Invasive marine macroalgae - community invasibility, invasion process, and their ecological role in the ecosystem functioning.

Invasive marine macroalgae – understanding community invasibility, invasion process, and their ecological role in the ecosystem functioning Fátima de Carvalho Vaz Pinto Tese de doutoramento em Ciências do Mar e do Ambiente 2012 Fátima de Carvalho Vaz Pinto Invasive marine macroalgae – understanding community invasibility, invasion process, and their ecological role in the ecosystem functioning Tese de Candidatura ao grau de Doutor em Ciências do Mar e do Ambiente (Especialidade em: Oceanografia e Ecossistemas Marinhos) submetida ao Instituto de Ciências Biomédicas de Abel Salazar da Universidade do Porto. Orientador – Doutor Francisco Arenas Parra Categoria – Investigador Afiliação – Laboratório de Biodiversidade Costeira, Centro Interdisciplinar de Investigação Marinha e Ambiental, Porto. Co-orientadora – Doutora Celia Olabarria Uzquiano Categoria – Professora Afiliação – Departamento de Ecoloxía e Bioloxía Animal, Faculdade de Bioloxía, Universidade de Vigo, Espanha. LEGAL DETAILS In compliance with what is stated in Decret-Law nº 216/92 of October 13th, it is hereby declared that the author of this thesis participated in the creation and execution of the experimental work leading to the results shown, as well as in their interpretation and the writing of respective manuscripts. This thesis also includes three scientific papers published in international journals and two submitted articles originating from part of the results obtained in the experimental work referenced to as: Vaz-Pinto F., Olabarria C., Arenas F. Propagule pressure and functional diversity: interactive effects on a macroalgal invasion process. Marine Ecology Progress Series (in press) Vaz-Pinto F., Olabarria C., Arenas F. Role of top-down and bottom-up forces on the invasibility of intertidal macroalgal assemblages. Journal of Sea Research (in press) Vaz-Pinto F., Olabarria C., Arenas F. Invasion of Sargassum muticum (Yendo) Fensholt: impacts on ecosystem functioning of marine macroalgal assemblages. Journal of Phycology (under review) Vaz-Pinto F., Martínez B., Olabarria C., Arenas F. The native Cystoseira humilis versus the invasive Sargassum muticum: competitive interactions between two fucoid algae. (in preparation) Vaz-Pinto F., Olabarria C., Gestoso I., Cacabelos E., Incera M., Arenas F. Functional diversity and climate change: effects on the invasibility of macroalgal assemblages. Biological Invasions (accepted) Table of Contents Acknowledgements _______________________________________________________ i Figures index ___________________________________________________________ iii Tables index ___________________________________________________________vii Summary ___________________________________________________________ xi Resumo __________________________________________________________ xv Part I - General introduction Chapter 1 – General introduction ____________________________________________ 3 1.1 Invasion ecology ___________________________________________________ 4 1.2. The invasion process _______________________________________________ 5 1.2.1. The arrival of introduced species ___________________________________ 6 1.2.2. The success of invasions _________________________________________ 8 1.3. Introduced macroalgae and known impacts _____________________________ 11 1.4. NIS and ecosystem functioning in marine ecosystems _____________________ 15 1.5. Global environmental change: new scenarios ___________________________ 17 1.6. Model invader species & Model system ________________________________ 19 1.6.1. Sargassum muticum: range distribution & general aspects of its biology ___ 19 1.6.2. Life in intertidal rock pools _______________________________________ 22 1.6.3. The use of synthetic macroalgal assemblages ________________________ 24 1.7. Objectives and structure of the thesis __________________________________ 25 1.8. References ______________________________________________________ 26 Part II - The process of macroalgal invasions Chapter 2 – Propagule pressure and functional diversity: interactive effects on a macroalgal invasion process ___________________________________ 43 2.1. Abstract _________________________________________________________ 46 2.2. Introduction ______________________________________________________ 47 2.3. Materials and Methods _____________________________________________ 48 2.3.1. Study area ___________________________________________________ 48 2.3.2. The invader, Sargassum muticum _________________________________ 49 2.3.3. Macroalgal assemblages and experimental design ____________________ 49 2.3.4. Invasion procedure _____________________________________________ 51 2.3.5. Measurements of invasibility and sampling procedure __________________ 52 2.3.6. Statistical analyses _____________________________________________ 53 2.4. Results _________________________________________________________ 54 2.4.1. Resources availability ___________________________________________ 54 2.4.2. Invasion success at different invasion stages ________________________ 56 2.4.3. Instantaneous mortality rates _____________________________________ 56 2.5. Discussion _______________________________________________________ 59 2.6. Acknowledgements ________________________________________________ 62 2.7. References ______________________________________________________ 63 Chapter 3 – Role of top-down and botton-up forces on the invasibility of intertidal macroalgal assemblages ______________________________________ 67 3.1. Abstract _________________________________________________________ 70 3.2. Introduction ______________________________________________________ 71 3.3. Material and methods ______________________________________________ 73 3.3.1. Study site ____________________________________________________ 73 3.3.2. Experimental design and Sampling procedure ________________________ 74 3.3.3. Statistical analysis _____________________________________________ 76 3.3.4. Effectiveness of treatments ______________________________________ 77 3.4. Results _________________________________________________________ 78 3.4.1. Early recruitment of Sargassum muticum ____________________________ 78 3.4.2. Invasibility of macroalgal assemblages _____________________________ 79 3.4.3. Top-down and bottom-up regulation on macroalgal assemblage structure __ 82 3.5. Discussion _______________________________________________________ 86 3.6. Acknowledgments _________________________________________________ 89 3.7. References ______________________________________________________ 90 3.8. Suplementary data ________________________________________________ 94 Part III - Biotic interactions & Ecosystem functioning Chapter 4 – Neighbourhood competition in coexisting macroalgae: the native Cystoseira humilis vs the invasive Sargassum muticum ______________ 99 4.1. Abstract ________________________________________________________ 102 4.2. Introduction _____________________________________________________ 103 4.3. Materials and methods ____________________________________________ 105 4.3.1. Field study site _______________________________________________ 105 4.3.2. Sargassum muticum and Cystoseira humilis ________________________ 106 4.3.3. Nutrients over a tidal cycle ______________________________________ 108 4.3.4. Nutrient uptake experiment in the laboratory ________________________ 109 4.3.5. Statistical analysis ____________________________________________ 110 4.4. Results ________________________________________________________ 111 4.4.1. Algal growth rate and tissue C and N ______________________________ 111 4.4.2. Nutrients over a tidal cycle ______________________________________ 112 4.4.3. Nutrient uptake _______________________________________________ 116 4.5. Discussion ______________________________________________________ 119 4.6. Acknowledgements _______________________________________________ 122 4.7. References _____________________________________________________ 123 Chapter 5 – Ecosystem functioning impacts of the invasive seaweed Sargassum muticum (Phaeophyta) _______________________________________ 129 5.1. Abstract ________________________________________________________ 132 5.2. Introduction _____________________________________________________ 133 5.3. Materials and methods ____________________________________________ 135 5.3.1. Macroalgal assemblages _______________________________________ 135 5.3.2. Incubation procedures _________________________________________ 137 5.3.3. Ecosystem functioning surrogates ________________________________ 138 5.3.4. Data analysis ________________________________________________ 138 5.4. Results ________________________________________________________ 139 5.4.1. November 2010 ______________________________________________ 139 5.4.2. May 2011 ___________________________________________________ 140 5.5. Discussion ______________________________________________________ 146 5.6. Acknowledgments ________________________________________________ 151 5.7. References _____________________________________________________ 152 Part IV - NIS & Climate change Chapter 6 – Functional diversity and climate change: effects on the invasibility of macroalgal assemblages _____________________________________ 161 6.1. Abstract ________________________________________________________ 164 6.2. Introduction _____________________________________________________ 165 6.3. Methods _______________________________________________________ 168 6.3.1. Laboratory mesocosm system ___________________________________ 168 6.3.2. Environmental parameters in tanks _______________________________ 168 6.3.3. Synthetic assemblages _________________________________________ 169 6.3.4. Invasion of assemblages _______________________________________ 170 6.3.5. Development of assemblages in the field ___________________________ 171 6.3.6. Statistical analyses ____________________________________________ 171 6.4. Results ________________________________________________________ 173 6.4.1. Environmental parameters in tanks _______________________________ 173 6.4.2. Resistance to invasion and invasiveness of S. muticum _______________ 173 6.4.3. Development of assemblages in the field ___________________________ 178 6.5. Discussion ______________________________________________________ 180 6.6. Acknowledgments ________________________________________________ 183 6.7. References _____________________________________________________ 184 vi respectively); b) Identity effect; adjusted mean (+SE, n = 16). See Fig. 6.1 for functional diversity abbreviations. _________________________________________ 175 Fig. 6. 3. Mean (+SE, n = 16) percentage survivorship of Sargassum muticum settled germlings. Interaction term Temperature x pCO2 at a) 3 days after experimental conditions were imposed; b) 10 days after experimental conditions were imposed. ________________________________________________________ 177 Fig. 6. 4. Mean (+SE, n = 16) length (mm) of Sargassum muticum settled germlings. Interaction term Temperature x pCO2. ______________________________________ 178 Fig. 6. 5. Mean (+SE, n= 8) percentage cover of Ulva spp. and Sargassum muticum within Functional diversity in each pCO2 experimental treatments. See Fig. 6.1 for functional diversity abbreviations. _________________________________________ 179 Fig. 6. 6. Mean (+SE, n = 16) number of Sargassum muticum recruits (64 cm2) at each temperature and pCO2 treatment conditions. ____________________________ 179 vii Tables index Table 1. 1. List of the non-indigenous macroalgae from the northwest Iberian Peninsula (Araújo et al., 2009; Bárbara et al., 2005; Bárbara et al., 2008). NP, Northern Portugal; G, Galicia. _____________________________________________ 14 Table 2. 1. Sargassum muticum. Analysis of deviance revealed by generalized linear models fitted to the explanatory variables Propagule pressure (PP), Functional diversity (FD), Photosynthetic active radiation (PAR) and Space availability. Variable’s parameters described for the minimum adequate model for the response variable a) number of S. muticum germlings, b) number of S. muticum recruits and c) presence-absence data for survivorship of S. muticum to adult stage. ____________ 58 Table 2. 2. Sargassum muticum. Results of the three-way ANOVA testing for the effects of Propagule pressure (PP), Time (T) and Functional diversity (FD) on rates of instantaneous mortality. * Two missing replicates were replaced by the average value of the three replicates remaining in the treatment, and 2 df subtracted from Residuals. ____________________________________________________________ 59 Table 3. 1. Results from Generalized Linear Models (GLMs) for a) the length, b) the biomass and c) the number of S. muticum recruits at the end of the experimental period. Model fitted to the explanatory variables Nutrients, Herbivory intensity, Number of native taxa (native), Number of non-indigenous species (NIS), Percentage cover of crustose algae (Crustose), Percentage cover of leathery algae (Leathery), Percentage cover of corticated algae (Corticated), Percentage cover of articulated calcareous algae (Artcalcareous), Percentage cover of filamentous and foliose algae (Ephemeral) and biologically meaningful interactions. The NULL (intercept only) model is y ~1, reflecting no dependence of y on the explanatory variables. Akaike information criterion (AIC). 1, 2, 3, indicate the family of distribution applied to the full model. _________________________________________________ 81 Table 3. 2. Summary of similarity percentage (SIMPER) analysis. Differences in average percentage cover of species contributing to dissimilarities between initial and final sampling dates (February 2010 and March 2011, respectively). A cut-off of a cumulative percentage dissimilarity of 60 was applied. ________________________ 82 Table 3. 3. Analysis of Wald statistic table from Generalized Estimating Equations (GEE) examining the main and interactive effects of Nutrient and Herbivory intensity on number of native species, number of non-indigenous species (NIS), Percentage viii cover of crustose algae, Percentage cover of leathery algae, Percentage cover of corticated algae, Percentage cover of articulated calcareous algae, Percentage cover of ephemeral filamentous and foliose algae (Ephemeral). Data from February 2010 to March 2011. ____________________________________________________ 84 Table 4. 1. Two-way analysis of variance (ANOVA) for growth rate (g DW month-1) of the focal species, a) S. muticum (n = 5) and b) C. humilis (n = 6). Density of macroalgal assemblages and neighbour identity were fixed factors. ______________ 112 Table 4. 2. Two-way analysis of variance (ANOVA) for C:N ratio of the focal species, a) S. muticum (n = 5) and b) C. humilis (n = 6). Density of macroalgal assemblages and neighbour identity were fixed factors. ___________________________________ 112 Table 4. 3. Estimations (standard error) of the uptake kinetic parameters for S. muticum and C. humilis using either linear or Michaelis–Menten (MM) functions. Vmax is expressed in µmol g DW−1 h−1 and Ks in µM. ___________________________ 118 Table 5. 1. November 2010. Linear model analysis testing the main and interactive effects of invasion of Sargassum muticum and a) species richness or b) species evenness, on the functional responses of macroalgal assemblages (per 64 cm2), i.e. Respiration, maximum net primary productivity (Max NPP), and alpha. *P < 0.05; **P < 0.01; ***P < 0.001; ns: not significant. _________________________________ 140 Table 5. 2. May 2011. Linear model analysis testing the main and interactive effects of invasion of Sargassum muticum and a) species richness or b) species evenness, on the functional responses of macroalgal assemblages (per 64 cm2), i.e. Respiration, maximum net primary productivity (Max NPP) and alpha. *P < 0.05; **P < 0.01; ***P < 0.001; ns: not significant. ____________________________________ 142 Table 5. 