macroalgae to deter grazing by herbivores (Rosenthal and Janzen 1979, Geiselman and McConnell 1981, Steinberg 1984, 1985, 1988, Steinberg and van Altena 1992). We found polyphenolic content of Sargassum to be low relative to the native macroalgae tested, and similar to that of the native macroalga Egregia (Fig. 4.3). This suggests polyphenolic content is not influencing herbivore choice in this system. These results may suggest that Sargassum allocates relatively few resources to polyphenolic defenses. Phenotypic plasticity or selection for low polyphenolics in the invaded range may explain the relatively low concentration. Additional properties that may influence herbivore grazing could be differences in macroalgal morphology (i.e “toughness” or “palatability”), nutrient or caloric content, other non-polyphenolic chemical constituents, the chemical identity or molecular size of  85  the phlorotannins, or behavioural properties of the herbivore. Some herbivore species may require more time to detect the presence of, consume, and digest an individual of one macroalgal species than another (Underwood and Clarke 2007). Littler and Littler (1980) emphasized the importance of morphological defenses in influencing herbivory. Sargassum has a radially branched morphology (see Fig. 4.3), which is distinct from native low intertidal macroalgae in Barkley Sound. The three native macroalgal species exhibit different thallus structures, but are all solid forms compared to Sargassum. Its unique morphology may deter handling or consumption by some herbivores (i.e. Pugettia producta, Pagurus samuelis). Further, macroalgal toughness can influence grazing if macroalgae utilized differ in tissue structure. Also, the herbivore species used in the grazing trials differ in their feeding apparatus. A structurally tough macroalga offered to a herbivore that is not morphologically capable of triturating it would show low assimilation efficiency regardless of polyphenolic content (Targett et al. 1995). Further, differences in the gut characteristics of some herbivore species (such as chemical environment) influence the reactivity of polyphenolic compounds (Martin and Martin 1984, Martin et al. 1985, Hay and Steinberg 1992, Boettcher and Targett 1993, Targett et al. 1995). Chemicals in marine macroalgae other then polyphenolic compounds can also be important in determining herbivory rates (Geiselman 1974). For example, perhaps the best known macroalgal invader, Caulerpa taxifolia contains the defensive toxins caulerpicin and caulerpin, which concentrate in herbivore tissues (Doty and AguilarSantos 1970). The tropical macroalga Rhipocephalus phoenix contains rhipocephenal,  86  which has been shown to be toxic to fish (Sun and Fenical 1979). Some macroalgal substances are even toxic to humans, i.e. dinoflagellates responsible for shellfish poisoning (Geiselman 1974). One drawback to the Folin-Ciocalteu method used here is that it measures all hydroxylated aromatic compounds, not just phlorotannins. Thus, when reporting and comparing total polyphenolic content of different macroalgal species, it is assumed that all the phlorotannins quantified have similar functionality. In reality this group consists of polymers of phloroglucinol with a wide range of molecular sizes and structure, and common functionality may not be the case. For example smaller phlorotannins are generally more active as antioxidants than larger polymers (Nakamura et al. 1996), while they may not affect herbivore physiology to the same degree as higher molecular weight phlorotannins (Geiselman and McConnell 1981, Boettcher and Targett 1993). Using the Folin-Ciocalteu method alone, it was not possible to resolve the relative abundance and chemical identities of various phlorotannin species present in the tissue of different macroalgal species, which may be critical to determining their relative antiherbivory potential. Further, this work measures concentrations of polyphenolic compounds as indicators of the level of chemical defenses in S. muticum, but other compounds may also influence herbivore choice in this system. Geiselman (1974) recognized none of the above traits to be mutually exclusive, and concluded that a combination of different traits present in different macroalgal species likely determine their palatability. Polyphenolic content of the four brown macroalgal species studied showed significant interspecific variation within and between sites (Table 4.2B, Fig. 4.3). The  87  polyphenolic content of marine macroalgae is governed by both intrinsic (life history stage, Van Alstyne et al. 2001; tissue age, Pedersen 1984; reproductive status, Ragan and Jensen 1978), and environmental factors. The difference in polyphenolic content with site could be the result of a number of factors, including (but not limited to) differences in salinity, temperature, exposure, tidal height, and nutrient concentrations. Consistent with our findings, Van Alstyne et al. (1999) found significant site variation in polyphenolic content of vegetative blades of Alaria marginata (<1 - >3 % DW) and Egregia menziesii (<1 - >2 % DW) at similar levels to ours. Striking variation in polyphenolic content between brown macroalgal species also exists (Steinberg 1989, Van Alsytne et al. 2001, Connan et al. 2004). Connan et al. (2004) suggested an explanation for differences in polyphenolic content between two brown macroalgal species occupying the same habitat in relation to the potential protection provided by respective macroalgal morphologies. Species with long, flexible branches sweep the substratum around the base restricting herbivore access, therefore needing less chemical deterrents. In our study, Fucus displayed the highest mean polyphenolic content (~ 5.80% DW, Fig. 4.3) which is unsurprising as species of this genus typically exhibit high polyphenolic concentration (F. spiralis < 3% DW, Van Alstyne et al. 2001, F. vesiculosus ~ 5.80% DW, F. serratus ~ 4.27% DW, Connan et al. 2004). Fucoids are small (typically < 30 cm in BS) with rigid, upright thalli that are buoyant when submerged and have little impact on the surrounding substrate. Sargassum and Egregia both have long primary lateral branches (~ 2 metres in BS), with numerous branches arising from a single holdfast which frequently disturb the substrate through frond  88  sweeping. These species had the lowest polyphenolic content (Fig. 4.3). This suggests support for a trade-off in resources between chemical and structural defense, assuming polyphenolic compounds function as defensive compounds in Sargassum, a finding not supported by our results. Community level studies of the ERH such as this, while testing impacts of release from co-evolved enemies in the recipient community, fail to address biogeographic patterns. Support for the ERH at a biogeographical scale has emerged from a number of studies (reviewed in Colautti et al. 2004). A comparison of the levels of chemical defences of S. muticum in source and invaded populations will be presented separately. A previous study by the authors demonstrated that S. muticum displaces native macroalgae in this system through competition, which is conferred through aspects of its morphology (shading of smaller macroalgae by large, lateral fronds, White and Shurin, chapter three). The mechanisms driving the competitive advantage of Sargassum in this system are still unclear, but are not conferred through reduced grazing pressure by native herbivores. Patterns of consumption of native herbivores did not appear to be influenced by polyphenolic content. While enemy-release has been shown to play an important role in some exotic invasions (Schierenbeck et al. 1994, Lesica and Miles 1999, Goergen and Daehler 2001, Torchin et al. 2001, Wolfe 2002, Siemann and Rogers 2003, DeWalt et al. 2004) suggesting support for competitive release, other studies report the opposite trend where exotic species were more affected by native enemies in the exotic range (Clay 1995, Bellingham 1998, Gross et al. 2001, Agrawal and Kotanen 2003) suggesting other mechanisms are important. By example, grazing preference experiments using the  89  amphipod species Dexamine spinosa show that S. muticum was consumed in preference to the native macroalgae Saccharina latissima, Halidrys siliquosa, and Fucus serratus (Strong et al. 2009). These contrasting results suggest that release from natural enemies may contribute to invasion success of some exotic species, but may not generalize to all exotic species.  90  LITERATURE CITED Agrawal AA and Kotanen PM (2003) Herbivores and the success of exotic plants: a phylogenetically controlled experiment. Ecol. Lett. 6: 712-715. Barker KM, Chapman ARO (1990) Feeding preferences of periwinkles among four species of Fucus. Mar. Biol. 106: 113-118. Bellingham PJ (1998) Shrub succession and invasibility in a New Zealand montane grassland. Aust. J. Ecol. 23: 562-573. Boettcher AA and Targett NM (1993) Role of polyphenolic molecular size in reduction of assimilation efficiency of Xiphister mucosus. Ecology 74: 891-903. Blossey B and Notzold R (1995) Evolution of increased competitive ability in invasive nonindigenous plants: a hypothesis. J. Ecol. 83: 887-889. Byers JE (1999) The distribution of an introduced mollusk and its role in the long-term demise of a native confamilial species. Biol. Invas. 1: 339-352. Callaway RM and Ridenour WM (2004) Novel weapons: invasive success and the evolution of increased competitive ability. Front. Ecol. Environ. 2: 419-426. Clay K (1995) Correlates of pathogen species richness in the grass family. Can. J. Bot. 73: S42-S49. Colautti RI, Ricciardi A, Grigorovich IA, MacIsaac HJ (2004) Is invasion success explained by the enemy release hypothesis? Ecol. Lett. 7: 721-733. Connan S, Delisle F, Deslandes E, Ar Gall E (2006) Intra-thallus phlorotannin content and antioxidant activity in Phaeophyceae of temperate waters. Bot. Mar. 49: 39-46. Connan S, Goulard F, Stiger V, Deslandes E, Ar Gall E (2004) Phlorotannins in beltforming brown algae of a sheltered shore. Bot. Mar. 47: 410-416. Conover JT and Sieburth J McN (1964) Effect of Sargassum distribution on its epibiota and antibacterial activity. Bot. Mar. 6: l47-157. Denton A and Chapman ARO (1991) Feeding preferences of gammarid amphipods among four species of Fucus. Mar. Biol. 109: 503-506. Denton A, Chapman ARO, Markham J (1990) Size-specific concentrations of phlorotannins (antiherbivore compounds) in three species of Fucus. Mar. Ecol. Prog. Ser. 65: 103-104.  91  DeWalt SJ, Denslow JS, Ickes K (2004) Natural-enemy release facilitates habitat expansion of the invasive tropical shrub Clidemia hirta. Ecology 85: 471-483. Doty MS and Aguilar-Santos G (1970) Transfer of toxic algal substances in marine food chains. Pac. Sci. 24: 351-355. Estes JA and Steinberg PD (1988) Predation, herbivory, and kelp evolution. Paleobiology 14: 19-36. Geiselman JA (1974) Ecology of chemical defenses of algae against the herbivorous snail, Littorina littorea, in the New England Rocky Intertidal Community. PhD thesis, Massachusetts Institute of Technology and the Woods Hole Oceanographic Institution. Geiselman JA and McConnell OJ (1981) Polyphenols in brown algae Fucus vesiculosus and Ascophyllum nodosum: chemical defenses against the marine herbivorous snail, Littorina littorea. J. Chem. Ecol. 7: 1115-33. Goergen E and Daehler CC (2001) Inflorescence damage by insects and fungi in native pili grass (Heteropogon contortus) versus alien fountain grass (Pennisetum setaceum) in Hawai’i. Pac. Sci. 55: 129-136. Gross EM, Johnson RL, Hairston NG (2001) Experimental evidence for changes in submersed macrophyte species composition caused by the herbivore Acentria ephemerella (Lepidoptera). Oecologia 127: 105-114. Hamback PA, Agren J, Ericson L (2000) Associational resistance: insect damage to purple loosestrife reduced in thickets of sweet gale. Ecology 81: 1784-1794. Hamback PA, Pettersson J, Ericson L (2003) Are Associational Refuges SpeciesSpecific? Func. Ecol. 17: 87-93. Hay ME and Steinberg PD (1992) The chemical ecology of plant-herbivore interactions in marine versus terrestrial communities. In: Herbivores: their interaction with secondary metabolites, evolutionary and ecological processes. Eds: Rosenthal JA and Berenbaum MR. Academic Press, San Diego, pp. 371-413. Hemmi A, Honkanen T, Jormalainen V (2004) Inducible resistance to herbivory in Fucus vesiculosus - duration, spreading and variation with nutrient availability. Mar. Ecol. Prog. Ser. 273: 109-120. Irelan CD, Horn MH (1991) Effects of macrophyte secondary chemicals on food choice and digestive efficiency of Cebidichthys violaceus (Girard), an herbivorous fish of temperate marine waters. J. Exp. Mar. Biol. Ecol. 153: 179-194.  92  Jormalainen V and Ramsay T (2009) Resistance of the brown alga Fucus vesiculosus to herbivory. Oikos 118: 713-722. Keane RM and Crawley MJ (2002) Exotic plant invasions and the enemy release hypothesis. Trends Ecol. Evol. 17: 164-170. Koivikko R, Loponen J, Honkanen T, Jormalainen V (2005) Contents of soluble, cellwall-bound and exuded phlorotannins in the brown alga Fucus vesiculosus, with implications on their ecological functions. J. Chem. Ecol. 31: 195-212. Lesica P and Miles S (1999) Russian olive invasion into cottonwood forests along a regulated river in north-central Montana. Can. J. Bot. 77: 1077-1083. Liszka D and Underwood AJ (1990) An experimental design to determine preferences for gastropod shells by a hermit-crab. J. Exp. Mar. Biol. Ecol. 137: 47-62. Littler MM and Littler DS (1980) The evolution of thallus form and survival strategies in benthic marine macroalgae: field and laboratory tests of a functional form model. Am. Nat. 116: 24-44. Lubchenco J (1983) Littorina and Fucus: effects of herbivores, substratum, heterogeneity, and plant escapes during succession. Ecology 64: 1116-1123. Lüder UH and Clayton MN (2004) Induction of phlorotannins in the brown macroalga Ecklonia radiata (Laminariales, Phaeophyta) in response to simulated herbivory – the first microscopic study. Planta 218: 928-937. Martin MM, Martin JS (1984) Surfactants: their role in preventing the precipitation of proteins by tannins in insect guts. Oecologia 61: 342-345. Martin MM, Rockholm DC, Martin JS (1985) Effects of surfacrants, pH, and certain cations on precipitation of proteins by tannins. J. Chem. Ecol. 11: 485-494. Martinez EA (1996) Micropopulation differentiation in phenol content and susceptibility to herbivory in the Chilean kelp Lessonia nigrescens (Phaeophyta, Laminariales). Hydrobiologia 326/327: 205-211. McLachlan J and Craigie JS (1966) Antialga1 activity of some simple phenols. J. Phycol. 2: l33-l35. Mollo E, Gavagnin M, Carbone M, Castelluccio F, Pozone F, Roussis V, Templado J, Ghiselin MT, Cimino G (2008) Factors promoting marine invasions: a chemoecological approach. Proc. Natl. Acad. Sci. USA. 105: 4582-4586.  93  Nakamura T, Nagayama K, Uchida K, Tanaka R (1996) Antioxidant activity of phlorotannins isolated from the brown alga Eisenia bicyclis. Fish. Sci. 62: 923-926. Novotny V, Miller SE, Cizek L, Leps J, Janda M, Basset Y, et al. (2003) Colonising aliens: caterpillars (Lepidoptera) feeding on Piper aduncum and P. umbellatum in rainforests in Papua New Guinea. Ecol. Entomol. 28: 704-716. Pavia H, Cervin G, Lindgren A, Åberg P (1997) Effects of UV-B radiation and simulated herbivory on phlorotannins in the brown alga Ascophyllum nodosum. Mar. Ecol. Prog. Ser. 157: 139-146. Pavia H, Toth G, Aberg P (1999) Trade-offs between phlorotannin production and annual growth in natural populations of brown seaweed Ascophyllum nodosum. J. Ecol. 87: 761-771. Pavia H and Toth G (2008) Macroalgal models in testing and extending defense theories. Algal Chemical Ecology. pp. 147-172, in C. D. Amsler (ed.). Springer, Berlin. Pedersen A (1984) Studies on phenol content and heavy metal uptake in fucoids. Hydrobiologia 116/117: 498-504. Peterson CH, Renaud PE (1989) Analysis of feeding preference experiments. Oecologia 80: 82-86. Pfister CA, Hay ME (1988) Associational plant refuges: convergent patterns in marine and terrestrial communities result from differing mechanisms. Oecologia 77: 118129. Poulin R and Mouillot D (2003) Parasite specialisation from a phylogenetic perspective: a new index of host specificity. Parasitology 126: 473-480. Prider J, Watling J, Facelli JM (2009) Impacts of a native parasitic plant on an introduced and a native host species: implications for the control of an invasive weed. Annals of Botany 103: 107-115. Radho-Toly S, Majer JD, Yates C (2001) Impact of fire on leaf nutrients, arthropod fauna and herbivory of native and exotic eucalyptus in Kings Park, Perth, Western Australia. Aust. Ecol. 26: 500-506. Ragan MA and Jensen A (1978) Quantitative studies on brown algal phenols. II. Seasonal variation in polyphenol content of Ascophyllum nodosum (L.) Le Jol. and Fucus vesiculosus (L.) J. Exp. Mar. Biol. Ecol. 34: 245-258.  94  Rönnberg O and Ruokolahti C (1986) Seasonal variation of algal epiphytes and phenolic content of Fucus vesiculosus in a northern Baltic Archipelago. Ann. Bot. Fenn. 23: 317- 323. Rosenthal GA, Janzen DH (1979) Herbivores, their interaction with secondary plant metabolites. Academic Press, New York. Schierenbeck KA, Mack RN, Sharitz RR (1994) Effects of herbivory on growth and biomass allocation in native and introduced species of Lonicera. Ecology 75: 16611672. Schoenwaelder MEA and Clayton MN (1999) The presence of phenolic compounds in isolated cell walls of brown algae. Phycologia 38: 161-166. Schmitt TM, Hay ME, Lindquist N (1995) Constraints on chemically mediated coevolution: multiple functions for seaweed secondary metabolites. Ecology 76: 107-123. Sieburth JM (1968) The influence of algal antibiosis on the ecology of marine microorganisms. In: Advances in microbiology of the sea. Eds: Droop MR, Wood EJF Academic Press, New York, pp. 63-94. Siemann E and Rogers WE (2003) Increased competitive ability of an invasive tree may be limited by an invasive beetle. Ecol. Applic. 13: 1503-1507. Steinberg PD (1984) Algal chemical defenses against herbivores: allocation of phenolic compounds in the kelp Alaria marginata. Science 223: 405-7. Steinberg PD (1985) Feeding preferences of Tegula funebralis and chemical defenses of marine brown algae. Ecol. Monogr. 55: 333-49. Steinberg PD (1988) The effects of quantitative and qualitative variation in phenolic compounds on feeding in three species of marine invertebrate herbivores. J. Exp. Mar. Biol. Ecol. 120: 221-37. Steinberg PD (1989) Biogeographical variation in brown algal polyphenolics and other secondary metabolites: comparison between temperate Australasia and North America. Oecologia 78: 374-83. Steinberg PD, Edyvane K, de Nys R, Birdsey R, van Altena FA (1991) Lack of avoidance of phenolic rich brown algae by tropical herbivorous fishes. Mar. Biol. 109: 335343.  95  Steinberg PD (1992) Geographical variation in the interaction between marine herbivores and brown algal secondary metabolites. In: Ecological Roles of Marine Natural Products. Ed: Paul VJ. Cornell University Press, Ithaca, pp. 51-92. Steinberg PD and van Altena IA (1992) Tolerance of Australasian marine herbivores to brown algal phlorotannins. Ecol. Monogr. 