3. May 2011. Linear model analysis testing the main and interactive effects of invasion of Sargassum muticum and a) species richness or b) species evenness, on the functional responses of macroalgal assemblages, i.e. Respiration, maximum net primary productivity (Max NPP), and alpha. Data was corrected by macroalgal biomass (grams dry weight). (·)P < 0.06; *P < 0.05; **P < 0.01; ***P < 0.001; ns: not significant. ___________________________________________________________ 144 Table 6. 1. Two-way analysis of covariance (ANCOVA) for the number of Sargassum muticum germlings. Functional diversity was a fixed factor and Tank was a random factor, Biomass was a covariate (n = 4). Homogeneous variances. ___ 174 ix Table 6. 2. Four-way ANOVA for survivorship of Sargassum muticum settlers at a) 3 days and b) 10 days after experimental conditions were imposed (n = 4). Temperature, pCO2 and Functional diversity were fixed factors and Tank was a random factor. Homogeneous variances. ___________________________________ 176 Table 6. 3. Four-way analysis of variance (ANOVA) for length (mm) of Sargassum muticum settlers (n = 4). Temperature, pCO2 and Functional Diversity were fixed factors and Tank was a random factor. Variances heterogeneous (Cochran’s test, C = 0.2382, P < 0.01). Significance level at P = 0.01. ____________________________ 177 xi Summary pecies invasions have increased to unprecedented rates due to globalization of human activities around the world. High rates of biodiversity loss and biological invasions put ecosystems under enormous stress. Hence, it is critical not only to understand how the loss of a species influences the stability and function of the ecosystems we rely on, but also the implications of the addition of non-indigenous species. Life in today’s seas is changing in an alarming rate. In the present thesis I examine different aspects of the processes governing biological invasions in coastal habitats, using the invasive brown macroalga Sargassum muticum Yendo (Fensholt) as a model species. Firstly, I give a general overview of invasion ecology in Chapter 1, highlighting registered impacts of non-indigenous species (NIS). The following two chapters refer to the interaction between the native ecosystem and the invasive macroalgae. Chapter 2 investigated the interactive effects of functional diversity and propagule pressure on the recruitment success of S. muticum through different invasion stages, i.e., settlement, recruitment and colonisation. It is widely accepted that high diverse assemblages are more resistant to invasions because of a more complete or efficient utilization of resources. Invasion success is ultimately determined by the interaction of NIS arrival (propagule pressure) and the availability of resources needed by the NIS to establish. The initial hypothesis was that there is a positive relationship between invasibility of macroalgal assemblages and propagule pressure of the invader, but its shape is determined by the diversity of the recipient assemblages. I found that invasion success was affected by the interaction of environmental factors that covary with species diversity and identity at different invasion stages. Chapter 3 examined how the availability of nutrients and grazers operate interactively to determine invasion success. Consumers and nutrient availability play a crucial role in the regulation of intertidal habitats. The theory of fluctuating resource availability suggests that the invasibility of a community changes as the amount of unused resources fluctuates. The hypothesis was that a greater availability of resources through an increase in nutrients concentration or an increase in herbivory would affect the invasibility of macroalgal assemblages. Results showed that early recruitment of S. muticum was enhanced by low nutrient enrichment, but no effect of grazers was observed. In contrast, at the end of the experiment (9-months after invasion) top-down control, together with the number of non-indigenous species and the percentage cover of ephemerals, was a significant predictor for the invasion success of S. muticum. S xii Species interactions are fundamental in shaping community structure and dynamics. Thus, Chapter 4 and 5 linked the presence of S. muticum with competitive interactions and ecosystem-level impacts on native macroalgal assemblages. Chapter 4 explored the mechanisms involved in the competition between two functionally similar fucoid species, the invasive S. muticum and the native macroalga Cystoseira humilis. Specifically, this chapter intends to explore the effects of neighbors in local interactions. We expected that species would display different resource allocations when subjected to intraversus interspecific competition and that S. muticum would perform better than C. humilis when in interaction. Our results showed a negative density-dependent effect on biomass production for S. muticum, independently of neighbour identity. Contrary to our expectations, the nutrient uptake experiment revealed a greater competitive response of the native species compared to S. muticum. Moreover, our short-term study C. humilis showed a greater increased in biomass when in mixed assemblages with S. muticum than in monospecific assemblages, suggesting that S. muticum neighbours display a positive interaction with C. humilis when in mixed assemblages. Chapter 5 compared the dynamics of respiration and production in native macroalgal assemblages and assemblages invaded by S. muticum, to investigate ecosystem-level impacts of macroalgal invasions. Native species with a long history of co-evolution are expected to partition resources among them and promote ecosystem functioning throughout resource use complementary effects. In contrast, newly introduced species, probably enhance ecosystem functioning by sampling effects where the influence of the invader is well beyond its proportion. This chapter aimed to test empirically this hypothesis. I found that S. muticum altered ecosystem functioning by increasing assemblage’s productivity, although this change was seasonal. Results imply that S. muticum was more productive on an area basis than on a biomass-specific basis, suggesting that the impact of S. muticum may be related to its high dominance in the invaded assemblages, varying drastically with the season. Finally, global change drivers are expected to induce higher susceptibility of marine communities to invasions. This hypothesis was tested with intertidal macroalgal assemblages. Chapter 6 tested the effects of disturbance by climate change-related drivers on early survivorship of S. muticum germlings and the legacy effects over recruitment success. Specifically, this chapter examines the combined effects of increase temperature and CO2 partial pressure and functional diversity in shaping the invasion success of S. muticum. Our results showed an interaction of temperature and pCO2 in the early survivorship of settled germlings (3 and 10 days after invasion). Moreover, after 6 months in the field, legacy effects of laboratory treatments remained, with S. muticum xiii reaching higher cover in most assemblages previously subjected to ambient pCO2, but ephemeral green algae appearing disproportionately after elevated-pCO2 treatment. Future climatic scenarios of increasing temperature and pCO2 may not be beneficial to S. muticum. Lastly, Chapter 7 provides an overview of the invasion success of S. muticum and its implications for biodiversity ecosystem function research in the current and future scenario of global change. This thesis contributed to a wider understanding of marine macroalgal invasions in general and of the ecology and physiology of the invasive brown alga S. muticum in particular. xv Resumo introdução de espécies não indígenas tem vindo a aumentar com a intensificação da globalização do planeta. A redução global da biodiversidade e o aumento das invasões biológicas levam a alterações e aumento do stress nos ecossistemas. Neste contexto, torna-se critico não só entender qual o efeito da perda de espécies na estabilidade e funcionamento dos ecossistemas, mas também quais as implicações da introdução de espécies não indígenas. Especificamente, a vida nos nossos oceanos está a mudar a um ritmo alarmante. Nesta tese de doutoramento foram analisados diferentes aspetos do processo de invasões biológicas em habitats costeiros, usando a alga castanha invasora Sargassum muticum Yendo (Fensholt) como espécie modelo. Inicialmente é feita uma revisão sobre a ecologia das invasões biológicas no Capítulo 1, descrevendo alguns impactos conhecidos de espécies não indígenas. Os dois capítulos seguintes integram a relação entre os ecossistemas nativos e a as macroalgas invasoras. O Capítulo 2 investiga a interação entre a diversidade funcional e a pressão de propágulos no recrutamento de S. muticum ao longo dos vários estágios do processo de invasão: assentamento, recrutamento e colonização. É amplamente aceite que comunidades com maior diversidade são mais resistentes a invasões devido a uma mais completa e eficiente utilização dos recursos disponíveis. O sucesso de uma invasão biológica é determinado pela interação entre a chegada dos organismos não nativos (pressão de propágulos) e a disponibilidade de recursos necessários para o estabelecimento da espécie não nativa. A hipótese inicial foi que havia uma relação positiva entre a invasibilidade das comunidades de macroalgas e a pressão de propágulos da espécie não nativa, mas o tipo de relação é determinada pela diversidade da comunidade recetora. Os resultados mostram que o sucesso da invasão foi afetado pela interação entre fatores ambientais que co-variam com a riqueza específica e identidade das espécies em diferentes estágios da invasão. O Capítulo 3 examina como a disponibilidade de nutrientes e a herbivoria interagem no processo de invasão. Estes dois fatores têm um papel crucial na regulação dos habitats intertidais. A teoria da disponibilidade de recursos (Theory of fluctuating resource availability) sugere que a invasibilidade de uma comunidade varia com a disponibilidade de recursos não usados. A nossa hipótese foi que uma maior disponibilidade de recursos através de um aumento da concentração de nutrientes ou um aumento na herbivoria afetaria a invasibilidade das comunidades de macroalgas. Os resultados revelaram que o recrutamento inicial de S. muticum aumentou com o baixo enriquecimento de nutrientes, mas não foi encontrado efeito de herbivoria. Contrariamente, no final da experiência (9 meses após a invasão) o A Chapter 1 4 Preface This thesis examines different aspects of the invasion success of the macroalga Sargassum muticum (Yendo) Fensholt (Phaeophyta, Fucales). In this general introduction, I define and describe the process of biological invasions, the mechanisms involved and summarise the known impacts of invasion, particularly of introduced marine macroalgae. Then, I make a general overview of interactive effects of global climate change, specifying vulnerability and adaptations of non-indigenous species to those changes. This introduction also provides background information about S. muticum: its status, distribution, invasion history and impacts in native assemblages. The thesis chapters have been written as journal articles. Thus, to allow autonomy of each chapter, some repetition of introductory and methodological information was necessary. The general introduction chapter may be seen as an overview where I describe the invasion process, our model organism, S. muticum, and our model system, intertidal rock pools. 1.1 Invasion ecology ince the 19th century, pioneering naturalists – namely Charles Darwin, Alphonse De Candolle, Joseph Hooker and Charles Lyell – have mentioned invasive species in their writings (Ludsin and Wolfe, 2001; Richardson and Pysek, 2007), although then invasive species were not perceived as a threat to global biodiversity. Charles Elton book The Ecology of Invasions by Animals and Plants (1958) is considered as the seminal work of invasion ecology, classified as one of the first to address biological invaders as key drivers of ecosystem change. More than 50 years after being released, Elton’s book is the most cited source in invasion biology (Richardson and Pysek, 2008). Elton’s 1958 book most notable arguments are related to the diversity-stability and diversity-invasibility relationships (Richardson and Pysek, 2007). Additionally, its chapter 5 entitled “Changes in the Sea” was the first global overview of marine invasions. Invasion ecology encompasses the study of the human-mediated introduction of organisms, especially introductions to areas outside the natural range expansions of given organisms (Carlton, 1996b; Mack et al., 2000). Over the last decades the study of patterns and processes behind biological invasions, their interactions with resident organisms in recipient locations and the consideration of costs and benefits from their presence have grown as research topics (Carlton, 1996b; Richardson et al., 2010; Stachowicz et al., 1999). In the beginning, terrestrial and freshwater systems were the most studied S General introduction 5 systems, but during the last two decades marine systems have been studied intensely (Grosholz, 2002). 1.2. The invasion process Biological invasions occur when a species enters and spreads into areas beyond its natural range of distribution (Vermeij, 1991). Beyond the changes in species range, which occur constantly in natural communities (Vermeij, 1991), it is the accelerating rate and the magnitude of those changes that pose a threat to global biodiversity (Lodge, 1993). Despite many non-indigenous species (NIS) can arrive into new locations, the development of NIS populations is a dynamic, very complex process and consists of several transitional stages (Williamson, 2006). There is, however, little agreement about how to name these stages, and how many are there. In this research, the definitions were used following two different authors (Colautti and MacIsaac, 2004; Richardson et al., 2000). In particular, Richardson et al. (2000) critically defined the minimum key terms linked to the process of invasion, i.e. introduction, naturalization and invasion. Colautti and MacIssac (2004) suggested a well defined framework linking the invasion process with different filters that have to be overcome to arrive to subsequent stages. Here, the process of invasion is described as a set of stages, critically defined as follows: - Non-indigenous species (NIS) are defined as individual(s) of a species that enters a given area outside of its historic or native range, due to human activity (synonyms: alien species, non-native species) (Richardson et al., 2000); - Introduction (stage II) describes the arrival of individuals (or propagules) to a given area outside of its historic or native range after being transported through human activities. Survival of introduced NIS to adult individuals, establishment, is set as stage III. Stage 0 and stage I correspond to residency in a potential donor region and transportation, respectively. - Naturalization represents the stage where NIS have a regular reproduction and develop sustainable widespread populations (Stage IVa) that do not necessarily invade recipient ecosystems (stage IVb). - Invasion requires naturalized NIS to reproduce abundant reproductive offspring also in areas distant from parent plants and thus may be described as widespread and dominant (stage V). Chapter 1 6 Thus, during the process of invasion, NIS must overcome several ecological barriers at different stages of invasion before they are able to inflict ecological or economic harm (Fig. 1.1). Abiotic conditions may be the main ecological barriers NIS will encounter. Traits such as geographic origin, native range extent and dispersal ability may be particularly related to successful transportation (Goodwin et al., 1999; Theoharides and Dukes, 2007). Additionally, high propagule pressure has been suggested to increase the likelihood that species survive transport (Kolar and Lodge, 2001; Lockwood et al., 2005). Repeated introduction events are also likely to increase the likelihood of a species to arrive at the new location, increasing invasion risks (Drake and Lodge, 2006). Upon arrival, abiotic conditions at the recipient ecosystem, biotic interactions and demographic processes may interfere with NIS establishment. The process of invasion will then be completed by the spread/integration of the NIS within the recipient ecosystem, subjected to ecological and evolutionary processes (e.g. competition, predation, hybridization) of the new ecosystem. Understanding which factors and to what extent they influence invasion success (Kolar and Lodge, 2001; White and Shurin, 2007) through the different stages is of extreme importance in invasion ecology (Theoharides and Dukes, 2007). This information will allow us to disentangle the interaction between what we know as barriers to invasion and the process of invasion (i.e. number of introduction events, species traits, and ecosystem characteristics). 