62: 189-222. Stiger V, Deslandes E, Payri CE (2004) Phenolic contents of two brown algae, Turbinaria ornata and Sargassum mangarevense on Tahiti (French Polynesia): interspecific, ontogenic and spatio-temporal variations. Bot. Mar. 47: 402-409. Strong JA, Maggs CA, Johnson MP (2009) The extent of grazing release from epiphytism for Sargassum muticum (Phaeophyceae) within the invaded range. J. Mar. Biol. Ecol. UK. 89: 303-314. Sun HH and Fenical W (1979) Diterpenoids of the brown seaweed Glossophora galapagensis. Phytochemistry 18: 340-341. Swanson AK and Druehl LD (2002) Induction, exudation and the UV protective role of kelp phlorotannins. Aquat. Bot. 73: 241-253. Targett NM, Boettcher AA, Targett TE, Vrolijk NH (1995) Tropical marine herbivore assimilation of phenolic-rich plants. Oecologia 103: 170-179. Torchin ME, Lafferty KD, Kuris AM (2001) Release from parasites as natural enemies: increased performance of a globally introduced marine crab. Biol. Inv. 3: 333-345. Underwood AJ, Chapman MG, Crowe TP (2004) Identifying and understanding ecological preferences for habitat or prey. J. Exp. Mar. Biol. Ecol. 300: 161-187. Underwood AJ and Clarke KR (2007) More response on a proposed method for analyzing experiments on food choice. J. Exp. Mar. Biol. Ecol. 344: 13-115. Van Alstyne KL (1988) Grazing increases polyphenolic defenses in the intertidal brown alga Fucus distichus. Ecology 69: 655-663. Van Alstyne KL 1995. A comparison of three methods for quantifying brown algal polyphenolic compounds. J. Chem. Ecol. 21: 45-58. Van Alstyne KL and Paul VJ (1990) The biogeography of polyphenolic compounds in marine macroalgae: temperate brown algal defenses deter feeding by tropical herbivorous fishes. Oecologia 84: 158-163.  96  Van Alstyne KL, Mccarty III JJ, Hustead CL, Duggins DO (1999a) Geographic variation in polyphenolic levels of Northeastern Pacific kelps and rockweeds. Mar. Biol. 133: 371-379. Van Alstyne KL, Whitman SL, Ehlig JM (2001) Differences in herbivore preferences, phlorotannin production, and nutritional quality between juvenile and adult tissues from marine brown algae. Mar. Biol. 139: 201-210. White LF and Shurin JB (2007) Diversity effects on invasion vary with life history stage in marine macroalgae. Oikos 116: 1193-1203. Wolfe LM (2002) Why Alien Invaders Succeed: Support for the Escape-from-Enemy Hypothesis. Am. Nat. 160: 705-711. Yates JL and Peckol P (1993) Effects of nutrient availability and herbivory on polyphenolics in the seaweed Fucus vesiculosus. Ecology 74: 1757-66.  97  CHAPTER FIVE BIOGEOGRAPHIC PATTERNS OF SIZE AND CHEMICAL DEFENSE OF AN INVASIVE MARINE MACROALGA  INTRODUCTION  One unexplained pattern in the ecology of invasive species is how exotic plants occurring at low densities in their native ranges attain high densities in their introduced ranges (Callaway and Ridenour 2004). This is manifested as more vigorous growth and reproduction in the invasive than in the native distribution (Crawley 1987, Willis and Blossey 1999, Leger and Rice 2003) and can strongly influence the magnitude of the impacts of the invader on native communities. The vigour and success of aliens in nonnative regions may be attributed to increased resource availability as a result of release from natural phytophagous enemies (the Enemy-Release Hypothesis, Keane and Crawley 2002). The Evolution of Increased Competitive Ability hypothesis (EICA, Blossey and Nötzold 1995) invokes optimal defence theory to predict that plants with limited resources show trade-offs between resource investments among maintenance, growth, storage, reproduction, and defense (Coley et al. 1985). In the absence of coevolved herbivores, selection will favour genotypes that allocate fewer resources to herbivore defense, and more to growth and reproduction, thereby improving competitive abilities (Callaway and Ridenour 2004). Competitive abilities can be maximized by increasing 1  A version of this chapter will be submitted for publication. White LF and Nienhuis SB (2010) Biogeographic patterns of size and chemical defense of an invasive marine macroalga.  98  vegetative growth or reproductive efforts depending on which is more important for success in a particular new environment (Blossey and Notzold 1995). Empirical evidence is mixed; some studies support the predictions of the EICA hypothesis (Blossey and Notzold 1995, Willis and Blossey 1999, Ledger and Rice 2003, Grosholz and Ruiz 2003, Brown and Eckert 2005), while others have found the opposite trend (Evolutionary Reduced Competitive Ability, Vilá et al. 2003, van Kleunen and Schmid 2003, Bossdorf et al. 2004). Whether release from native enemies results in increased competitive abilities of exotic species remains poorly understood (Callaway and Ridenour 2004). Macroalgal invasions are a growing problem worldwide often with dramatic effects on ecosystem structure and function (Walker and Kendrick 1998). Analogous to terrestrial plants, marine macroalgae are defended against herbivores both mechanically and chemically. Mechanical defenses include structural form (i.e. turf and encrusting forms) and tissue ‘toughness’ (i.e. calcification and structural carbohydrates). Chemical defenses can be expressed as polyphenolic compounds, which are metabolic constituents of marine macroalgae. In the brown macroalgae (Phaeophyceae), polyphenolic compounds are expressed as phlorotannins, which are well studied (reviewed by Amsler and Fairhead 2006). These polyphenolic compounds can posses primary metabolic functions as cell wall components (Schoenwaelder and Clayton 1999), and often display a putative secondary function in the chemical defense of the thallus (algal body) against bacterial infections (Sieburth 1968), endophytes (Geiselman 1974), UV damage (Swanson and Druehl 2002), and fouling (Schmitt et al. 1995), and serve as allelopathic (McLachlan and Craigie 1966) and herbivory deterrents. Polyphenolic compounds have  99  been shown to deter grazing by fishes, urchins, gastropods and crustacean herbivores in both lab assays (Geiselman and McConnell 1981, Steinberg 1984, 1985, 1988, Irelan and Horn 1991, Steinberg and van Altena 1992, Luder and Clayton 2004) and the natural environment (Jormalainen and Ramsay 2009), although their effectiveness as herbivore deterrents is variable (reviewed by Pavia and Toth 2008). The ability of this class of compounds to deter grazing varies widely depending on the size and chemical identity of the phlorotannin polymers, and on the digestive physiology of the herbivore (Hay et al. 1988 a,b). Similar polyphenolic compounds, or even the same chemical species, can have variable effects on different herbivores (Hay 1991, Hay and Steinberg 1992).  The Invader Sargassum muticum (Yendo) Fensholt (Phaeophyceae: Fucales) is one of a handful of prevalent macroalgal aliens to become a dominant invasive in non-native regions. Native to Japan, S. muticum was an unintentional introduction with Japanese oysters (Crassostrea gigas) transplanted to British Columbia around 1940 for aquaculture (Scagel 1956). Sargassum muticum subsequently became well established in lower intertidal habitats, expanding in range north to south-eastern Alaska (Wallentinus 1999) and south to Baja, California (Nunez Lopez and Valdez 1998). From British Columbia, Japanese oysters were transplanted to Atlantic France, unintentionally translocating S. muticum with them, as predicted by Druehl (1973). It was first recorded from the Isle of Wight in 1971 (Farnham et al. 1973) where vegetative drift across the English Channel is believed responsible for its spread to and within England. Introduced populations in  100  Europe extend along the coasts of France, Scandinavia, the Baltic Sea, Helgoland, the Netherlands, the Iberian Peninsula, and into the Mediterranean from Italy and the Adriatic Sea (Thomas 2002). It has become a conspicuous dominant on sheltered shores, in many places displacing native macroalgal species, and reducing the native community richness and biomass (Britton-Simmons 2004, Sanchez and Fernandez 2005, Harris et al. 2007, Kraan 2007, White and Shurin chapter three). It has been suggested that S. muticum is more successful in non-native regions, attaining larger sizes, higher densities, and a wider depth distribution than in the native range of Japan (Norton 1977b, Critchley 1983, Rueness 1989). However, there is limited information about the ecology of S. muticum in its native range. Norton (1977b), Critchley (1983), and Hirata et al. (2003) reported S. muticum as a relatively minor component of the native Japanese algal flora. Supporting this, Tamada (1955) listed 70 species of Sargassum from Japan and its adjacent regions and no mention was made of S. muticum, which was taxonomically recognized as a separate species by this time. Further, the Japanese distribution of S. muticum appears to be restricted to areas influenced by warm currents (Critchley 1983), and is restricted to the subtidal (Hanyuda, pers comm), which likely contributes to the paucity of information on this species in its native range. Accounts of the size of S. muticum in Japan vary. Josefsson and Jansson (2006) reported S. muticum as 0.75-1.2 metres long, and Curiel et al. (1998) reported S. muticum averages 1.2 metres. Similarly, Rueness (1989) reported S. muticum off the coast of Japan attains lengths of 1-1.5 metres, but is highly invasive in non-native regions. However, no  101  data are presented to support these apparent patterns. The only morphological data we could find on S. muticum in its native range was a study of the drift macroalgae of central Japan (Hirata et al. 2003), which showed the maximum length of adult S. muticum rarely exceeded 45 cm. While not attached to the substrate, individuals collected were whole, and therefore likely similar in size to attached individuals. In the northern (British Columbia) portion of its introduced Pacific Coast range, low intertidal populations of S. muticum are common (Sánchez and Fernández 2005, White and Shurin 2007), growing to depths of ~2 metres (Cohen 2005), and attaining intertidal densities of 64 individuals m–2 (L White unpub. data), and subtidal densities of 126 individuals m–2 in nearby Washington (Britton-Simmons 2004). DeWreede (1978) reported low intertidal fronds reaching maximum lengths of ~1.05 metres in this region, and Druehl (2000) frond lengths commonly between 1-3 metres. In southern England S. muticum grows deeper, from the low intertidal to 6-8 metres depth (Cohen 2005), and commonly exceed intertidal lengths of 6 metres (Critchley et al. 1986). Densities of 130300 individuals m–2 have been observed in European waters (Fernández et al. 1990). In other invaded regions, S. muticum reaches lengths of 1.5-2 metres in Swedish waters, 6-7 metres in French waters, and up to 8.5 metres in Norwegian waters (Josefsson and Jansson 2006), showing considerable plasticity between invaded regions. Davidson (1996) reported S. muticum fronds of >4 metres common in Northern Ireland, and Cohen (2005) fronds of 3-5 metres in southern California. These data suggest S. muticum may be bigger in non-native regions, but no studies have directly tested this, and sampling of S. muticum from Japan is needed to establish the size of individuals in its native region.  102  The phenology of S. muticum’s life cycle also varies regionally. Reproductive periods of populations in Washington, British Columbia, and southern England are similar (DeWreede 1978, Deysher 1984). Lateral branches reach their maximum height in June, are reproductive from July to August, germlings are released after the spring tides (Fletcher 1975), and individuals begin to senesce in mid- to late- August (Deysher 1984, Britton-Simmons 2004). This is significantly later than populations from the Sea of Japan where S. muticum is reproductive April and May (pers. obs in Deysher 1984, Hirata et al. 2003), has a shorter reproductive cycle (Curiel et al. 1998), and gamete release occurs during the spring tides (Okuda 1981). White and Nienhuis (chapter three) in a test of the Enemy Release Hypothesis (ERH, Keane and Crawley 2002) in marine macroalgae in western Canada, showed nonnative S. muticum had similar concentrations of defensive polyphenolic compounds to some native brown macroalgae. While polyphenolic content of native and non-native macroalgae are similar in the invaded region, we do not know how levels of these chemical defenses vary between the native and non-native regions. Here we make preliminary comparisons of the size (as a measure of performance), and levels of chemical defenses of S. muticum from both its native and invaded regions. If the EICA theory explains invasion success of S. muticum, individuals in the invaded regions should be larger, and have lower levels of chemical defenses relative to the native region.  103  MATERIALS AND METHODS Sample collections Sargassum muticum individuals were randomly selected from 14 distinct populations within the native and invaded regions (Table 5.1, Fig. 5.1). We could find no information on the natural history status of S. muticum in Korea, so we termed these populations cryptogenic.  Table 5.1 Location of Sargassum muticum collection sites (populations 1-14) within regions (A-E).  Population  Coordinates  Native Region, western Japan 1 D Oh-ura, Takeno, Hyōgo Prefecture 2 D Toyooka, Hyōgo  35°39'35"N, 134°44'54"E 35°30'45"N, 134°10'43"E  Cryptogenic Region, South Korea 3 D Jangsungpo-dong, Geoje-si, Kyungsangnam-do 4 D Jeongdo-ri, Wando-gun, Chollanam-do 5 D Sorok-do, Goheung-gun, Chollanam-do  34°51'43"N, 128°43'33"E 34°17'33"N, 126°41'42"E 34°31'08"N, 127°06'28"E  Non-Native Regions British Columbia (BC), western Canada 6 B Outer Dixon Island, Barkley Sound, BC 7 B Kelp Bay, Barkley Sound, BC 8 B Ross Islets, Barkley Sound, BC 9 C Brockton Pt. North, Vancouver, BC 10 C Brockton Pt., East, Vancouver, BC 11 C Brockton Pt. West, Vancouver, BC Plymouth, southern England 12 E Church Reef, Wembury, Devon 13 E The Hoe, Plymouth, Devon 14 E Salcombe, Devon  48°51'43"N, 125°06'49"W 48°51'01"N, 125°07'07"W 48°52'18"N, 125°08'11"W 49°03'22"N, 123°07'01"W 49°01'45"N, 123°06'42"W 49°02'03"N, 123°07'03"W 50°18'58"N, 04°04'54"W 50°21'48"N, 04°08'40"W 50°13'57"N, 03°46'15"W  104  B  75km  A British Columbia  Vancouver Island  Barkley Sound  Burrard Inlet  6km  2km  Barkley Sound  C  NE Pacific Burrard Inlet  Fig. 5.1 Collection regions (A) British Columbia (western Canada), (B) Barkley Sound, (C) Burrard Inlet, (D) SE Asia, (E) Plymouth, England.  Vancouver  D  E England  Sea of Japan  South Korea  Japan  NW Pacific 400km  English Channel France 100km  Polyphenolic concentrations are phenotypically plastic depending on a suite of environmental factors. All individuals were collected from the lower intertidal zone of each site to control for variation due to tidal height (Connan et al. 2004). No rock pool individuals were selected to control for effects of emersion time (Martínez 1996). This  105  a priori site selection requirement also controlled for differences in S. muticum morphology between exposed substrata and rockpools (DeWreede 1978). Visually undamaged, non-epiphytized individuals were collected, as correlations between damage and polyphenolic concentrations have been reported (Van Alstyne 1988, Lüder and Clayton 2004). In an attempt to standardize effects of seasonal or temporal variation within the different collection regions, the five Asian populations (two Japan, three South Korea, Fig. 5.1D) were collected in early April 2007. The six western Canada (Fig. 5.1B,C) and three southern England populations (Fig. 5.1E) were collected in late June and early July 2007. These collection times correspond to pre-reproductive phenologies in the different regions, when individuals are typically at their maximum size (DeWreede 1978, Deysher 1984, Hirata et al. 2003, Britton-Simmons 2004). Macroalgae from each site were collected on the same tidal cycle. Japan is surrounded by the Japan and East China Seas and the Pacific Ocean, and South Korea by the Yellow, Japan and East China Seas. These different water masses range from temperate to subtropical in temperature. Asian samples were all collected from sites along the temperate Sea of Japan (or Korea East Sea as it is known in South Korea, Fig. 5.1D). Visible epifauna were removed before individuals were bagged separately and shipped on ice in a dark cooler to prevent photodegradation. Samples were frozen at -80˚ prior to shipping on dry ice. All storage and chemical analyses were performed at Bamfield Marine Sciences Centre (herein BMSC).  106  Size measurements The length of each macroalgal individual was measured prior to chemical analysis. As not all individuals were collected with the holdfasts intact, we measured the length of the annual primary lateral, from the top of the basal axes to the apex (Fig. 5.2), which regenerates each spring from perennial basal axes. Samples from South Korea and southern England comprised apical segments only, so individuals were not measured. For one Japanese population (population 1, Table 5.1), only eight of 15 individuals were measured as the rest were collected as fragments.  Polyphenolic bioassays To measure polyphenolic concentration, we excised 0.5 g of apical tissue from 15 individuals from each population (blotted mass). We specifically selected nonreproductive tissue to standardize our assays as polyphenolic concentration and physode density (vacuoles in which polyphenolics are housed) vary with tissue type (Pavia et al. 2002). Tissue samples were homogenized in liquid nitrogen by hand with a mortar and pestle, and extracted in 80% aqueous methanol in darkness at 4ºC for 12 hours. Polyphenolic concentrations were measured using the methods described in chapter four following Van Alstyne (1995). Absorbance of each solution was read spectrophotometrically at 765 nm (Biochrom Ultrospec 2100 pro) using phloroglucinol (Fisher Scientific) as standard. Total polyphenolic content was expressed as percent of dry weight (DW) in grams, based on the mean of triplicate assays per extract.  107  Data analysis To assess whether S. muticum individuals were larger in non-native regions, we tested for differences in size of individuals between the native region (Japan) and invaded regions of western Canada using a univariate analysis of variance (ANOVA). To analyze for differences in polyphenolic content of S. muticum between sites, we used a 2-factor ANOVA (% DW), nesting the factor ‘population’ (random) within the factor ‘region’ (fixed). Significant differences between all the collection regions were explored using Tukey-Kramer HSD for unequal sample sizes. Polyphenolic concentrations (% DW) were arcsine square root transformed, and size measurements log transformed to conform to the assumptions of ANOVA. All analyses were performed in JMP 4.0.4 (SAS Inst.).  RESULTS Sargassum muticum from the invaded regions in western Canada (Barkley Sound and Brockton Pt) were on average 53 cm larger than individuals from Japan (F2 = 63.35, P < 0.0001, Fig. 5.2). Polyphenolic content (% DW) of S. muticum varied significantly between regions (F4 = 38.24, P <0.0001), but population differences within regions were marginally insignificant (F9 = 1.87, P = 0.06). Post hoc tests revealed S. muticum had significantly higher levels of polyphenolic compounds in its native range, compared to all other regions (Tukey-Kramer HSD, Fig. 5.3). South Korea had similar polyphenolic levels to southern England, which were significantly higher than the two western Canada regions (Tukey-Kramer HSD Fig. 5.3).  Mean S. muticum length (cm)  108  native invaded  200  n = 45  150  n = 45 100  n = 23  50 0  Japan  Barkley Sound Brockton Pt.  Region  Figure 5.2 Mean length of Sargassum muticum individuals from the native region (Japan) and invaded regions of western Canada (Barkley Sound and Brockton Pt.). Measurements are the length of the longest primary lateral, from top of the basal axes to the apex. Significant differences in size between regions were demonstrated with ANOVA (P < 0.0001). Error bars are SE.  109  Polyphenolic DW (%)  4  native cryptogenic invaded  3  2  1  0  1  2  Japan  3  4  5  6  7  8  South Korea Barkley Sound  9  10  11  12  13  14  Brockton Pt. southern England  Population  Figure 5.3 Mean polyphenolic content (% DW) of Sargassum muticum individuals from distinct populations (1-14) within different geographic regions. Data are the mean of 15 replicate adult individuals, each analyzed in triplicate. Significant differences in polyphenolic content between regions were demonstrated with ANOVA (P < 0.0001). Error bars are SE.  