1.2.1. The arrival of introduced species Geographic barriers such as oceans, mountains, ice sheets, and river valleys represent boundaries to the movement of individuals between populations of the same species. However, since humans have had the ability to disperse across continents they have accelerated the invasion process by overcoming biogeographical barriers to dispersal (Vermeij, 1991). Associated with human-mediated introductions is its unpredictable character, independent of the natural barriers of space and time. As such, it is estimated that a marine species will be introduced to a new environment on a weekly to daily basis (NOAA, 2012). Possible vectors of marine invasions are shipping, canals, aquaculture, fisheries, ornamental species and live food, marine leisure and research (Minchin et al., 2009). Identifying how a species arrived, however, is not always possible. Nevertheless, shipping and aquaculture are recognized as the most important vectors involved in marine species introductions (Minchin, 2007; Ruiz et al., 2000; Streftaris et al., 2005). General introduction 7 Fig. 1. 1. Schematic integrated approach to define the different stages of the invasion process. Potential invaders begin as residents propagules from a donor region (stage 0). The invasion process is described from transportation (stage I) to invasion (stage V) following Colautti and MacIsaac (2004). Potential invaders have to pass through a series of filters that may stop the transition to subsequent stages. We present a set of characteristics that may favour invaders in a particular stage since introduction. Chapter 1 8 Marine sessile species may be transported by ships both attached to the hull and inside the ship in solid or water ballast (Minchin et al., 2009). In particular, the transport of ballast water provides opportunities for the spread of entire assemblages of marine species (Carlton and Geller, 1993). As an example, it has been estimated that 10000 marine species are transported in ballast water daily (Carlton, 1999). In addition, the accumulation of sediments in the bottom of ballast tanks provides a refuge for infaunal species (Minchin et al., 2009). However, fouling of ship hulls has been recognized as a more important vector for seaweeds than ballast water (Johnson and Chapman, 2007). Aquaculture-related activities are the second most frequent vector of marine introductions (Minchin, 2007). Many non-native species of fish, invertebrates and seaweeds are cultivated worldwide. For instance, the Japanese oyster, Crassostrea gigas, and the Japanese seaweed, Undaria pinnatifida, are examples of introduced species with commercial value. Additionally, aquaculture species may unintentionally introduce associated biota, including pests, parasites and diseases (Minchin et al., 2009). This last pathway accounted for 15% of the marine invasions reported by Ruiz et al. (2000) in North America and 19% of the marine invasions in Europe (Streftaris et al., 2005). A recent review from marine benthic communities has found that the introduction of macroalgae is mainly related to aquaculture while shipping is described as the major introduction vector for the animals (McQuaid and Arenas, 2009). 1.2.2. The success of invasions Particularly interesting in invasion ecology is the dramatically greater abundance of some NIS in new ecosystems compared to where they are native. Thus, after a general description of the invasion process, we now intend to disentangle the mechanism behind an invasion success and the maintenance of biodiversity. Overall, three factors are usually cited as determining the fate of invasions: the biology of the introduced species (species invasiveness), number and frequency of introductions (propagule pressure) and the susceptibility of the native community to invasion (community invasibility) (Lonsdale, 1999). Nonetheless, invasion studies are usually focused on only one of these factors (Lonsdale, 1999). Understanding the mechanisms behind resistance to biological invasions is of particular interest due to the General introduction 9 fact that NIS have produced large ecological and economic impacts (Mooney and Drake, 1989; Scalera, 2010; Vitousek et al., 1997). Taken as a whole, there are many and diverse theories of invasion success (Mitchell et al., 2006; Sakai et al., 2001), although most share the key prediction that successful invaders must be fundamentally different from native species (e.g. Daehler, 2003; Titman, 1976). According to the Diversity Resistance Hypothesis, generally credited to Charles Elton (1958), species diversity enhances resistance to biological invasions. This Hypothesis suggests that high diversity communities maintain occupied most available niches compared to species-poor communities, believed to contain more unoccupied niches. Through his studies of native and exotic species on oceanic islands, Elton (1958) showed evidence that richer communities were more stable, less vulnerable to disturbances and thus less likely to be invaded. However, an “invasion paradox” has emerged as researchers describe that at regional scales species-rich communities may be more susceptible to invasions (e.g. Lonsdale, 1999; Stohlgren et al., 1999), emphasizing the role of spatial scale in the diversity-invasibility relationship (Fridley et al., 2007; Levine, 2000). Because competition is regarded as a key biotic resistance mechanism acting upon NIS arrival, special attention has been given to the role of natural enemies, i.e. predators and parasites, in the invasion process. Invasion success of NIS has been related to the Enemy Release Hypothesis, which states that the release from natural enemies drives the increase in distribution and abundance of NIS in its new range (Keane and Crawley, 2002), by being unrecognized or unpalatable to native enemies. On the other hand, a key consideration is whether NIS are less vulnerable to native enemies than are native species in its new range. Both hypotheses are not, however, mutually exclusive. Upon arrival into a new region, NIS left behind many of its natural enemies (enemy release), but may also encounter new generalists enemies for which no defenses had been developed (biotic resistance) (Morrison and Hay, 2011; Verhoeven et al., 2009). In the marine realm, however, because marine herbivores are often generalists (Hay, 1991; Morrison and Hay, 2011), the Enemy Release Hypothesis might be of limited use to explain invasion success of marine algae. The Fluctuating Resources Theory (Davis et al., 2000) emerges as an integration of several existing hypotheses regarding community invasibility. This theory suggests that the invasibility of a community changes as the amount of unused resources fluctuates, i.e. a community becomes more susceptible to invasion whenever resource availability (the difference between gross resource supply and resource uptake) increases (Davis Chapter 1 10 et al., 2000). Availability of resources may fluctuate either due to a pulse in resource supply, a decline in resource uptake, or both which consequently will decrease competition for that resource. Thus, the susceptibility of a community to invasion is not a fixed property and fluctuates in the same way other community properties fluctuate (e.g. nutrient availability, predation intensity, space availability). Experimental studies reinforce this theory, emphasizing resource fluctuation and competition as the proposed mechanisms affecting invasibility (Davis and Pelsor, 2001) and not species richness (Dunstan and Johnson, 2007). Moreover, manipulative field experiments using macroalgal functional groups found that resource availability was mediated by algal species identity (and not species richness), suggesting that a functional group approach may better describe the different mechanisms of species coexistence acting within a community (Arenas et al., 2006b). In addition to biotic and abiotic factors, the relative importance of propagule pressure has been largely acknowledged (see Simberloff, 2009 for a review). Contrasting to speciesand community-level particular traits, which remain constant across repeated introductions events separated by some relatively small timeframe, propagule pressure is characteristic of a particular introduction event, differing between introduction events (Lockwood et al., 2005). Thus, propagule pressure could explain some of the idiosyncratic nature of introduction success. The invasibility of an ecosystem is the outcome of several factors, from ecological interactions between NIS and native species to the region’s climate and related interactions, the disturbance regime, unutilized resource availability and propagule supply (Davis et al., 2005; Lonsdale, 1999). Thus, despite many hypotheses have been proposed to explain why some ecosystems are more susceptible to invasion than others, it is unlikely that any single hypothesis will apply to all different environments. Whether biotic resistance is mainly linked with community diversity and associated resource use (e.g. Stachowicz and Byrnes, 2006; Tilman, 1999) or other biotic processes such as competition and facilitation (Crawley et al., 1999) is an open debate in ecology (Occhipinti-Ambrogi, 2007). Moreover, recent work has also suggested a varying effect of diversity with life history stage of invader (Britton-Simmons, 2006; White and Shurin, 2007). For example, an experimental study testing the invasion success of the introduced brown macroalga Sargassum muticum found that although native algal richness enhanced initial settlement of invaders, an overall negative effect of species richness was observed for the invader abundance (White and Shurin, 2007). Hence, the study of the invasion General introduction 11 process should focus on the mechanisms affecting each stage of the invasion. Upon arrival, the processes that determine species distribution and abundance should be similar for both native and NIS (Davis et al., 2000) where stable coexistence requires competitors to differ in their niches. Thus, species with similar patterns of resource use and habitat requirements (functionally similar species) are expected to increase interspecific competition. Classic niche theory, however, does not explain how NIS without niche differences establish nor why invasive species produce large impacts when their establishment depend on different niche utilization than the native species (MacDougall et al., 2009). For example, it has been described a high invasion success of nitrogen (N) fixers in habitats with only non-N-fixers (Vitousek et al., 1987), supporting the empty niche theory. However, the use of two different niches should facilitate coexistence and not the registered N-fixers impact on community composition (Vitousek et al., 1987). Recently, an integration of the current mechanisms determining the fate of invasions uses niche differences versus fitness differences as a framework to better understand the invasion success (MacDougall et al., 2009). Fitness differences describe species differences that drive competitive dominance. Thus, while niche differences favour species coexistence, fitness differences drive competitive exclusion (MacDougall et al., 2009). Integrating both aspects of species differences may help to explain small controversies from single theories. For instance, for a NIS to become invasive it must use resources also required by resident species, i.e. little niche difference, but in a superior effective way, i.e. large fitness difference. On the other hand, NIS that become established based on niche differences are unlikely to exert major impacts and coexist with resident species. Therefore, this integrative framework highlights the fact that some NIS do not require niche differences to invade (MacDougall et al., 2009). 1.3. Introduced macroalgae and known impacts The extent and current rate of species introductions in the marine environment are astonishing. Marine surveys have identified at least 298 introduced algal and invertebrate species in North America (Ruiz et al., 2000), 573 NIS in the Mediterranean Sea (Galil, 2009), 129 NIS in Australian waters (Hayes et al., 2005) and 987 marine NIS in Europe (www.europe-aliens.org). A recent review described 277 records of introduced seaweed, which make up a significant proportion of marine introduced species (Schaffelke et al., 2006; Williams and Smith, 2007). Overall, there are five wellknown algae (Fig. 1.2) which are reported to have invaded multiple regions around the world, i.e. Codium fragile (Suringar) Hariot spp. tomentosoides (Van Goor) Silva, Chapter 1 12 Caulerpa taxifolia (Vahl) C. Agardh, Grateloupia turuturu Yamada, Sargassum muticum (Yendo) Fensholt and Undaria pinnatifida (Harvey) Suringar (e.g. Araújo et al., 2011; Boudouresque and Verlaque, 2002; Critchley et al., 1983; Occhipinti-Ambrogi and Savini, 2003). The macroalge Caulerpa taxifolia and Undaria pinnatifida are ranked as two of the World’s 100 most invasive species (www.issg.org). Fig. 1. 2. Examples of high-impact invasive species recorded for the NE Atlantic. (A) green alga Codium fragile spp. tomentosoides, native to the Asian Pacific, (c) David Villegas; (B) Japanese wireweed Sargassum muticum, native to SE Asia, (c) David Villegas; (C) red alga Grateloupia turuturu, also native to the Asian Pacific Region (c) Fátima Pinto; (D) Wakame Undaria pinnatifida, native to Japan, (c) David Villegas. Marine macroalgae, i.e. seaweeds, dominate the rocky intertidal in most oceans, and in temperate and Polar regions cover rock surfaces in the shallow subtidal, describing General introduction 13 particularly important ecological functions. Seaweeds make a substantial contribution to marine primary production (Mann, 1973), and thus form much of the basis for intertidal food webs (Little and Kitching, 1996). Increasing evidence suggests that invasion of macroalgae may cause ecological perturbations in natural communities (Grosholz, 2002). Impacts of non-indigenous macroalgal populations 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). Macroalgae can alter light availability to other species, change nutrient cycling, affect herbivory intensity (Britton-Simmons, 2004; Sánchez et al., 2005; Yun and Molis, 2012), modify ecosystem properties and ultimately they may decrease native diversity (Casas et al., 2004). However, some studies describe no significant impact of non-indigenous marine algae in specific areas (e.g. Cecere et al., 2011; Forrest and Taylor, 2002), highlighting unpredictable nature of invasions. For example, a three-year study of sheltered low shore assemblages found little impact from the introduction of Undaria pinnatifida (Forrest and Taylor, 2002) while large ecological consequences were described after U. pinnatifida introduction in sites without large canopy species (Casas et al., 2004; Forrest and Taylor, 2002). Other seaweed that has spread dramatically during the last century is Codium fragile ssp. tomentosoides. It has been reported as a fast growing species, growing up to 170 thalli m-2, with high impacts on recipient assemblages by damaging and replacing native kelp (Laminaria spp.) forests with potential impacts on associated fauna (Levin et al., 2002; Trowbridge, 1995). Moreover, the accumulation of masses of C. fragile ssp. tomentosoides rotting on beaches of the NW Atlantic, Mediterranean, and New Zealand produces a foul odor that drives away visitors. Despite its known capacity for spreading and replacing indigenous species (Carlton and Scanlon, 1985; Nyberg and Wallentinus, 2005), there are some sites where no impact has been registered for this species, as for example in the Azores (Cardigos et al., 2006) and in the eastern North Atlantic Ocean (Chapman, 1998). So far, no link has been found between the ability of an introduced species to spread rapidly (its invasiveness) and the likelihood that it will have a strong impact on the recipient community (Ricciardi and Cohen, 2007). Thus, it is critical to understand the underlying ecological principles behind successful marine macroalgal invasions. The North East Atlantic, in particular, has been home to a large number of introductions (Arenas et al., 2006a; Eno, 1996; Rueness, 1989). Recent reviews on introduced marine macroalgae include assessment of introduction vectors as well as mechanisms that may influence invasion success (Hewitt et al., 2007; Schaffelke et al., 2006; Chapter 1 20 although in N Portugal intertidal it can only be observed in channels and tidepools (Incera et al., 2011; Monteiro et al., 2009). Longevity estimates of S. muticum in Portugal, using a matrix model are of 2.8–8.4 years (Engelen and Santos, 2009). Intrinsic traits such as being a fast growing species, high fecundity, monoecious, selffertile and pseudo-perennial life history (Fig. 1.3), among others, have been cited as responsible for the success of S. muticum as invader (Norton, 1976). The thallus of S. muticum is highly differentiated and has several distinct morphological features, with a marked seasonal variation. It consists of a perennial holdfast from which main axes and annual lateral branches grow. This species has 4 characteristics growth phases, suggesting a highly specialized adaptation to life in a seasonal environment (Fig. 1.3): 1) initial growth phase with large basal leaves which increase photosynthetic surface area; 2) elongation growth where the presence of gas-bladders maintain the thallus erect and closer to the light; 3) reproductive growth (April/May), where energy allocation is now shifted to the production of reproductive tissue (receptacles) and 4) a senescence period (from August onwards) where primary laterals degenerate and float away (Arenas et al., 1995; Deysher, 1984). Fig. 1. 