DISCUSSION The present study suggests that in non-native regions, S. muticum attains larger sizes (Fig. 5.2) with lower levels of defensive polyphenolic compounds (Fig. 5.3) relative to the native region, as predicted by the EICA hypothesis. S. muticum is the largest macroalga present in the mid to lower intertidal belt it inhabits in British Columbia,  110  overlying smaller native species. In chapter three, I showed that S. muticum out-competes native macroalgae in this system through the monopolization of space and light, restricting these resources for smaller, underlying native macroalgae. These results indirectly suggest that the increased size of S. muticum in this system confers competitive superiority over native macroalgae, contributing to the success of this invasive species in British Columbia. This study is the first to examine traits of this global invader between its native and non-native regions. Our results provide preliminary support for an increased size and reduced chemical defense of S. muticum in non-native regions. However, our results hinge on the assumption that the thalli collected were of maximal size based on the limited information available on S. muticum’s phenology in the native region. The larger size of S. muticum outside of its native range is often reported in the literature (Rueness 1989, Critchley 1983), but supported by limited data. Studies of S. muticum are largely from the non-native regions where it is often a conspicuous component of the local macroalgal community, unlike Japan where it achieves only relatively low densities (Critchley 1983). With limited literature for comparison, sampling of more populations from the native range is needed to rigorously establish the size patterns found in this paper. Reports from the literature suggest S. muticum attains its largest size and densities in southern England (Critchley et al. 1986, Rueness 1989). Sampling populations from this region would determine whether S. muticum is more successful there than in other invaded regions.  111  Polyphenolic variability This study does not attempt to identify the mechanisms responsible for the reduced polyphenolic content of S. muticum in non-native regions. Polyphenolic compounds have been shown to vary through time (Connan et al. 2004), and with life history stage in some brown macroalgae (Van Alstyne 2001, Pavia et al. 2002). Differences in the collection times between regions could explain the observed differences in polyphenolic content. Sargassum muticum’s phenology varies regionally (Deysher 1984), and the different collection times correspond to pre-reproductive phenologies of S. muticum in the region of collection. Samples from southern England were collected at the same time as the western Canada populations, but were similar in polyphenolic content to samples from South Korea which were collected eight weeks earlier. The similar polyphenolic levels between southern England and South Korea suggest differences in collection times may not explain the differences in polyphenolic content. The homogeneity of polyphenolic content between populations within regions was surprising as environmental factors also vary at this level. Marine brown macroalgae have evolved the ability to induce production of polyphenolic defenses when the benefits of that defense can be realized, instead of maintaining costly constitutive defenses in low or sporadic periods of herbivory (Pavia et al. 1999). This cost-saving strategy is common across brown macroalgal species (Pavia and Toth 2000, Hemmi et al. 2004, Toth and Pavia 2007, Jormalainen and Honkanen 2008, Jormalainen and Ramsay 2009). While the ability to induce polyphenolic defenses in S. muticum have not been directly shown, they have with other temperate brown  112  macroalgal species, i.e. Fucus distichus (Van Alstyne 1988), and Ascophyllum nodosum (Pavia and Toth 2000). If S. muticum up-regulates biosynthesis of defense metabolites upon herbivory, this could explain the higher polyphenolic content in the native region where grazing pressure is likely stronger in the presence of co-evolved herbivores.  Cost of chemical defense Two possible mechanisms have been suggested to explain variation in polyphenolic content in non-native regions; a plastic phenotypic response to novel conditions in a new environment (i.e. exotic species losing defenses under reduced herbivore pressure), or genetic changes in invasive populations. Both can occur if the production of herbivore deterrent compounds imposes a fitness cost that reduces resources available to support growth and reproduction, as well as secondary metabolite production (Targett and Arnold 1998). It follows that the resources expended on chemical defense could be reallocated to metabolic processes which could result in increased growth and reproduction. The costs of defending against predators or grazers have been studied extensively in terrestrial plants (reviewed in Bergelson and Purrington 1996, Koricheva 2002). In macroalgae, negative correlations between polyphenolic levels and fitness components have been demonstrated at the phenotypic level (Steinberg 1984, Pfister 1992, Yates and Peckol 1993, Agrawal 1998, Pavia et al. 1999). However, several studies in other systems have failed to detect trade-offs in resource allocation (Simms and Rausher 1989, Rousi et al. 1991, Adler et al. 1995). Variable results suggest that the expression of costs may be condition-dependent (Strauss et al. 2002).  113  To determine whether trade-offs in resource allocation occur in S. muticum across geographic regions, studies of growth or reproduction and polyphenolic content where environment is controlled, i.e. a common garden design, such as those carried out on some terrestrial plants (i.e. Donaldson et al. 2005, Feng et al. 2009) would be required, as environmental variation among regions may cause phenotypic covariances to differ in both sign and magnitude (Rausher 1992). Our study measured phenotypic correlations between a single defense trait (polyphenolic content) and a single morphological trait (size) of S. muticum. Trade-offs between plant growth and reproduction also occur, such secondary trade-offs may preclude the detection of fitness costs if only one trait (i.e. growth) is measured (Mole 1994). Further, Steinberg (1995) found temporal variation in resource trade-offs in the phenotypic correlation between polyphenolic production and growth. Analysis of temporal variation in the expression levels of polyphenolic compounds in S. muticum individuals from different populations would inform as to whether the observed levels are maintained, suggesting selective pressures are at play, or induced in response to variable grazing pressure.  114  LITERATURE CITED Adler LS, Schmitt J, Bowers MD (1995) Genetic variation in defensive chemistry in Plantago lanceolata (Plantaginaceae) and its effect on the specialist herbivore Junonia coenia (Nymphalidae). Oecologia 101: 75-85. Agrawal AA (1998) Algal defense, grazers, and their interactions in aquatic trophic cascades. Acta Oecologica 19: 331-337. Amsler CD and Fairhead VA (2006) Defensive and sensory chemical ecology of brown algae. Adv. Bot. Res. 43: 1-91. Bergelson J and Purrington CB (1996) Surveying patterns in the cost of resistance in plants. Am. Nat. 148: 536-558. Blossey B and Nötzold R (1995) Evolution of Increased Competitive Ability in Invasive Nonindigenous Plants: A Hypothesis. J. Ecol. 83: 887-889. Bossdorf O, Prati D, Auge H, Schmid B (2004) Reduced competitive ability in an invasive plant. Ecol. Lett. 7: 346-353. Britton-Simmons KH (2004) Direct and indirect effects of the introduced alga Sargassum muticum on benthic, subtidal communities of Washington State, USA. Mar. Ecol. Prog. Ser. 277: 61-78. Brown JS and Eckert CG (2005) Evolutionary increase in sexual and clonal reproductive capacity during biological invasion in an aquatic plant Butomus umbellatus (Butomaceae). American Journal of Botany 92: 495-502. Callaway RM and Ridenour WM (2004) Novel weapons: invasive success and the evolution of increased competitive ability. Front. Ecol. Environ. 2: 419-426. Cohen AN (2005) Guide to the Exotic Species of San Francisco Bay. San Francisco Estuary Institute, Oakland, CA. www.exoticsguide.org. Accessed June 2009. Coley PD, Bryant JP, Chapin Fs III (1985) Resource availability and plant antiherbivore defense. Science 230: 895-899. Connan S, Goulard F, Stiger V, Deslandes E, Ar Gall E (2004) Phlorotannins in belt forming brown algae of a sheltered shore. Bot. Mar. 47: 410-416. Crawley MJ (1987) What makes a community invasible? In: Colonization, Succession and Stability (Eds Gray AJ, Crawley MJ and Edwards PJ). Blackwell Scientific Publications, Oxford, pp. 429-453.  115  Critchley AT (1983) Sargassum muticum. A taxonomic history including world-wide and western Pacific distributions. J. Mar. Biol. Ass. UK. 63: 617-625. Critchley AT, Farnham WF, Morrell SL (1986) An account of the attempted control of an introduced marine alga, Sargassum muticum, in southern England. Biol. Conserv. 35: 313-332. Davison DM (1996) Sargassum muticum in Strangford Lough, 1995-1998; a review of the introduction and colonisation of Strangford Lough MNR and cSAC by the invasive brown algae Sargassum muticum. Report to the Environment and Heritage Service, D.o.E. (N.I.) DeWreede RE (1978) Phenology of Sargassum muticum (Phaeophyta) in the Strait of Georgia, British Columbia. Syesis 2: 1-9. Deysher LE (1984) Reproductive phenology of newly introduced brown alga, Sargassum muticum (Yendo) Fensholt. Hydrobiologia 116/117: 403-7. Donaldson JR, Kruger EL, Lindroth RL (2005) Competition- and resource-mediated tradeoffs between growth and defensive chemistry in trembling aspen (Populus tremuloides). New Phytologist. 169: 561-570. Druehl L (1973) Marine transplantation. Science 179: 112 Druehl L (2000) Pacific Seaweeds. Harbour Publishing. Farnham W, Fletcher RL, Irvine LM (1973) Attached Sargassum found in Britain. Nature 243: 231-232. Feng YL, Lei YB, Wang RF, Callaway RM, Valiente-Banuet A, Li YP, Zheng YL (2009) Evolutionary tradeoffs for nitrogen allocation to photosynthesis versus cell walls in an invasive plant. Proc. Natl. Acad. Sci. USA. 106: 1853-1856. Fletcher RL (1975) Studies on the recently introduced brown alga Sargassum muticum (Yendo) Fensholt I. Ecology and reproduction. Bot. Mar. 18: 149-156. Geiselman JA (1974) Ecology of chemical defenses of algae against the herbivorous snail, Littorina littorea, in the New England Rocky Intertidal Community. PhD thesis, Massachusetts Institute of Technology and the Woods Hole Oceanographic Institution. Geiselman JA and McConnell OJ (1981) Polyphenols in brown algae Fucus vesiculosus and Ascophyllum nodosum: chemical defenses against the marine herbivorous snail,  116  Littorina littorea. J. Chem. Ecol. 7: 1115-33. Grosholz ED and Ruiz GM (2003) Biological invasions drive size increases in marine and estuarine invertebrates. Ecol. Lett. 6: 705-710. Harries DB, Harrow S, Wilson JR, Mair JM, Donnan DW (2007) The establishment of the invasive alga Sargassum muticum on the west coast of Scotland: a preliminary assessment of community effects J. Mar. Biol. Ass. U.K. 87: 1057-1067. Hay ME (1991) Fish-seaweed interactions on coral reefs: effects of herbivorous fishes and adaptations of their prey. In: The Ecology of Fishes on Coral Reefs. Chp. 5 Ed: Sale PF. Academic Press Inc., New York. Hay ME and Fenical W (1988) Marine plant-herbivore interactions: the ecology of chemical defense. Ann. Rev. Syst. Ecol. 19: 111-145. Hay ME, Renaud PE, Fenical W (1988) Large mobile versus small sedentary herbivores and their resistance to seaweed chemical defenses. Oecologia 75: 246-252. Hay ME, Duffy JE, Fenical W, Gustafson K (1988b) Chemical defense in the seaweed Dictyopteris deliculata: differential effects against reef fishes and amphipods. Mar. Ecol. Prog. Ser. 48: 185-192. Hay ME and Steinberg PD (1992) The chemical ecology of plant-herbivore interactions in marine versus terrestrial communities. In: Herbivores: their interactions with secondary plant metabolites, Volume II: Ecological and evolutionary processes. Eds: Rosenthal GA and Berenbaum MR. Academic Press, San Diego, CA, pp. 371413. Hemmi A, Honkanen T, Jormalainen V (2004) Inducible resistance to herbivory in Fucus vesiculosus - duration, spreading and variation with nutrient availability. Mar. Ecol. Prog. Ser. 273: 109-120. Hirata T, Tanaka J, Iwami T, Ohmi T, Dazai A, Aoki M, Ueda H, Tsuchiya Y, Sato T, Yokohama Y (2003) Ecological studies on the community of drifting seaweeds in the south-eastern coastal waters of Izu Peninsula, central Japan. II: Seasonal changes in plants showing maximum stipe length in drifting seaweed communities. Phycological Research. 51: 186-191. Irelan CD, Horn MH (1991) Effects of macrophyte secondary chemicals on food choice and digestive efficiency of Cebidichthys violaceus (Girard), an herbivorous fish of temperate marine waters. J. Exp. Mar. Biol. Ecol. 153: 179-194. Jormalainen V and Honkanen T (2008) Macroalgal chemical defenses and their roles  117  in structuring temperate marine communities. In: Algal Chemical Ecology. Ed: Amsler CD. Springer-Verlag Berlin Heidelberg. Jormalainen V and Ramsay T (2009) Resistance of the brown alga Fucus vesiculosus to herbivory. Oikos 118: 713-722. Josefsson M and Jansson K (2006): NOBANIS – Invasive Alien Species Fact Sheet – Sargassum muticum. - From: Online Database of the North European and Baltic Network on Invasive Alien Species-NOBANIS www.nobanis.org. Accessed May 2009. Keane RM and Crawley MJ (2002) Exotic plant invasions and the enemy release hypothesis. Trends Ecol. Evol. 17: 164-170. Koricheva J (2002) Meta-analysis of sources of variation in fitness costs of plant antiherbivore defenses. Ecology 83: 176-190. Kraan S (2007) Sargassum muticum (Yendo) Fensholt in Ireland: an invasive species on the move. J. Appl. Phycol. 20: 825-832. Leger EA and Rice KJ (2003) Invasive California poppies (Eschscholzia californica Cham.) grow larger than native individuals under reduced competition. Ecol. Lett. 6: 257-264. Lüder UH and Clayton MN (2004) Induction of phlorotannins in the brown macroalga Ecklonia radiata (Laminariales, Phaeophyta) in response to simulated herbivory: the first microscopic study. Planta 218: 928-937. Macaya EC and Thiel M (2008) In situ tests on inducible defenses in Dictyota kunthii and Macrocystis integrifolia (Phaeophyceae) from the Chilean coast. J. Exp. Mar. Biol. Ecol. 354: 28-38. Martinez EA (1996) Micropopulation differentiation in phenol content and susceptibility to herbivory in the Chilean kelp Lessonia nigrescens (Phaeophyta, Laminariales). Hydrobiologia 326/327: 205-211. McLachlan J and Craigie JS (1966) Antialga1 activity of some simple phenols. J. Phycol. 2: l33-l35. Mole S (1994) Trade-offs and constraints in plant-herbivore defense theory: a life-history perspective. Oikos 71: 3-12. Norton TA (1977b) The growth and development of Sargassum muticum (Yendo) Fensholt. J. Exp. Mar. Biol. Ecol. 26: 41-53.  118  Nunez Lopez RA. and Valdez MC (1998) Seasonal variation of seaweed biomass in San Ignacio Lagoon, Baja Californa Sur, Mexico. Bot. Mar. 41: 421-426. Okuda T (1981) Egg liberation in some Japanese Sargassaceae (Phaeophyceae). Xth International Seaweed Symposium, Berlin. Pp. 197-202. Pavia H and Toth GB (2008) Macroalgal models in testing and extending defense theories. In: Amsler, C. D. (ed.), Algal chemical ecology. Springer, pp. 147-172. Pavia H, Toth GB, Aberg P (2002) Optimal defense theory: elasticity analysis as a tool to predict intraplant variation in defenses. Ecology 83: 891-897. Pavia H and Toth GB (2000) Influence of light and nitrogen on the phlorotannin content of the brown seaweeds Ascophyllum nodosum and Fucus vesiculosus. Hydrobiologia 440: 299-305. Pavia H, Toth G, Åberg P (1999) Trade-offs between phlorotannin production and annual growth in the brown seaweed Ascophyllum nodosum. J. Ecol. 87: 761-771. Pfister CA (1992) Costs of reproduction in an intertidal kelp: patterns of allocation and life history consequences. Ecology 73: 1586-1596. Rausher MD (1992) The measurement of selection on quantitative traits: biases due to environmental covariances between traits and fitness. Evolution 46: 616-626. Rousi M, Tahvanainen J, Uotila I (1991) A Mechanism of Resistance to Hare Browsing in Winter-Dormant European White Birch (Betula pendula). Am. Nat. 137: 64-82. Rueness J (1989) Sargassum muticum and other introduced Japanese macroalgae: biological pollution of European coasts. Mar. Poll. Bull. 20: 173-176. Sánchez I and Fernandez C (2005) Impact of the invasive seaweed Sargassum muticum (Phaeophyta) on an intertidal macroalgal assemblage. J. Phycol. 41: 923-930. Scagel RF (1956) Introduction of a Japanese alga, Sargassum muticum, into the northeast Pacific. Fish. Res. Pap. Wash. Dept. Fish. 1: 49-59. Schoenwaelder MEA and Clayton MN (1999) The presence of phenolic compounds in isolated cell walls of brown algae. Phycologia 38: 161-166. Schmitt TM, Hay ME, Lindquist N (1995) Constraints on chemically mediated coevolution: multiple functions for seaweed secondary metabolites. Ecology 76: 107-123.  119  Sieburth JM (1968) The influence of algal antibiosis on the ecology of marine microorganisms. In: Advances in microbiology of the sea. Eds: Droop MR, Wood EJF. Academic Press, New York, pp. 63-94. Simms EL and Rausher MD (1989) The evolution of resistance to herbivory in Ipomoea purpurea. II. Natural selection by insects and costs of resistance. Evolution 43: 573585. Steinberg PD (1984) Algal chemical defense against herbivores: allocation of phenolic compounds in the kelp Alaria marginata. Science 223: 405-406. Steinberg PD (1985) Feeding preferences of Tegula funebralis and chemical defenses of marine brown algae. Ecol. Monogr. 55: 333-49. Steinberg PD (1988) The effects of quantitative and qualitative variation in phenolic compounds on feeding in three species of marine invertebrate herbivores. J. Exp. Mar. Biol. Ecol. 120: 221-37. Steinberg PD (1992) Geographical variation in the interaction between marine herbivores and brown algal secondary metabolites. In: Ecological Roles of Marine Natural Products. Ed: Paul VJ. Cornell University Press, Ithaca, pp. 51-92. Steinberg PD (1995) Seasonal variation in the relationship between growth rate and phlorotannin production in the kelp Ecklonia radiata. Oecologia. 102: 169-173. Steinberg PD and van Altena IA (1992) Tolerance of Australasian marine herbivores to brown algal phlorotannins. Ecol. Monogr. 62: 189-222. Strauss SY, Rudgers JA, Lau JA, Irwin RE (2002) Direct and ecological costs of resistance to herbivory. Trends Ecol. Evol. 17: 278-285. Swanson AK and Druehl LD (2002) Induction, exudation and the UV protective role of kelp phlorotannins. Aquatic Botany 73: 241-253. Targett NM and Arnold TM (1998) Predicting the effects of brown algal phlorotannins on marine herbivores in tropical and temperate oceans. J. Phycol. 34: 195-205. Thomas DN (2002) Seaweeds. The Natural History Museum, London. Toth GB and Pavia H (2007) Induced herbivore resistance in seaweeds: a meta-analysis. J. Ecol. 95: 425-434.  