3. Sargassum muticum life-cycle. General introduction 21 A fertile individual of S. muticum bears thousands of receptacles (reproductive structure) and bear both male and female conceptacles, thus self-fertilization is common (Deysher and Norton, 1982). Gamete expulsion and release has a semilunar periodicity, peaking around full and new moon (Monteiro et al., 2009). Fertilized eggs are retained in the external surface of the receptacles until the development of tiny germlings with adhesive rhizoids (Norton, 1981). Embryos are then liberated, they sink and generally reattach to the substratum within a few meters of the parent plant (Deysher and Norton, 1982). The characteristic egg expulsion by pulses, separated of several days (Norton, 1981) can also be seen as an advantageous characteristic of S. muticum when facing non-appropriate environmental conditions. However, multiplerange dispersal mechanisms have been cited for this species, including germling settlement and drifting fertile thalli (Norton, 1976). The characteristic air bladders provide a good mode of dispersal for dislodged individuals (Rueness, 1989). Replacement of native species, increase of filamentous epiphytic algae, changes in composition of flora and fauna, increased sedimentation, interference with coastal fisheries large accumulations of drift algae, blocking of narrow sounds and harbours, and interference with recreational activities are some of the known impacts of S. muticum (Critchley et al., 1986; Mack et al., 2000; Pedersen et al., 2005). For example, competition between the giant kelp Macrocystis pyrifera and S. muticum has shown that S. muticum inhibited the recruitment of M. pyrifera (Ambrose and Nelson, 1982). Moreover, indirect effects affecting abiotic conditions, such as shading or alteration of the current by the dense canopy during the summer months may affect settlement and recruitment of benthic organisms (Britton-Simmons, 2004; Critchley et al., 1990; Staehr et al., 2000). Because S. muticum also undergoes a faster and more complete decomposition than that of the native flora it has replaced, it has been suggested to alter the energy flow, by increasing the turnover rate and regeneration of nutrients (Pedersen et al., 2005; Rossi et al., 2011) and rate of primary production (Cebrián and Duarte, 1995). Replacement of native species, increase of filamentous epiphytic algae, changes in composition of flora and fauna, increased sedimentation, interference with coastal fisheries large accumulations of drift algae, blocking of narrow sounds and harbours, and interference with recreational activities are some of the known impacts of S. muticum (Critchley et al., 1986; Mack et al., 2000; Pedersen et al., 2005). For example, competition between the giant kelp Macrocystis pyrifera and S. muticum has shown that S. muticum inhibited the recruitment of M. pyrifera (Ambrose and Nelson, 1982). Chapter 1 22 Moreover, indirect effects affecting abiotic conditions, such as shading or alteration of the current by the dense canopy during the summer months may affect settlement and recruitment of benthic organisms (Britton-Simmons, 2004; Critchley et al., 1990; Staehr et al., 2000). Because S. muticum also undergoes a faster and more complete decomposition than that of the native flora it has replaced, it has been suggested to alter the energy flow, by increasing the turnover rate and regeneration of nutrients (Pedersen et al., 2005; Rossi et al., 2011) and rate of primary production (Cebrián and Duarte, 1995). Although non-indigenous species are often linked to negative impacts on the indigenous biota, some positive effects have also been described. In the North Sea, S. muticum provides habitats for epibiota otherwise absent in sediments, with consequent strong effects on diversity of soft sediments but not of hard substrata, where native species perform the same function (Buschbaum et al., 2006). Moreover, a recent study in the Galician coast revealed that S. muticum offered a suitable habitat for many invertebrates (Gestoso et al., 2012). 1.6.2. Life in intertidal rock pools In the NW coast of Portugal and Galicia, Sargassum muticum is very abundant in midand low-intertidal rock pools and research was mostly conducted in these habitats (Fig. 1.5). Tidepools are a special habitat on the rocky intertidal platforms, which provide refuge from desiccating conditions during low tide (Metaxas and Scheibling, 1993). It has been suggested that tidepools do not represent an intertidal habitat, due to not being emerged during low tide (Underwood, 1981), nor reflect life in the sublittoral zone, due to dramatic environmental fluctuations in these habitats (Little and Kitching, 1996). Environmental conditions in rock pools are, however, regulated by the tidal cycle (Metaxas and Scheibling, 1993). Each rock pool exhibits specific physic-chemical parameters such as temperature, salinity, oxygen, carbon dioxide and pH (Huggett and Griffiths, 1986), as those vary particularly in relation to height on the shore, but also with pool size, shape and assemblage composition (Metaxas and Scheibling, 1993). However, tidepools not only vary in environmental conditions but also in the availability of food (e.g. Underwood, 1984), level of predation (e.g. Connell, 1970) and availability of recruits (Coleman and Brawley, 2005). General introduction 23 Particularly for macroalgal assemblages, there is evidence that pool depth is of major importance (Martins et al., 2007), probably due to larger fluctuations of physic-chemical parameters in shallow pools (Metaxas and Scheibling, 1993). During daytime, algae absorb carbon dioxide and produce oxygen by photosynthesis whereas at night, when photosynthesis stops, respiration absorbs much of the available oxygen (Morris and Taylor, 1983). This physiological activity provokes changes in the pH of the water, with an inverse relationship with pCO2 (Morris and Taylor, 1983). Moreover, the physicchemical conditions within tidepools are related to the biomass of fauna and flora, and show a marked seasonal variation. Thus, the referred diurnal changes are the result of interactions between physic-chemical and biological processes (Morris and Taylor, 1983). Overall, like other intertidal habitat pools are very amenable for experimental studies. Also, because manipulations are easy to carry, tidepools can be used as experimental mesocosm providing an ideal system in which to study invasion processes. Fig. 1.5. Intertidal rocky shore in Ria de Vigo (Galicia, Spain). © Eva Cacabelos. Chapter 1 24 1.6.3. The use of synthetic macroalgal assemblages Some authors argue that certain experimental designs may cause artifacts due to the assumption of random loss of species from the total species pool (Loreau et al., 2001) in experiments such as mesocosms or with synthetic assembled assemblages. It has been demonstrated that non random changes in biodiversity of macroalgal assemblages can have different effects on ecosystem functioning to random changes (Bracken et al., 2008), however, only few studies have focused on the selective removal of key species in the field, to simulate non random changes in patterns of distribution and abundance of organisms (Díaz et al., 2003). Studies based on “natural” ecosystems are scarce (but see Arenas et al., 2009), which leads to a very pertinent point on how comparable are synthetic versus natural assemblages. It is important to recognise that all manipulative experiments are in fact abstractions from reality and all approaches have significant strengths and weaknesses. The great advantage of laboratory experimentation is the opportunity to control the system conditions and to take precise measurements of response variables. Its disadvantages are a measure of its artificiality and possible lack of applicability to the natural world as it is isolated from many of the biotic interactions (e.g. competition, predation, facilitation) that occur in a natural system. On the other hand, the lack of full controlled experimentation in field studies will tolerate possible effects over and above patterns of natural variation, which will represent an interaction closer to reality. It is, however, suggested that biodiversity effects are significantly weaker in less-controlled experimental systems (Hooper et al., 2005; Loreau et al., 2001). Ideally, we should aim to design experiments to test the same hypotheses using the same assemblages in both the laboratory and the field (e.g. Tait and Schiel, 2010). The intensity of interactions among sessile individuals varies with the spatial arrangement of competitors, and thus, the relative proximity of neighbouring individuals may be an important factor influencing the interactions among seaweeds (Carpenter, 1990; Kim, 2002). Moreover, macroalgal assemblages in tidepools are very patchy systems at very small scales, where extremely small spatial scales of variation seem to be among the most important sources of heterogeneity (Archambault and Bourget, 1996). Thus, synthetic macroalgal assemblages resembling the patchy “tile-like” pattern of intertidal algal assemblages are expected to describe quite realistically the natural system. General introduction 25 1.7. Objectives and structure of the thesis The general objective of this thesis was to provide critical knowledge of the processes governing biological invasions in intertidal rocky shores. Specifically, this PhD Thesis aimed to determine the relevant physical and biological processes affecting the invasion success and potential impacts of the brown macroalga Sargassum muticum. The specific objectives were: i) To assess the interactive effects of functional diversity and propagule limitation on different macroalgal invasion stages; ii) To study how bottom-up (resources) and top-down (grazers) processes operate interactively to determine invasion success; iii) To investigate ecosystem-level impacts of macroalgal invasions and discriminate the mechanisms involved behind the impact; iv) To test the effects of disturbance by climate change-related drivers on early survivorship of S. muticum germlings and legacy effects over recruitment success. The two first objectives, tackled in Chapter 2 and 3, are centered on the interaction between native ecosystem and the invasive success of the macroalga S. muticum, employing laboratory and empirical field experiments using intertidal macroalgal assemblages. The third objective is assessed in Chapter 4 and 5, which investigated the dynamics of production and respiration in native macroalgal assemblages and assemblages invaded by S. muticum, and the mechanisms involved in the competition between the native macroalga Cystoseira humilis and the invasive macroalga S. muticum. Lastly, Chapter 6 examined the effect of climate change drivers on the invasion success of S. muticum and on the susceptibility of marine macroalgal assemblages to invasions. Chapter 1 26 1.8. References Abbott IA, Hollenberg GJ (1976) Marine algae of California. Stanford University Press, California Ambrose RF, Nelson BV (1982) Inhibition of giant kelp recruitment by an introduced brown alga. Bot Mar 25:265-268 Araújo R, Bárbara I, Santos G, Rangel M, Sousa-Pinto I (2003) Fragmenta Chorologica Occidentalia. 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Marine Biology 159:1475-1484 Part II The process of macroalgal invasions 43 Chapter 2 Propagule pressure and functional diversity: interactive effects on a macroalgal invasion process Chapter 2 45 Propagule pressure and functional diversity: interactive effects on a macroalgal invasion process Fátima Vaz-Pinto1,2,*, Celia Olabarria3, Francisco Arenas1 1 Laboratory of Coastal Biodiversity, CIIMAR – Centro Interdisciplinar de Investigação Marinha e Ambiental, Universidade do Porto, 4050-123 Porto, Portugal 2 ICBAS – Instituto de Ciências Biomédicas Abel Salazar, Universidade do Porto, 4050313 Porto, Portugal 3 Departamento de Ecología y Biología Animal, Facultad de Ciencias del Mar, Universidad de Vigo, 36310 Vigo, Spain Vaz-Pinto F., Olabarria C., Arenas F. Propagule pressure and functional diversity: interactive effects on a macroalgal invasion process. Marine Ecology Progress Series, in press Chapter 2 52 remained for 12 months. The selected rock pool was S. muticum free, as all the individuals were ripped off by its holdfast days before the start of the experiment. During the course of the experiment the rock pool was continuously monitored and new individuals of S. muticum removed with a scraper. 2.3.5. Measurements of invasibility and sampling procedure Invasion of Sargassum muticum was monitored over three stages of invasion; settlement, recruitment and colonisation. Settlement of S. muticum germlings was assessed 2 to 4 days after lab exposure to propagule rain, prior to transportation of assemblage plates to the field. Recruitment discs from each assemblage plate were used to record germlings densities with the use of a stereo microscope with additional dimmed light. Plates were then transported to the field. In order to assess the effect of transport, 26 additional recruitment discs (13 per PP treatment) with known densities of S. muticum germlings were taken for field transportation and returned to the laboratory where germlings were recounted. Mortality rate (± SE) due to transportation (4.3% ± 0.91) was insignificant when compared to 2-months field natural mortality rate of ≈ 99% (± 0.34) (F1,50 = 2673.16, P < 0.0001). Thus, mortality due to transportation was not taken into account when interpreting results. Two months after lab exposure to propagule rain (September 2009) recruits were visible to the naked eye and recruitment was assessed by counting Sargassum muticum recruits in the assemblage plates at the substratum level. Colonisation success was assessed by quantifying adult S. muticum (> 20 cm length) survivorship in assemblage plates from September 2009 to April 2010. In addition, we also estimated percentage cover of algae (primary and secondary cover) and bare rock surface. Algal cover was estimated by visually dividing each rock cube in 4 and attributing a score from 0 to 4 to the functional groups present and adding up the 16 estimates (Dethier et al., 1993). The above canopy area was measured as secondary cover when subcanopy species were present. Available resources (space and photosynthetic active radiation, PAR) were also monitored for each macroalgal assemblage prior to the invasion procedure. Space availability was assessed during algae cover monitoring by quantifying percentage of bare rock at the substratum level. PAR penetration (measured as the ratio of PAR under the algal canopy to that above it, or in the case of encrusting species functional group at the substratum level to that 10 cm above it) was registered with a spherical Chapter 2 53 quantum scalar irradiance sensor (1.9 cm diameter sensor, Mod. QSL–2100 Biospherical Instruments - Inc., USA). PAR availability at substratum level was related to the irradiance intercepted per functional group, and represented a measure of how much PAR reached the substratum. Experimental plates were submerged in a 5-L tank during PAR availability measurements. For each plate, the sensor was moved throughout the substrata (below assemblage canopy) and 10 cm above (or over assemblage canopy) for 30 seconds each. 2.3.6. Statistical analyses Differences in resources availability among functional diversity treatments was assessed with a one-way ANOVA. Functional diversity was a fixed factor with 4 levels (i.e. encrusting, turf, subcanopy and high functional diversity). Both, space and PAR data had homogeneous variances (Cochran’s test, C = 0.3518 and C = 0.3511, respectively; P > 0.05). No Sargassum muticum recruitment was recorded in plates from control PP treatment over the 12 months experimental period. All new S. muticum recruits counted were assumed to come from the laboratory artificial invasion and thus analyses were only performed for low and high PP treatments. Generalized linear models (GLM) were fitted for settlement and recruitment data to investigate the influence of the explanatory variables over the variation in the number of S. muticum germlings and recruits, respectively. For both sets of count data, we assumed a negative binomial distribution (an extension of GLMs) to account for high overdispersion using the glm.nb function from the MASS package (Venables and Ripley, 2002). Colonisation success was assessed with presence-absence data where 1 represented survivorship of S. muticum and 0 represented no S. muticum survivorship to adult stage. For presence-absence data we used a binomial GLM and a logit link function. Initial fitted models were built using PP (2 levels, high vs low), FD (5 levels this time: bare rock, monospecific assemblages of encrusting species, monospecific assemblages of turf-forming species, monospecific assemblages of subcanopy species and high functional diversity assemblages) and two continuous variables, PAR and space, as predictors. The continuous variable space availability was not included in the predictive model performed for settlement due to the use of recruitment discs in this stage. We also included biologically reasonable two-way interactions in the initial model. To obtain the minimum adequate model we performed stepwise variable selection minimizing the Akaike’s information criterion (AIC) and then hypothesis testing (Zuur et Chapter 2 54 al., 2009). A posteriori contrasts were performed to assess statistical significance between the various levels of the categorical factor Functional diversity (Crawley, 2009). To assess collinearity between variables we calculated the variance inflation factor (VIF) with vif function in the car package (Fox and Weisberg, 2011). All VIF values were below 8, indicating low collinearity, so we kept all variables in the analyses. Residuals from the selected model were plotted against the fitted values for model validation. Rates of instantaneous mortality of the invader were used to test whether or not the mortality rates were constant over the different invasion stages and were calculated following Viejo et al. (1999) as: Ln [(Nt/Nto) + 0.0001] / t where Nt/Nto is the proportion surviving per macroalgal assemblage and t is the time after invasion (0-2 months, settlement period; 2-6 months, recruitment period; 612, colonisation period). At the end of the experiment 2 out of the 60 experimental plates were lost due to rough weather and, thus colonisation data were analysed with three replicates instead of the initial four in two treatments. The design was balanced by adding the averaged values of the other three replicates to the model and recalculating the F-values (see Underwood, 1997). Rates of instantaneous mortality were analysed with a three-way ANOVA with PP (2 levels), time (3 levels) and FD (5 levels) as orthogonal fixed factors. Variance was homogeneous (Cochran’s Test, C = 0.1109; P > 0.05). GLMs were carried out using the R-program 2.12.1 (R Development Core Team, 2010). Analysis of variance for resources availability and instantaneous mortality was performed with WinGMAV 5 (http://sydney.edu.au/science/bio/eicc). 2.4. Results 2.4.1. Resources availability PAR availability at substratum level varied significantly among functional diversity treatments (ANOVA, F3,28 = 35.20, P < 0.001). Monospecific assemblages of encrusting species showed the greatest values of PAR reaching the substratum level (Fig. 2a). PAR availability was particularly low at assemblages composed of subcanopy species, i.e. monospecific assemblages of subcanopy species and assemblages of high Chapter 2 55 functional diversity, whereas assemblages of turf-forming species showed intermediate levels (Fig. 2.2a). Space availability varied significantly among functional diversity treatments (ANOVA, F3,28 = 29.52, P < 0.0001). Monospecific assemblages of subcanopy species and turfforming species presented around 40 and 30%, respectively, of bare rock surface while monospecific assemblages of encrusting species exhibit lowest values (≈ 9%). Plates assembled with the three functional groups presented intermediate levels of availability of primary space (Fig. 2.2b). Fig. 2. 2. Availability of resources (mean +SE, n = 8) at the substratum level over the different functional diversity treatments, including those made only of bare rock. FD indicates functional diversity. A) Photosynthetic active radiation (PAR) availability; B) space availability as percentage cover of bare rock. Different letters indicate significant differences at P < 0.05. Chapter 2 56 2.4.2. Invasion success at different invasion stages Invasion success of Sargassum muticum was partially explained by some of the predictive variables included in the models, namely propagule pressure, functional diversity and PAR availability. However, the variables included in the minimum adequate models differed between the three stages of invasion, i.e. settlement, recruitment and colonisation. A different order of inclusion of the variables did not affect the results. Quantification of Sargassum muticum individuals at the different experimental treatments for the different stages of invasion are presented in Fig. 2.3. Settlement of S. muticum germlings was interactively influenced by propagule pressure and functional group diversity (Table 2.1a), with much higher settlement at high PP treatment (Fig. 2.3a). Overall, the explanatory variables from the model explained ≈ 92 % of the variability in the number of S. muticum settled germlings (Table 2.1a). The interactive effects of PP and FD resulted in contrasting significant numbers of settled germlings between macroalgal assemblages at low and high PP. In the low PP treatment, monospecific assemblages of turf-forming species and high functional diversity assemblages registered significantly more settled germlings compared to the other assemblages (Fig. 2.3a). However, in the high PP treatment an opposite pattern was observed (Fig. 2.3a). Unlike observed at settlement stage, propagule pressure was not a significant predictor of Sargassum muticum recruitment success. Thus, the fitted model for recruitment data, i.e. S. muticum juveniles recorded on assemblages at substratum level, included only functional diversity as a significant predictor and explained 52.4% of the variability in the number of S. muticum recruits (Table 2.1b). In particular, bare rock assemblages had significantly higher number of recruits than the other assemblage types (Fig. 2.3b). At the colonisation stage, survivorship of Sargassum muticum was mainly influenced by PAR availability at the substratum level (Table 2.1c). Survivorship of S. muticum to adult stage was only observed at high PAR availability (Fig. 2.4). 2.4.3. Instantaneous mortality rates The instantaneous mortality rate varied over time, independently from functional diversity of assemblages or propagule pressure (ANOVA, P < 0.0001; Table 2.2). Posthoc multiple comparisons showed that instantaneous mortality rates were significantly Chapter 2 57 greater during the first two months (mortality of settlers) than over the recruitment period (2 to 12 months after invasion). In addition, instantaneous mortality rates were also greater in the first 4 months (2 to 6 months after invasion) compared to the last 6 months of the experiment (Fig. 2.5). Fig. 2. 3. Sargassum muticum over a) settlement (per 2.26 cm2), b) recruitment (per 64 cm2) and c) colonisation (per 64 cm2). Mean (+SE, n = 4 per functional group and bare rock) number of individuals on assemblages of different functional diversity, including those made only of bare rock. FD indicates functional diversity, PP indicates propagule pressure. Chapter 2 58 Table 2. 1. Sargassum muticum. Analysis of deviance revealed by generalized linear models fitted to the explanatory variables Propagule pressure (PP), Functional diversity (FD), Photosynthetic active radiation (PAR) and Space availability. Variable’s parameters described for the minimum adequate model for the response variable a) number of S. muticum germlings, b) number of S. muticum recruits and c) presence-absence data for survivorship of S. muticum to adult stage. Variables df Deviance Residual df Residual Deviance P AIC a) Settlement Model 39 498.68 PP 1 434.47 38 64.21 <0.0001 FD 4 9.12 34 55.09 0.058 PP:FD 4 14.28 30 40.81 0.006 Minimum adequate model 530.77 Full model: PP*FD + PAR+ PP:PAR + FD:PAR, family = negative.binomial 531.43 b) Recruitment Model 39 88.372 FD 4 46.302 35 42.070 <0.0001 Minimum adequate model 209.7 Full model: PP*FD + PAR + Space + PP:PAR + PP:Space + FD:PAR + FD:Space, family = negative.binomial 228.57 c) Colonisation Model 39 42.653 PAR 1 15.425 38 27.228 <0.0001 Minimum adequate model 31.29 Full model: PP+FD+PAR+Space+PP:FD+FD:PAR, family = binomial 53.44 Fig. 2. 4. Sargassum muticum colonisation success. Variation in the survivorship of S. muticum to adult stage (from 2 to 9 months) with photosynthetic active radiation (PAR) availability at substratum level on macroalgal assemblages. Chapter 2 59 Table 2. 2. Sargassum muticum. Results of the three-way ANOVA testing for the effects of Propagule pressure (PP), Time (T) and Functional diversity (FD) on rates of instantaneous mortality. * Two missing replicates were replaced by the average value of the three replicates remaining in the treatment, and 2 df subtracted from Residuals. Source df MS F P PP 1 0.082 0.11 0.741 T 2 136.747 177.22 <0.0001 FD 4 0.112 0.14 0.967 PP x T 2 0.209 0.27 0.764 PP x FD 4 0.076 0.10 0.982 T x FD 8 0.712 0.92 0.504 PP x T x FD 8 1.433 1.86 0.076 Residuals 88* 0.772 Fig. 2. 5. Sargassum muticum. Instantaneous mortality rates (mean +SE, n = 40) during settlement (0 to 2 months), recruitment (2 to 6 months) and colonisation period (6-12 months after invasion). Different letters indicate significant differences in at P < 0.05. 2.5. Discussion We found that the main and interactive effects of propagule pressure and the functional structure of assemblages varied over the different stages of invasion, i.e. settlement, recruitment and colonisation. Moreover, PAR availability at the substratum level was an important variable explaining the invasion success of Sargassum muticum. In contrast to previous studies on S. muticum (e.g. Stæhr et al., 2000; Britton-Simmons, 2006), space availability did not play a significant role explaining recruitment success of the invader. Chapter 2 60 The role of propagule pressure in the invasion process was only relevant over the settlement stage (Clark and Johnston, 2009). Although high percentage cover of canopy, i.e. monospecific assemblages of subcanopy species, may act as increased biotic resistance under low propagule pressure, high propagule pressure seemed to overwhelm biotic resistance to invasion. The swaying of the algal fronds induced by water movement could free the substratum they covered, allowing for propagules to reach the substratum. Here, because of the use of discs for settlement assessment, it was impossible to separate and model the effect of space availability on the number of settled germlings. Over the settlement stage, the observed larger number of Sargassum muticum germlings in algal turfs than in areas covered by encrusting species has been also registered for Cystoseira spp. (Fucales, Phaeophyta) (Benedetti-Cecchi and Cinelli, 1992). Explanations such as that algal turf might provide a refuge from water displacement (Brawley and Johnson, 1991) or from herbivores (Benedetti-Cecchi and Cinelli, 1992) are not applicable within laboratory artificial invasion. During the process of settlement, propagules have first to locate the substratum and then to establish surface contact (Fletcher and Callow, 1992). Hence, a possible explanation is that the highly compacted algal turfs entrapped the germlings (Fletcher and Callow, 1992) allowing a higher probability of surface contact. Greater biomass of subcanopy species, on the other hand, could either prevent S. muticum germlings from reaching the substratum (Deysher and Norton, 1982) or remove germlings due to scour (Vadas et al., 1992). Nevertheless, results suggested that the positive effect of turf in high diverse assemblages could exceed the negative effect of subcanopy fronds. Despite the possibility of S. muticum settling over the fronds (Deysher and Norton, 1982), that information was not quantified in the present study. However, no recruits were observed over fronds latter on. The invader, Sargassum muticum, presents a highly successful reproductive strategy by releasing large amount of germlings intermittently at intervals of ≈15 days throughout its fertile period (Fletcher, 1980; Norton, 1981). This strategy will significantly increase the chances of at least one release coinciding with favorable conditions for settlement as suggested by studies manipulating disturbance (BrittonSimmons and Abbott, 2008; Clark and Johnston, 2009). Several studies have reported space and light availability as important resources limiting Sargassum muticum recruitment (Andrew and Viejo, 1998; Britton-Simmons, 2006). Results from the present study, however, suggested that the number of S. muticum at recruitment stage was not significantly affected by either resource. Nonetheless, the fact that bare rock assemblages showed the greatest number of S. muticum recruits suggest that space Chapter 2 61 availability is determinant for germlings settlement. Encrusting species have been described as suitable settlement substrate in some studies (Ang, 1985; Arenas et al., 2006), while in others they have been suggested to act as an effective barrier to colonisation (Deysher and Norton, 1982). In the present study, monospecific assemblages of encrusting species had the second greatest number of S. muticum recruits, highlighting the aptness of encrusting species as a suitable substratum for S. muticum. In fact, these results corroborate our field observations that indicate the presence of S. muticum adults over crustose substratum. Because the structure of assemblages explained ≈52 % of S. muticum recruitment, other factors related to postsettlement survivorship such as nutrient availability, herbivory or wave action might be related with the results found. Colonisation success of Sargassum muticum was positively affected by percentage of PAR reaching the substratum, suggesting that the presence of high values of canopy biomass may have a negative effect on the invader survivorship. Canopy may indeed affect significantly the survival of recruits older than 6 months of age (Kendrick, 