120  Van Alstyne KL (1988) Herbivore grazing increases polyphenolic defenses in the brown alga Fucus distichus. Ecology 69: 655-663. Van Alstyne KL (1995) A comparison of three methods for quantifying brown algal polyphenolic compounds. J. Chem. Ecol. 21: 45-58. Van Alstyne KL, Whitman SL, Ehlig JM (2001) Differences in herbivore preferences, phlorotannin production, and nutritional quality between juvenile and adult tissues from marine brown algae. Mar. Biol. 139: 201-210. van Kleunen M and Schmid B (2003) No evidence for an evolutionary increased competitive ability in an invasive plant. Ecology 84: 2816-2823. Vilá M, Goméz A, Maron JL (2003) Are alien plants more competitive than their native conspecifics? A test using Hypericum perforatum L. Oecologia 137: 211-215. Walker DI and Kendrick GA (1998) Threats to macroalgal diversity: marine habitat destruction and fragmentation. Bot. Mar. 41: 105-112. Wallentinus I (1999) Sargassum muticum. In: Exotics across the ocean. Case histories on introduced species: their general biology, distribution, range expansion and impact. Eds: Gollasch S, Minchin D, Rosenthal H, Voigt M. Logos Verlag, Berlin. Willis AJ and Blossey B (1999) Benign environments do not explain the increased vigour of non-indigenous plants: a cross-continental transplant experiment. Biocontrol Science and Technology 9: 567-577. Yates JC and Peckol P (1993) Effects of nutrient availability and herbivory on polyphenolics in the seaweed Fucus vesiculosus. Ecology 74: 1757-1766. Yoshida T (1983) Japanese species of Sargassum subgenus Bactrophycus (Phaeophyta, Fucales). Journal of the Faculty of Sciences, Hokkaido University Series V. Botany. 13: 99-246.  121  CHAPTER SIX GENERAL CONCLUSION  In chapter two, I investigated the role of native macroalgal diversity in the susceptibility of the community to invasion by Sargassum muticum. I conducted both observational field surveys of the correlation between native diversity and exotic cover, and experimental manipulations of native diversity in constructed 25 x 25 cm communities. Field surveys found higher cover of S. muticum in plots with low native diversity, suggesting a negative relationship between diversity and invasibility at the neighbourhood scale. The experiment found initial cover of S. muticum germlings was highest in plots with greater diversity. Over the duration of the experiment cover of settled germlings increased fastest in the low diversity plots, so that there was a weak negative effect of diversity on final cover of the invader after 77 days. The slope of the relationship reversed over time, with field patterns and experimental results converging at the end of the experiment. These results suggest native diversity has contrasting effects on different stages of invasion. Diversity facilitates invader recruitment of S. muticum but decreases growth and or survivorship. Phenological differences between S. muticum and native macroalgal taxa contributes to invasion success of S. muticum in Barkley Sound. The invader occupies a distinct temporal niche from the native macroalgae; peak S. muticum recruitment occurred in August, which coincided with a sharp decrease in native cover as mature macroalgae senesced, both in our plots and naturally on the shore, enabling population persistence of the exotic. One explanation for the negative  122  association between S. muticum and native diversity is that S. muticum excludes natives through interspecific competition for limiting resources, such as light, space, and nutrients. If the negative correlation between S. muticum cover and native diversity observed in the field is the result of exclusion by S. muticum, then experimental reductions of S. muticum should lead to increased native diversity. In chapter three, I tested the effects of S. muticum on native species richness at different densities by manipulating exotic density in natural communities. I examined effects of light competition by S. muticum, and both light and space competition (net effect), to investigate their effects on the cover and richness of native macroalgae. The effects of S. muticum on native macroalgal richness were both density and time dependent. High cover of S. muticum (40-60%) suppressed native richness through interspecific competition for a combination of limiting resources. Low cover of the exotic had both positive and neutral (no) effects on native richness. The positive and neutral effects of S. muticum at low cover appear to be overwhelmed by negative effects at higher cover, leading to negative associations between native diversity and exotic cover, a pattern reported from unmanipulated plots (chapter one). These results suggest some species benefitted from a modest amount of S. muticum cover, but other species suffered competitive effects. Effects of S. muticum were manifested through resource use, which was mediated through shading smaller, underlying macroalgae by large fronds combined with maintaining substrate through a perennial holdfast, which may be important for preventing the re-establishment of annual macroalgal species and maintaining population persistence of S. muticum. These treatment effects varied with time. These results  123  collectively demonstrate that each phase of the invaders life cycle may yield incremental insights into the overall invasion process, and should be considered in studies of plant and macroalgal invasions. In chapter four, I investigated whether release from co-evolved grazers in British Columbia contributes to invasion success of S. muticum. I conducted grazing trials to test whether native herbivores preferentially graze native, co-evolved macroalgae over nonnative S. muticum, both in the absence of choice and when a choice was offered. I measured the concentrations of polyphenolic compounds in the tissue of the exotic and three native macroalgal conspecifics to determine whether they express different levels of chemical defenses, which could influence grazing by native herbivores. I showed that generalist grazers native to the invaded region showed no clear pattern of preference between native macroalgae and S. muticum when offered a choice. Polyphenolic content of S. muticum was low relative to the native macroalgae tested, with similar levels to some native species, suggesting the exotic allocates relatively few resources to polyphenolic defenses. Release from native enemies does not translate to reduced grazing pressure of S. muticum in the invaded region of British Columbia. In chapter five, I investigated whether S. muticum in non-native regions attained greater sizes and lower levels of chemical defenses than conspecifics in the native region. I compared the size (as a measure of performance) and levels of chemical defenses of S. muticum from both its native and invaded regions. My results suggest that in nonnative regions, S. muticum attains larger sizes with lower levels of defensive compounds than the native region. Coupled with the results of chapter three which show that  124  S. muticum out-competes smaller, underlying native macroalgae through the monopolization of space and light, these results suggest that the increased size of S. muticum in this system confers competitive superiority. I have demonstrated throughout this thesis that invasion of S. muticum is a complex process. Colonization success is a function of phenological and life history traits; native diversity effects on invasion of S. muticum vary with life history stage of the exotic, and effects of S. muticum on the native community vary both with density of the exotic and time. The relatively limited amount of detailed work on macroalgal invasions shows that a given species might have very different impacts in different locales (Johnson and Chapman 2007). Invasion patterns are also variable at different spatial scales. The invasive macroalga Codium fragile ssp. tomentosoides reaches nuisance proportions in the western Atlantic, replacing kelp forests, but is quite rare in the eastern North Atlantic Ocean (Chapman et al. 2002). Macroalgal invasions can also be influenced by abiotic factors in the recipient region, such as disturbance (Britton-Simmons and Abbott 2008), wave exposure (Andrew and Viejo 1998), and resource availability (Stachowicz et al. 2002, Britton-Simmons 2006), and biotic traits such as grazing by native herbivores (Sjøtun et al. 2007) and diversity effects of native macroalgae (White and Shurin 2007). Other influential factors make invasions almost impossible to predict, including evolution in the invaded regions (i.e. increasing host range or adapting to climatic conditions). There is also often a time lag following introduction before an invader proliferates. Alien species can exist in relatively low numbers for decades before becoming invasive (Crooks 2005), or can mediate immediate impacts (i.e. direct consumption of natives by  125  exotics, Savidge 1987). The relative ability of multiple factors to influence invasion success of macroalgal species should be assessed in studies of invasion. Macroalgal introductions are increasing, with more than 120 species currently known, some of which aggressively dominate marine habitats (Mathieson et al. 2008), but our understanding of marine macroalgal invasions is still limited (Grosholz 2002). The mechanisms of macroalgal invasion success are poorly understood, as research has been reactive, following discoveries of introductions and largely focus on high profile species that have caused significant ecological and economic impacts (Schaffelke et al. 2006). Such species include Caulerpa taxifolia in the Mediterranean (Ceccherelli and Cinelli 1998), Codium fragile ssp. tomentosoides in the western Atlantic (Trowbridge 1998), Sargassum muticum in Southern England (Farnham et al. 1973) and Undaria pinnatifida in Australia and New Zealand (Valentine and Johnson 2004). Their diverse and widespread impacts affect all other marine conservation programs including marine protected areas, habitat management (effects of fishing), marine mammal conservation, etc. (Bax et al. 2003). However, when compared to other marine taxonomic groups of marine aliens, macroalgal species are understudied and often overlooked in policy and management initiatives (Johnson and Chapman 2007, McQuaid and Arenas 2009). This may be the result of poor appreciation of the economic and environmental costs of macroalgal invasions, due to their often hidden nature. A massive