1994), possibly explaining our non-significant results for 2 months recruits regarding PAR availability. Although highly diverse communities are expected to achieve a more complete utilization of resources (complementary resource use hypothesis), our experiment based on functional attributes of species provided little support for that. Nonetheless, colonisation of Sargassum muticum was more successful in less complex assemblages, i.e. monospecific assemblages of encrusting species, which presented higher levels of both resources, PAR and space availability. Additionally, rate of instantaneous mortality of S. muticum was significantly greater over the first two months of age indicating that early post-settlement processes are decisive to the overall invasion success. This fact suggests that post-settlement mortality not related with our predictors (e.g. density dependent mortality) eliminates the effect of propagule pressure in the final success of the invasion. Overall, our results did not support the theory that more diverse assemblages are less susceptible to invasion by non-native species (Elton, 1958; Lodge, 1993; Naeem et al., 2000). The traits inherent of each functional group have, however, strong effects on specific resources (Symstad, 2000; Britton-Simmons, 2006). This fact and our results suggested that the additive effects of each functional role in high diverse assemblages are important in modifying resources availability and may therefore interfere with community invasibility (Arenas et al., 2006). Our findings showed that the invasion Chapter 3 69 Role of top-down and bottom-up forces on the invasibility of intertidal macroalgal assemblages Fátima Vaz-Pinto1,2, Celia Olabarria3, Francisco Arenas1 1 Laboratory of Coastal Biodiversity, CIIMAR – Centro Interdisciplinar de Investigação Marinha e Ambiental, Rua dos Bragas 289, 4050-123 Porto, Portugal 2 ICBAS – Instituto de Ciências Biomédicas Abel Salazar, Rua de Jorge Viterbo Ferreira, Universidade do Porto, 4050-313 Porto, Portugal 3 Departamento de Ecología y Biología Animal, Facultad de Ciencias del Mar, Universidad de Vigo, 36310 Vigo, Spain Vaz-Pinto F., Olabarria C., Arenas F. Role of top-down and bottom-up forces on the invasibility of intertidal macroalgal assemblages. Journal of Sea Research, in press Chapter 3 70 3.1. Abstract Despite the available information regarding the negative effects of non-indigenous species (NIS) on ecosystem structure and functioning, the mechanisms controlling NIS invasion remain poorly understood. Here, we investigated the relative roles of top-down and bottom-up control on the invasion of intertidal macroalgal assemblages by the macroalga Sargassum muticum (Yendo) Fensholt. Using a factorial experiment, nutrient availability and intensity of herbivory were manipulated along an intertidal rocky shore. We found that early recruitment of S. muticum was enhanced by low nutrient enrichment but no effect of grazers was observed. In contrast, at the end of the experiment (9-months after invasion) top-down control, together with the number of NIS and the percentage cover of ephemerals, were significant predictors for the invasion success of S. muticum. In addition, both top-down and bottom-up forces played a significant role in structuring macroalgal assemblages, which indirectly could have influenced invasion success. Hence, by shaping community structure, main and interactive effects of bottom-up and top down forces may indirectly act on invasion. Our study highlights the importance of the recipient community structure on the invasion process and emphasizes the specific regulation of top-down and bottom-up forces in different stages of S. muticum invasion. Keywords: Invasion process; Top-down; Bottom-up; Sargassum muticum; NIS Chapter 3 71 3.2. Introduction nvasions by non-indigenous species (NIS), one of the main human-mediated impacts on ecosystems, can directly and indirectly induce major effects on the structure and functioning of ecosystems (Grosholz, 2002). Understanding which factors influence the susceptibility of communities to invasion has become a central concern in ecology. Although several factors are thought to play a role in the susceptibility of an ecosystem to invasion (Stachowicz and Byrnes, 2006), the composition and diversity of native species have been considered some of the primary causes (e.g. Arenas et al., 2006; BrittonSimmons, 2006; Elton, 1958). Commonly, reduced resource availability is the most cited mechanism driving community invasibility (e.g. Stachowicz et al., 1999; Britton-Simmons, 2006). Increased diversity of species and functional groups is expected to increase the efficiency of resource use, generating a negative relationship between native diversity and invasion success (Stachowicz and Byrnes, 2006). Despite some controversy, the relationship between native diversity and invasibility is well reported and widely accepted (Davis et al., 2000; Stachowicz et al., 2007). In this context, the Fluctuating Resources Availability Theory (FRAT) suggests that the invasibility of a community may change as the amount of unused resources fluctuates (Davis et al., 2000). Accordingly, the susceptibility of a community to invasion would increase whenever the availability of a limiting resource is increased. In coastal systems, resource supplies and consumers play a crucial role in the regulation of intertidal habitats, i.e. bottom-up versus top-down control (Menge, 2000; Worm et al., 2000). Specifically, it has been suggested that bottom-up forces, e.g. nutrient levels, are of great importance in regulating species composition while top-down regulation, e.g. herbivory, exert stronger effects over macroalgal biomass (Burkepile and Hay, 2006). Increased availability of resources in an environment can be induced either by introducing resources at a faster rate than the community can sequester them (e.g. nutrient addition) or by reducing resource uptake by the community (e.g. decrease of species biomass) (e.g., Grime, 1977; Davis et al., 2000). Furthermore, complex interactions among herbivory and nutrients levels are also frequent (Worm et al., 1999; Worm et al., 2002; Hillebrand, 2003; Masterson et al., 2008). Although a large number of studies have addressed the influence of top-down and bottom-up forces in structuring benthic communities (e.g. Worm et al., 2002; Burkepile and Hay, 2006; Bulleri et al., 2012), few experiments have been conducted to address the relative roles of nutrient supply and herbivore processes on invasion success of I Chapter 3 72 macroalgae (but see Vermeij et al., 2009). Roughly 20% of the marine invasive species in the world are macroalgae. These invaders can cause important ecological and economic damage by altering the structure and functioning of recipient ecosystems (Schaffelke et al., 2006). Understanding the mechanisms and processes that allow marine invasive macroalgae to become abundant is crucial for prevention and management of marine invasions in the future. In the marine environment, macroalgal assemblages depend on resources such as space, nutrients and light (Carpenter, 1990). Thus, the susceptibility to invasion by a macroalgal community is expected to be influenced by the availability of those limiting resources (Arenas et al., 2006). For example, space availability and nutrient enrichment facilitated the establishment and spread of the brown invasive alga Sargassum muticum on an intertidal shore (Sánchez and Fernández, 2006). In addition, it has been suggested that susceptibility to invasion should be highest when disturbance is accompanied by nutrient enrichment (Burke and Grime, 1996). The presence of grazers in a habitat can, however, induce disturbance of different kinds either by increasing space availability or by reducing algal biomass. Hence, disturbance by grazers may reduce resource use by decreasing the biomass of algal assemblages (Davis et al., 2000). This phenomenon would increase the amount of resources available and create a favourable time period for new recruitment. Nonetheless, experimental studies addressing the relative roles of herbivory and nutrients in the recruitment of algae indicate complex and possible interactive effects of both factors. For instance, a study in the Great Barrier Reef that simultaneously manipulated intensity of nutrients and herbivory, found that herbivory effects on algal recruitment (density and size) were not consistent among species and were stronger within nutrient-enriched treatments (Diaz-Pulido and McCook, 2003). The invasion process is characterised by different transitional stages of invasion (Williamson, 2006) and survivorship through each stage is the key to success of invasive NIS. Here, we used a field factorial experiment using intertidal assemblages as a model system to test the individual, combined and potential interactive effects of nutrient enrichment and intensity of herbivory on the invasibility of intertidal macroalgal assemblages by the invasive macroalga Sargassum muticum (Yendo) Fensholt. Previous studies on S. muticum, both from the intertidal and low subtidal, suggested that invasibility of assemblages was mediated by different resources at different stages of invasion (Britton-Simmons, 2006; Sánchez and Fernández, 2006). Thus, experiments conducted from early life stages throughout adult stages are of great importance in macroalgal invasion ecology (White and Shurin, 2007). Chapter 3 73 Sargassum muticum Yendo (Fensholt) is a brown alga native to East Asia, and it is considered a highly invasive species (Critchley et al., 1983; Norton, 1976). Attached individuals were first recorded in Europe in the British Isles in 1973 (Critchley et al., 1983). The invasive seaweed S. muticum arrived to the rocky shores of the northern Spanish coast in the 1980s (Fernández et al., 1990) and is now present from Portugal in the south to Norway in the North. Characteristics such as being a fast growing species, monoecious, self-fertile and pseudo-perennial, among others, have been cited as responsible for the success of this species as an invader (Norton, 1976). A fertile individual of S. muticum bears thousands of receptacles (reproductive structure). Fertilized eggs are retained on the external surface of the receptacles until the development of tiny germlings with adhesive rhizoids (Norton, 1981). Embryos are then liberated, they sink and generally reattach to the substratum within a few meters of the parent plant (Deysher and Norton, 1982). Adult plants stay attached to the substratum by a perennial holdfast. Seasonal patterns of growth and reproduction have been reported for this species (Deysher, 1984; Arenas and Fernández, 1998). In the study area, S. muticum has a spring-summer reproductive period while a senescence period is observed at the end of the summer (Arenas and Fernández, 1998; Fernández, 1999). Biomass shows a marked seasonality with a period of rapid growth starting around January and a period of decrease in biomass at the end of the reproductive period when fronds detach themselves from the thallus. By experimental manipulation of nutrients and herbivory levels and by assessing success of the invader after an artificial invasion, we tested the hypothesis that a greater availability of resources, through an increase in herbivory or due to nutrient enrichment, would affect the invasion of macroalgal assemblages. Additionally, we expected an interactive effect between nutrient concentrations and intensity of herbivory, where the magnitude of top-down forces of herbivory should increase with the bottom-up forces of nutrient availability. 3.3. Material and methods 3.3.1. Study site The field study was conducted from February 2010 to March 2011 along an intertidal rocky shore in Cabo Estai (42º11’ N, 8º48 W), Ria de Vigo (Galicia, Spain). Seawater temperature for the experimental period was about 15ºC (±0.84, mean ±SE), while salinity remained stable at around 34.5‰ (www.meteovigo.es). The site can be described as a gently sloping granitic rocky shore, with a wide variability of fauna and flora. Natural assemblages at low intertidal were characterised by a range of Chapter 3 74 macroalgae with differing morphologies such as the leathery alga Bifurcaria bifurcata Ross, articulated calcareous algae such as Corallina sp. Linnaeus and Jania rubens (Linnaeus) Lamouroux, the corticated alga Stypocaulon scoparium Linnaeus, and green filamentous and foliose algae (e.g. Ulva spp). NIS such as Sargassum muticum (Yendo) Fensholt, Asparagopsis armata Harvey, Grateloupia turuturu Yamada and Undaria pinnatifida (Harvey) Suringar were also present at the site. Invertebrate assemblages were dominated by gastropod grazers, including several topshells from the genera Gibbula, the limpet Patella spp. and the common topshell Monodonta lineata Da Costa. 3.3.2. Experimental design and Sampling procedure Forty-five plots (35 x 35 cm) of natural assemblages were manipulated in the low intertidal (0.4-0.8 m above the lowest astronomical tide) following a two-way factorial design including the intensity of Nutrients and Herbivory as orthogonal fixed factors. All replicate plots were distributed randomly within the study site along 200 m of sea shore, separated by at least 2 m. The corners of the experimental plots were marked with epoxy putty to allow repeated sampling. To study the effect of increased nutrient concentration, three levels were included in factor Nutrients: “Control”, corresponding to no nutrient addition and mesh bags control; “Low nutrients”, corresponding to an addition of 1 Kg of nutrient pellets, N+; and “High nutrients”, corresponding to an addition of 2 Kg of nutrient pellets, N++. Concentrations of nutrients were manipulated using slow controlled-release fertilizer pellets (Multicote®) in small mesh bags with 1 mm mesh size. Fertilizer pellets consisted of 15% N (8% NH4+, 7% NH3), 7% P (PO2), 15% K (K2O) and 2% MgO. Each experimental plot had 2 mesh bags (500 g each or 1 kg each, for low and high nutrients, respectively) on opposite sides anchored to the substrate with cable ties fasten to two steel screw eyes. Control mesh bags were filled with a plastic bag containing sand. Mesh bags were replaced every 3 months, or before if lost due to rough sea conditions, to ensure continuous delivery of nutrients. Herbivory intensity also had three levels: Natural herbivory, “H+”, where herbivores had free access to experimental plots; Low herbivory, “H-“, where antifouling paint was used to avoid herbivore access to experimental plots; and Procedural control, “PC”, to test for possible artifacts due to the use of antifouling paint, where antifouling paint was only used on two opposite sides of the plots. There were 5 replicates per treatment combination. Experimental manipulation of nutrients and herbivory was maintained for three months prior to invasion. Prior to the invasion procedure, individuals of Sargassum muticum were Chapter 3 75 removed from the experimental plots and immediate surroundings. To assure a similar propagule pressure by S. muticum in all experimental plots, we invaded the plots artificially. Artificial invasion was set on 13th of May, 1 day before the new moon, due to the fact that S. muticum has a semilunar periodicity of egg expulsion coinciding to 24-48h around new or full moons (Monteiro et al., 2009; Norton, 1981). Fertile individuals of S. muticum (≈ 30 cm long) were collected from an intertidal rocky-shore nearby. Approximately 500 g wet weight of algae were fixed to opposite sides in each plot using cable ties and 1 steel screw eye and were left for 1 week. Algae percentage cover and density of herbivores were recorded immediately before treatment manipulation (at the beginning of the experiment, February 2010) and every 2-3 months thereafter until March 2011. Plots were monitored using a 35 x 35 cm quadrat divided into 100 sub-quadrats of 3.5 x 3.5 cm each. Primary and secondary algal cover was estimated by summing up 1% cover of each taxon individually and adding up the total. By joining primary and secondary algae cover, total cover may greatly exceed 100% within each plot. In addition, percentage cover of sessile invertebrates and bare rock were also assessed adding up to the assemblage structure of the experimental plot. Macroalgae were identified to the most detailed level of taxonomic resolution achievable in the field. Grazers were individually counted and were left in or taken out of the plot depending on the herbivory treatment. Macroalgal species were assigned to different functional groups following Steneck and Dethier (1994). Nutrient enrichment is known to often increase the biomass of opportunistic fast-growing ephemeral species (Kraufvelin et al., 2010). Thus, in order to examine the potential indirect effect of the presence of these particular species, we chose to incorporate together species known as being highly opportunistic into a single category, Ephemerals. We incorporated together annual filamentous and foliose algae belonging to the genera Ulva, Ceramium, Pterosiphonia and Polysiphonia (Supplementary data, Table A3.1). Additionally, known NIS were identified and included as a single variable named “NIS” (excluding S. muticum). Overall, the macroalgal categories used were: corticated, foliose, leathery macrophytes, crustose, articulated calcareous, ephemerals and NIS. S. muticum individuals were first visible in experimental plots in December 2010. At the end of the experiment (March 2011), recruits of S. muticum were scraped from each experimental plot, taken to the laboratory and oven-dried at 60°C for 48 hours to estimate biomass (grams of dry weight). Several variables such as number of recruits, length and biomass were used as estimates of invasion success (Arenas et al., 2006; BrittonSimmons and Abbot, 2008). Chapter 3 76 3.3.3. Statistical analysis Due to the specific amount of nutrient release in each plot, we assumed dependent repeated measures sampling. In this context, effectiveness of nutrients enrichment was examined using a Repeated measures ANOVA (rmANOVA) with Nutrients as a fixed factor and Month as within-subjects variable. The rm-ANOVA assumption of sphericity was evaluated using Mauchly’s criterion. Bonferroni post-hoc comparisons were performed between pairs of variables. Furthermore, effectiveness of herbivory treatment was analysed with a two-way ANOVA with Herbivory intensity (3 levels) as a fixed factor and Month (3 levels) as a random factor (n = 15). Firstly, we tested for differences in assemblage structure between patches assigned to each experimental plot, to assure that any differences found later on were due to experimental treatments. Structure of assemblages (measured as percentage cover of algae, bare rock and sessile invertebrates) was compared through multivariate analysis of variance based on permutations (PERMANOVA). Based on a root-transformed BrayCurtis similarity matrix, PERMANOVA analysis included 2 fully crossed fixed factors, Nutrients (3 levels) and Herbivory intensity (3 levels) with 5 replicates per factorial combination of treatments. Additionally, PERMANOVA analyses were also performed at the end of the experimental period to compare structure of assemblages at the beginning and at the end of the experiment. For the three-way permanova analysis, we included Nutrients (3 levels), Herbivory intensity (3 levels) and Time (2 levels) as fixed factors. SIMPER analysis was used to identify taxa contributing most to multivariate differences between experimental assemblages. To identify the main variables that influenced early recruitment and invasion success of S. muticum we used generalized linear models (GLMs). For early recruitment, count data expressed as number of recruits was analysed assuming a quasipoisson distribution to account for overdispersion. For final recruitment data we assumed a negative binomial distribution to account for high overdispersion using the glm.nb function from the MASS library (Crawley, 2009) in R. We assumed a Gaussian distribution for biomass and length data and used glm() function with an identity-link. Predictive variables for both set of analyses were the factors Nutrients and Herbivory intensity, and the variables number of native taxa, number of non-indigenous taxa, percentage cover of morpho-functional groups (crustose, leathery, corticated, articulated calcareous) and percentage cover of ephemeral species. Chapter 3 77 Generalized Estimating Equations (GEEs), an extension of GLMs for repeated measures data analysis (Liang and Zeger, 1986), were used to assess how the structure of assemblages within each experimental plot varied over the experimental period due to nutrient enrichment and herbivory intensity. The specific response of assemblages was analysed through changes in percentage cover of functional groups, ephemerals and species richness. We used the geepack package (Halekoh et al., 2006) and the geeglm() function in the R-program. For the analysis, a link function and an error structure for the residuals were specified as in GLM models described above. Hence, for count data we assumed a Poisson distribution and for continuous percentage cover data we assumed a Gaussian distribution. We also specified a first order autoregressive model, using the corAR1 function, in the whole analysis assuming time dependence for each experimental plot unit. All univariate analyses were carried out using the R-program 2.14.1 (R Development Core Team, 2011). Multivariate analyses were performed with PERMANOVA + for PRIMER v.6 (Anderson et al., 2008; Clarke and Gorley, 2006). 3.3.4. Effectiveness of treatments Effectiveness of the nutrient enrichment was assessed by measuring nutrient concentrations in experimental plots in May, August and December 2010. Water samples were collected following low tide, with water 10-15 cm above experimental plots, using acid cleaned 50-ml plastic syringes with microfiber filters (Fisherbrand® MF 300). Immediately after collection, samples were placed on ice, returned to the laboratory, and frozen at -20 ºC until analysis (less than three months after). Analyses of NO3and PO43+ were carried out using a continuous-segmented flow autoanalyzer (Bran+Luebbe AA3). Analysis revealed marginally significant differences in nitrate availability over the different treatments (rm-ANOVA, F2,42 = 3.055, P = 0.058). Within treatment levels, the Control differed from the High nutrient treatments while the Low nutrient treatment showed intermediate values, not significantly different from the Control or High nutrient treatments. Phosphate concentration was not significantly affected by nutrient enrichment (rmANOVA, F2,42 = 1.443, P = 0.248). To evaluate the effectiveness of the herbivory treatments, the intensity of herbivory in each experimental plot was estimated using wax-discs, as described by Thompson et al. (1997). Three wax discs (1 cm in diameter) were used per experimental plot. Small holes slightly larger than the discs were made with a small drill and filled with epoxy putty where discs were deployed for periods of 10 days every four months. Discs were examined Chapter 3 84 Table 3. 3. Analysis of Wald statistic table from Generalized Estimating Equations (GEE) examining the main and interactive effects of Nutrient and Herbivory intensity on number of native species, number of non-indigenous species (NIS), Percentage cover of crustose algae, Percentage cover of leathery algae, Percentage cover of corticated algae, Percentage cover of articulated calcareous algae, Percentage cover of ephemeral filamentous and foliose algae (Ephemeral). Data from February 2010 to March 2011. Number of native species Number of NIS Crustose algae Leathery algae Corticated algae Articulated calcareous Ephemeral df χ2 P χ 2 P χ 2 P χ 2 P χ 2 P χ 2 P χ 2 P Nutrients (N) 2 34.4 <0.0001 14.45 0.001 5.76 0.056 16.44 0.0003 3.63 0.163 2.1 0.350 14.08 0.001 Herbivory (H) 2 6.5 0.039 2.01 0.365 5.81 0.055 8.73 0.013 14.37 0.001 2.2 0.330 2.17 0.338 N x H 4 2.6 0.635 7.10 0.132 23.63 <0.0001 8.25 0.083 14.12 0.007 48.4 <0.0001 9.84 0.043 Chapter 3 85 Fig. 3. 5. Mean (+SE, n = 30) percentage cover of crustose, corticated, articulated calcareous and ephemeral species across Nutrient and Herbivory intensity treatments over 6 sampling dates. Means calculated by averaging data from each plot unit (5) over time (6). Abbreviations as in Figures 3.1 and 3.3. Chapter 3 86 3.5. Discussion By maintaining similar propagule pressure and disturbance, this study gave evidence of the importance of inherent features of the invasive species (invasiveness) and recipient communities (invasibility) to invasion success. A key finding from our study was that invasion success of Sargassum muticum was governed by several variables and the magnitude of each effect depended on the invasion metrics used (i.e. abundance, growth and survivorship of recruits). Additionally, top-down and bottom-up forces induced specific regulation in different stages of the invasion. Over the initial stage of the invasion, early recruitment of S. muticum was positively influenced by low nutrient enrichment whereas high nutrient enrichment and grazers had limited effects. Two justifications could explain the lack of response to high nutrient enrichment during S. muticum recruitment. Used to low summer nutrient concentration, S. muticum may use nutrients more efficiently at low concentrations than at high concentrations. This fact could explain their greater survival at low nutrient environments. This first explanation is, however, not in agreement with previous manipulative experiments with S. muticum in which higher nutrient enrichment played an important role in controlling community invasibility (Sánchez and Fernández, 2006). Another possible explanation is based on the community structure at the time of the invasion. Over the artificial invasion period a substantial increase in the percentage cover of fast-growing ephemeral species can be observed (Supplementary data, Fig. B3.1). The observed bloom of ephemerals could have prevented the settlement and/or early survival of S. muticum by reducing available nutrient resources or simply by settling over the germlings preventing their survival. A study along the rocky shores of New England suggested that Enteromorpha could have outcompeted the long-lived Chondrus crispus by settling on top of its thallus which became bleached and then disappeared (Lubchenco, 1978). A recent study in northern Portugal (Monteiro et al., 2012) described a varying effect of grazers’ exclosure depending on shore-height, suggesting differences in grazing pressure at midand low intertidal. Low intertidal habitats showed no effect of grazer exclosure (Monteiro et al., 2012), which could explain our results over the recruitment period. Differences in the assemblage and abundance of herbivores between habitats could be responsible for the patterns found (Monteiro et al., 2012). In contrast, at the end of the experiment the presence of grazers together with specific traits of the recipient assemblage influenced the invasion success of the invader. Here, top-down control, together with number of non-indigenous species and percentage cover of ephemerals were significant predictors for the number of S. muticum recruits, a proxy of post- Chapter 3 87 settlement survival. Additionally, nutrient enrichment might have promoted invasion success of S. muticum, though indirectly. Nutrient availability may modify food preference of grazers (Russell and Connell, 2005) and so bottom-up factors may be linked to topdown forces (Menge, 1992). The fast-growing strategy of S. muticum linked to its possible rapid nutrient uptake might induce grazing preference over later S. muticum recruits. The mechanisms promoting S. muticum recruitment are suggested to be the same behind the expansion of native and other non-indigenous species (Davis et al., 2000). Hence, competition for resources may explain our results if we assume NIS in general to be better competitors than natives. Moreover, ephemeral species could act as a barrier to invasion. Accordingly, our results indicated that the cover of ephemerals negatively affected the number of recruits, reinforcing this idea. Furthermore, no significant growth response of S. muticum to herbivory was observed in our study. The fact that there was no effect of herbivory on growth when it negatively affected the abundance of recruits could suggest that predation only influences S. muticum directly in the germling stage and decreases with growth and maturation. Several studies have indicated that phenolic compounds can act as a chemical defense in marine brown algae (Steinberg, 1985) and suggested that those defensive compounds increase with growth and maturation (Van Alstyne et al., 2001). Another possible explanation could be that the effect of grazers involves consumption of the entire germling, thus affecting abundance but not growth of the remaining individuals (Diaz-Pulido and McCook, 2003). We should also mention that the method used to exclude grazers was effective excluding those grazers that crawl and creep on the rock but was probably not efficient with those species that swim or "jump" from individual to individual. Strong top-down control in marine habitats may influence species production or biomass (Bracken and Stachowicz, 2007; Jenkins et al., 2008). This control was not evident in our study, where the number of NIS in the recipient community was the key predictor of the overall final S. muticum biomass. The presence of NIS has been suggested to create facilitative interactions among invaders, i.e. an “invasion meltdown” process. Invasion meltdown describes positive interactions among invaders with consequent increased probability of survival and/or ecological impact at the community level (Simberloff and Von Holle, 1999). In the present study, however, not only was no facilitation apparent but a negative impact on the final biomass of the invader was found, linked to the presence of NIS on the recipient assemblage. Consistently, the majority of described interactions between NIS are those in which individuals of one species are benefited and those of the other are negatively affected (Simberloff and Von Holle, 1999). Here, the presence of one NIS species seems to be an advantage to S. muticum biomass production, whereas the Chapter 3 88 presence of more than one species had a negative impact. These results suggested mutual interference or competition between NIS, giving evidence of a “biotic resistance” process. The mechanisms supporting such interference or competition were not, however, addressed by the present experimental design. Quantification of population-level impacts of NIS on one another are of great interest and should be addressed in future studies (Parker et al., 1999). Species coexistence in natural communities is driven by resource partitioning (Chesson, 2000; Schoener, 1974) which is dependent on species biological traits. Over the experimental period leathery algae responded to both manipulated factors independently, whereas all other functional groups showed interactive effects. Surprisingly, there was a substantial increase in leathery cover with herbivorous exclosure. This functional group is expected to be less palatable and more defended than other functional groups such as turf algae (Littler et al., 1983) and so significant top-down effects were not expected. Negative grazing effects on leathery species could also be related to selective herbivory. A recent study suggested that by grazing on specialized tissues, herbivores may reduce N uptake and thus have a greater effect on communities than expected (Bracken and Stachowicz, 2007). Conversely, it has been suggested that by grazing filamentous/turf algae to low biomass levels, herbivores may then be forced to feed on less palatable algae (Burkepile and Hay, 2006). This phenomenon may explain the observed effect of herbivores on leathery cover. In this experiment, there was significantly more grazing (reduction in percentage cover) over corticated algae exposed to elevated nutrients than ambient nutrients concentration. In contrast to corticated algae, crustose algae increased in percentage cover when elevated nutrients and high grazing intensity were present in combination. These results are in agreement with previous experiments in which crustose coralline algae were more abundant under high herbivore pressure (Belliveau and Paul, 2002). Although calcareous macroalgae may be less nutrient limited than fleshy species (Delgado and Lapointe, 1994), the presence of herbivores resulted in higher percent cover probably due to higher consumption of fleshy algae. By limiting algal biomass, grazing promotes patches of bare rock which are then quickly occupied by crustose coralline algae (Wai and Williams, 2005). In this study, both top-down and bottom-up forces played a significant role in structuring assemblages. Our results suggested that the susceptibility of a community to invasion might be affected by changing specific functional traits of the recipient community, due to potential indirect effects of bottom-up and top-down forces (Leibold et al., 1997). These results may, however, be context-dependent as indicated by a recent study across Chapter 3 89 regions of contrasting productivity in Australia (Bulleri et al., 2012). Additionally, a study in New Zealand’s perennial short tussock grasslands suggested that the characteristics of the resident community were more critical in determining invasion success than fluctuating resource availability (Walker et al., 2005). Although it may appear contradictory, our results agree and reinforce this perspective. The dynamics and structure of the recipient community are key components of an invasion success. Hence, by shaping community structure, main and interactive effects of bottom-up and top down forces may be considered indirect factors acting on invasion. 