invasion by a marine macroalga is often much less conspicuous than invasion by a terrestrial plant. Further, a general lack of baseline data on marine macroalgae (Wells et al. 2007) makes identifying the impacts of an invader even more difficult. As both shellfish and macroalgae  126  aquaculture and the aquarium trade are still expanding, macroalgal introductions also have the potential to increase (Padilla and Williams 2004). Although there is increasing evidence supporting the importance of marine macroalgal invasions, there is a need to understand the underlying ecological principles of macroalgal invasions in marine ecosystems (Inderjit et al. 2006). One generality that has emerged from studies of macroalgal invasions is the need for more extant research (Schaffelke and Hewitt 2007, Britton-Simmons 2004, Williams and Smith 2007). Single species studies such as the work presented here comprise the knowledge base on the species and communities invaded, filling gaps in scientific knowledge, which more effectively feed policy. Integrating observations of natural systems across a variety of scales with results from controlled experiments and ecological theory are necessary to understand the invasion process, impacts and options to manage invasions (Johnson and Chapman 2007). Despite the work of biologists worldwide resulting in marine introduced species emerging as a major management issue within the last 20 years (Campbell and Hewitt 1999), management of marine invasives is typically generic, and focuses on managing introduction and transfer pathways rather than on the problems posed by particular species or organisms, such as macroalgae (Doelle et al. 2007). The variability surrounding macroalgal invasions suggests invasive marine macroalgae should be considered on a case-by-case basis (Inderjit et al. 2006). Coupled with the large ecological and economic impacts sometimes associated with macroalgal invaders, more attention to this often overlooked group of marine aliens is warranted, especially as once  127  established, the eradication of macroalgal invaders is considered impossible (Dumont et al. 2004, Nyberg 2007). The acceleration in spread and impacts of introduced macroalgae poses a major challenge for management of marine ecosystems (Valentine et al. 2007), propelling public concern, scientific interest, and calls for policy and management actions (Ruiz and Carlton 2003). Ultimately governments determine responses to invasive species, not scientists or environmental agencies (Johnson and Chapman 2007), and as marine invasions transcend national boundaries, the problem should be considered an international one (Inderjit et al. 2006). Policy surrounding aquatic invasions has developed from global and regional policies including the International Convention on Wetlands (1975), the Convention on Migratory Species of Wild Animals (1983), the Convention on Biodiversity (1993), and the Law of the Sea (1994), but is fragmented at both global and regional levels and does not consider long-term local and global harm (reviewed by Doelle et al. 2007). By example, only Australia, New Zealand, USA, Canada, Switzerland, and Germany have legislation controlling introductions for aquaculture (Johnson and Chapman 2007). Reviews of aquatic invasions policy are given by Simberloff et al. (2005), Doelle et al. (2007), and Bax et al. (2003). Policy and legislation based on better knowledge of marine invasion ecology from detailed investigations of the organisms and mechanisms responsible for success will help prevent these impacts from changing ecosystem structure and function.  128  LITERATURE CITED Andrew NL and Viejo RM (1998) Effects of wave exposure and intraspecifc density on the growth and survivorship of Sargassum muticum (Sargassaceae: Phaeophyta). Eur. J. Phycol. 33: 251-258. Bax N, Williamson A, Aguero M, Gonzalez E, Geeves W (2003) Marine invasive alien species: a threat to global biodiversity. Marine Policy 27: 313-323. Britton-Simmons KH (2006) Functional group diversity, resource preemption and the genesis of invasion resistance in a community of marine algae. Oikos 113: 395-401. Britton-Simmons KH (2004) Direct and indirect effects of the introduced alga Sargassum muticum on benthic, subtidal communities of Washington State, USA. Mar. Ecol. Prog. Ser. 277: 61-78. Britton-Simmons KH and Abbott KC (2008) Short- and long-term effects of disturbance and propagule pressure on a biological invasion. J. Ecol. 96: 68-77. Campbell ML and Hewitt CL (1999) Vectors, shipping and trade. In: Marine biological invasions of Port Phillip Bay. Hobart, Australia. Eds: Hewitt CL, Campbell ML, Thresher RE, Martin RB. CSIRO Marine Research. pp. 45-60. Ceccherelli G, Cinelli F (1998) Habitat effect on spatio-temporal variability in size and density of the introduced alga Caulerpa taxifolia. Mar. Ecol. Prog. Ser. 163: 289294. Chapman AS, Scheibling RE, Chapman ARO (2002) Species introductions and changes in marine vegetation of Atlantic Canada. In: Alien invaders in Canada’s waters, wetlands, and forests. Eds: Claudi R, Nantel P, Muckle-Jeffs E. Natural Resources Canada, Canadian Forest Service Science Branch, Ottawa, pp. 133-148. Doelle M, McConnell ML, VanderZwaag DK (2007) Invasive seaweeds: global and regional law and policy responses. Bot. Mar. 50: 438-450. Dumont HJ, Shiganova TA, Niermann U (2004) The Ctenophores Mnemiopsis leidyi and Beroe in the Ponto-Caspian and other Aquatic Invasions. In: Aquatic Invasions in the Black, Caspian, and Mediterranean Seas. Springer Netherlands 35: pp. 301-305. Farnham WF, Fletcher RL, Irvine LM (1973) Attached Sargassum found in Britain. Nature 243: 231-232.  129  Grosholz ED (2002) Ecological and evolutionary consequences of coastal invasions. Trends Ecol. Evol. 17: 22-27. Inderjit, Chapman D, Ranelletti M, Kaushik S (2006) Invasive Marine Algae: An Ecological Perspective. The Botanical Review 72: 153-178. Johnson CR and Chapman ARO (2007) Seaweed invasions: introduction and scope. Bot. Mar. 50: 321-325. Mathieson AC, Pederson JR, Neefus CD, Dawes CJ, Bray TL (2008) Multiple assessments of introduced seaweeds in the Northwest Atlantic. ICES Journal of Marine Science 65: 730-741. McQuaid CD and Arenas F (2009) Biological Invasions: Insights from Marine Benthic Communities. In: Marine Hard Bottom Communities, Ecological Studies. Ed: Wahl M. Springer-Verlag Berlin Heidelberg, pp. 309-320. Nyberg CD (2007) Introduced marine macroalgae and habitat modifiers-their ecological role and significant attributes. PhD Thesis, Göteborg University. Padilla DK and Williams SL (2004) Beyond ballast water: Aquarium and ornamental trades as sources of invasive species in aquatic ecosystems. Front. Ecol. Environ. 2: 131-38. Ruiz GM and Carlton JT (2003) Invasion vectors: a conceptual framework for management. In: Invasive species: vectors and management strategies. Eds: Ruiz GM and Carlton JT. Island Press, pp 459-504. Savidge JA (1987) Extinction of an island forest avifauna by an introduced snake. Ecology 68: 660-668. Schaffelke B and Hewitt CL (2007) Impacts of introduced seaweeds. Bot. Mar. 50: 397417. Schaffelke B, Smith JE, Hewitt CL (2006) Introduced macroalgae - A growing concern. J. Applied Phycology. 18: 529-541. Simberloff D, Parker IM, Windle PN (2005) Introduced species policy, management, and future research needs. Front. Ecol. Environ. 3: 12-20. Stachowicz JJ, Fried H, Osman RW, Whitlatch RB (2002) Biodiversity, invasion resistance, and marine ecosystem function: reconciling pattern and process. Ecology 83: 2575-2590.  130  Sjøtun K, Eggereide SF, Høisæter T (2007) Grazer-controlled recruitment of the introduced Sargassum muticum (Phaeophyceae, Fucales) in northern Europe. Mar. Ecol. Prog. Ser. 342: 127-138. Trowbridge CD (1998) Ecology of the green macroalga Codium fragile (Suringar) Hariot 1889: Invasive and non-invasive subspecies. Oceanogr. Mar. Biol. Ann. Rev. 36: 164. Valentine JP, Magierowski RH, Johnson CR (2007) Mechanisms of invasion: establishment, spread and persistence of introduced seaweed populations. Bot. Mar. 50: 351-360. Valentine JP, Johnson CR (2004) Establishment of the introduced kelp Undaria pinnatifida following dieback of the native macroalga Phyllospora comosa in Tasmania, Australia. Mar. Freshw. Res. 55: 223-230. Wells E, Wilkinson M, Wood P, Scanlan C (2007) The use of macroalgal species richness and composition on intertidal rocky seashores in the assessment of ecological quality under the European Water Framework Directive. Marine Pollution Bulletin 55: 151-161. White LF and Shurin JB (2007) Diversity effects on invasion vary with life history stage in marine macroalgae. Oikos 116: 1193-1203. Williams SL and Smith JE (2007) Distribution, Taxonomy, and Impacts of Introduced Seaweeds. Annu. Rev. Ecol. Evol. Syst. 38: 327-59.  

Cite

Citation Scheme:

        

Citations by CSL (citeproc-js)

Usage Statistics

Share

Embed

Customize your widget with the following options, then copy and paste the code below into the HTML of your page to embed this item in your website.
                        
                            <div id="ubcOpenCollectionsWidgetDisplay">
                            <script id="ubcOpenCollectionsWidget"
                            src="{[{embed.src}]}"
                            data-item="{[{embed.item}]}"
                            data-collection="{[{embed.collection}]}"
                            data-metadata="{[{embed.showMetadata}]}"
                            data-width="{[{embed.width}]}"
                            data-media="{[{embed.selectedMedia}]}"
                            async >
                            </script>
                            </div>
                        
                    
IIIF logo Our image viewer uses the IIIF 2.0 standard. To load this item in other compatible viewers, use this url:
https://iiif.library.ubc.ca/presentation/dsp.24.1-0069842/manifest

Comment

Related Items