3.6. Acknowledgments We would like to thank I. Gestoso, E. Cacabelos, M. Incera, M. Rubal and P. Veiga for assistance over experimental set-up. M. Pérez and M. Valdes helped maintaining experimental conditions. M. Vaz-Pinto and I. Gestoso were of valuable help over the period of artificial invasion. We thank Stuart Jenkins for helpful English revision and two anonymous referees for valuable comments. F. Vaz-Pinto was supported by FCT scholarship SFRH / BD / 33393 / 2008. This work was funded by the Spanish Government through the Ministry of Science and Innovation-FEDER (CLINVA Project CGL200907205). Chapter 3 90 3.7. References Anderson MJ, Gorley RN, Clarke KR (2008) PERMANOVA+ for PRIMER: Guide to software and statistical methods. 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Nature 417:848-85 Chapter 4 101 Neighbourhood competition in coexisting macroalgae: the native Cystoseira humilis vs the invasive Sargassum muticum Fátima Vaz-Pinto1,2, Brezo Martínez3, Celia Olabarria4, Francisco Arenas1 1 Laboratory of Coastal Biodiversity, CIIMAR – Centro Interdisciplinar de Investigação Marinha e Ambiental, Universidade do Porto, Rua dos Bragas 289, 4050-123 Porto, Portugal 2 ICBAS – Instituto de Ciências Biomédicas Abel Salazar, Universidade do Porto, Rua de Jorge Viterbo Ferreira, 228, 4050-313, Porto, Portugal. 3 Área de Biodiversidad y Conservación, Universidad Rey Juan Carlos, c/ Tulipán s/n, 28933 Móstoles, Madrid, Spain 4 Departamento de Ecología y Biología Animal, Facultad de Ciencias del Mar, Universidad de Vigo, 36310 Vigo, Spain. Chapter 4 102 4.1. Abstract The introduction of non-indigenous species (NIS) is expected to have negative effects on native competitors, particularly between similar native and invasive species. We examined the role of species identity and density in the growth of marine macroalgal assemblages at the level of individual neighbourhood (a central, focal individual and its neighbours). A spatial neighbourhood approach was used to quantify intraand interspecific competition between two similar macroalgal species (Fucales: Sargassaceae), using the invasive Sargassum muticum (Yendo) Fensholt and the native Cystoseira humilis Schousboe ex Kützing known to co-occur in intertidal rocky shores from NE Atlantic. A total of 64 plates of monospecific and mixed macroalgal assemblages of S. muticum and C. humilis with two different densities of neighbours were created. Using either S. muticum or C. humilis as the focal species, we monitored the focal species’ growth responses despite the presence of the neighbours in field conditions and we assessed nutrient uptake rates in laboratory. Our results showed that C. humilis grew at a higher rate showing N accumulation if surrounded by S. muticum, whereas the invader showed growth and N accumulation at low density conditions. In addition, C. humilis presented greater N uptake rates comparing to S. muticum suggesting better competitive potential to exploit high N transient pulses. Our results support the fact that because NIS in introduced habitats can be highly abundant, many NIS may have been wrongly considered superior competitors. Keywords: Neighbourhood competition; NIS; macroalgae; Sargassum muticum; Cystoseira humilis; nutrient uptake rates Chapter 4 103 4.2. Introduction iological introductions in marine ecosystems are widely reported and have recently increased all around the world, mainly due to human activities such as international shipping, aquaculture and aquarium activity (Bax et al., 2001; Carlton and Geller, 1993). Understanding changes in biodiversity as an impact of species invasions is a major challenge in ecology, particularly in coastal waters which experience a high susceptibility to such events and are amongst the most invaded systems on the planet (Grosholz, 2002). When reaching high densities in a novel community, nonindigenous species (NIS) are expected to have negative effects on native competitors. Particularly, strong competitions arise between ecologically similar native and invasive species (Dudgeon et al., 1999). However, unlike predators and parasites, competitors have rarely been documented to cause extinctions. Thus, when competition occurs it does not necessarily imply that it will have vital consequences for ecological communities (Goldberg and Barton, 1992). Notwithstanding, a review of the mechanisms underlying the impacts of exotic plant invasions revealed that competition is, in most of the cases, suggested to be the process responsible for the registered impact in native communities (Levine et al., 2003). Competition, together with disturbance and physical stress, are key factors in determining distribution and abundance patterns of marine macroalgal assemblages (Paine, 1990). Thus, competitive interactions between NIS and native species need to be investigated in order to predict invasions outcome and the possible consequences on native communities (Brenneis et al., 2010; Staehr et al., 2000; Vermeij, 1991). Research on biological invasions tried to look for common patterns of species features that increased the likelihood of a successful invasion (Nyberg and Wallentinus, 2005). For instance, species with a wide environmental tolerance, high growth rates and reproductive output are expected to have a greater establishment success while traits such as morphology and habitat effects are linked to ecological impacts (Nyberg and Wallentinus, 2005). Also, the susceptibility of a community to invasion has been suggested to increase with high resource availability (Davis et al., 2000). This pattern may be associated with the previously mentioned traits generally shared by successful NIS that need high supply of resources. However, a recent study in resource-limited habitats registered higher rates of carbon assimilation in invasive species, demonstrating that they can also outperform cooccurring native species in resource use efficiency in habitats of limited resources (Funk and Vitousek, 2007). In the marine environment, macroalgal assemblages depend on resources such as space, nutrients and light (Carpenter, 1990). Nutrients are the primer B Chapter 4 104 resource limiting macroalgal growth in temperate marine ecosystems (Hanisak, 1979) and is an important bottom-up factor controlling the performance and structure of coastal macroalgal communities (Sánchez and Fernández, 2006; Worm and Lotze, 2006). Macroalgal species have developed different ecological strategies to cope with the seasonal nutrient limitation intimately related to their life-cycle. Perennials with slow growth present low uptake rates and develop large nutrient stored pools (Martínez et al., 2012), which allow them to be relatively tolerant to the nutrient summer limitation and dominate in pristine temperate geographic areas (Pedersen et al., 2010). Opportunistic ephemeral macroalgae, on the other hand, with high nutrient uptake rates describe an ecological strategy suited to sites or seasons of high nutrient and light availability to sustain their fast growth and reproductive maturation (Phillips and Hurd, 2003). Other annual species, known as summer-annuals, sustain active growth during late spring to summer, achieving their maximum size in these periods (Sears and Wilce, 1975). These species don’t present significant nutrient stored pools like previously mentioned for other perennials and they are not bloom-forming species. For example, the summer-annual Nemalion helminthoides showed intermediate N content and uptake affinity per unit of biomass compared to the perennial slow-growing strategy and the opportunistic ephemeral strategy, together with a dynamic-inducible activation of nitrate uptake (Martínez et al., 2012). Thus, summer-annuals describe an ecological strategy that takes advantage of light availability in periods of limiting nutrients that prevent the active growth of perennials and the survival of many ephemerals (Sears and Wilce, 1975). The brown macroalga Sargassum muticum (Yendo) Fensholt 1955 is one of the most well studied invasive seaweeds. Commonly distributed throughout the intertidal, S. muticum regularly invades the habitats of algal species from the family Cystoseiraceae (Arenas et al., 1995; Fletcher and Fletcher, 1975; Wernberg et al., 2000). Previous studies indicated variable impact of S. muticum on native assemblages (Britton-Simmons, 2004; Olabarria et al., 2009). Particularly on the north and northwestern Iberian Peninsula, the impact of the invasive species was evident on three distinct understory morpho-functional groups: filamentous, leathery and foliose algae (Olabarria et al., 2009; Viejo, 1997). These results suggest a higher impact of an invader species over morpho-funcional groups with similar resources use. Cystoseiraceae and Sargassaceae are closely related families (Rousseau and De Reviers, 1999) and several studies have investigated differences in life-history strategies between algal species from the family Cystoseiraceae and Sargassaceae (e.g., Arenas et al., 1995; Engelen and Santos, 2009; Engelen et al., 2008). Moreover, it has been suggested that indigenous Cystoseiraceae can be displaced by S. muticum (Fletcher and Fletcher, 1975; Viejo, 1997; Engelen and Santos, 2009). Chapter 4 105 On the semi-exposed rocky shores of southern Portugal, S. muticum co-occurs with Cystoseira humilis Schousboe ex Kützing 1860 in intertidal rock pools. Observational results suggested lower population growth of S. muticum in pools where the native algae C. humilis was dominant, compared to pools where S. muticum was dominant (Engelen and Santos, 2009). Mechanisms of these interactions, however, have not been fully studied. Because the intensity of interactions among sessile individuals varies with the spatial arrangement of competitors, the relative proximity of neighbouring individuals may be an important factor influencing the interactions among seaweeds (Carpenter, 1990; Kim, 2002a). Competitive ability of an organism include its ability to deplete resources in detrimental of the neighbours as well as its response to the presence of competitors, growth, survival and reproductive ability in the presence of neighbours (Goldberg, 1987). The morphology of competing individuals, biomass, population density and neighbour distance affect resource competition for light, nutrients and space (Carpenter, 1990; Choi and Norton, 2005; Kim, 2000b). The present study used both field and laboratory experiments to evaluate the relative competitive abilities of S. muticum and C. humilis by comparing the species’ growth responses and nutrient uptake rates. The specific objectives were to (1) compare the relative effects of intraand interspecific competition on growth, 2) determine whether density effects influence individual growth, and (3) determine the nutrient uptake kinetics from each species. In general, higher density assemblages are expected to have reduced growth, survivorship and reproductive output compared with those at lower densities. Also, C. humilis is expected to show perennial slow-growing strategy and S. muticum may be compared to seasonal annuals, with low and high nutrient uptake rates, respectively. 4.3. Materials and methods 4.3.1. Field study site This study was carried out in the intertidal of Praia de Moledo (Northern Portugal, 41º50′22″ N, 8º52′30″ W) from February to August 2011. Moledo is an exposed site with a semi-diurnal tidal regime, with the largest tidal range during spring tides of 3.5-4 m. Sea surface temperatures vary between 13ºC and 20ºC. The site can be described as a gently sloping granitic rocky shore, with a wide variability of fauna and flora. The experiment was performed in mid-intertidal and in lower-intertidal rock pools. Chapter 4 106 4.3.2. Sargassum muticum and Cystoseira humilis Sargassum muticum is native to East Asia and it is considered an invasive species all around the world (Critchley et al., 1983; Norton, 1976). In Europe, S. muticum was first introduced to the English and French coasts in the early 1970s (Critchley et al., 1983; Druehl, 1973). Its present distribution in Europe range from the Atlantic coast, from Portugal to Norway, to single introductions in the Mediterranean, l’Etang de Thau and the Lagoon of Venice (Curiel et al., 1998; Engelen et al., 2008; Wallentinus, 1999). Populations of this species are distributed mainly in sheltered or semi-exposed rocky shores, although it may also attach to hard substrates on soft-bottoms, such as stones or shells (Strong et al., 2006). The species is monoecious, self-fertile and it is considered a pseudo-perennial species consisting of a perennial basal holdfast and annual main and lateral branches (Arenas and Fernández, 1998; Norton, 1976). Additionally, the presence of gas-bladders maintains the thallus erect and closer to the light in the water column (Arenas et al., 1995). S. muticum has a vertical distribution from mid-intertidal pools to the shallow subtidal level. Cystoseira humilis is a member of the family Sargassaceae1and can grow up to 150 cm high. It lives attached to the substratum by a compact basal disc. The native brown species are the dominant algae in most intertidal rocky pools along the Atlantic south-west coast of Portugal (Engelen and Santos, 2009). It is described to occur in pools and in shallow standing waters in the eulittoral zone, in moderately wave-exposed sites (Gómez Garreta, 2001). C. humilis has a perennial life history (Engelen and Santos, 2009) and is also a monoecious species. Macroalgal assemblage plates and Experimental design Individuals from S. muticum and C. humilis were collected in the rocky intertidal of Viana do Castelo (41º42’25’’ N, 8º51’42’’ W, Northwest coast of Portugal) and Sines (37º53’12’’ N, 8º47’43’’ W, Southwest coast of Portugal), respectively, and taken to the laboratory. Assemblage plates with different species identity and density of both species were created in the laboratory. We constructed experimental plates made of PVC of two different sizes containing either individuals of S. muticum, C. humilis or both at two different final densities. We set 25 individuals per 500 cm2 in plates of 27 x 27 cm and 25 individuals per 250 cm-2 in plates of 20 x 20 cm, for low and high density treatments, 1 Cystoseiraceae according to traditional classification on morphological grounds (position of receptacles) but transferred to Sargassaceae based on DNA sequence (Draisma et al., 2010). Chapter 4 107 respectively. This gives final densities of 20 and 40 individuals per 400 cm2, respectively. These values mimic those observed in field conditions for the native species (16 to 49 holdfasts per 400 cm-2 with a mean (±SD) of 31 (±12), from three scraped quadrates of 20 x 20 cm). Experiments comparing species competition or performance in monoculture and mixed assemblages use additive or substitutive designs depending on whether the density is kept constant or not in the experimental plots (Snaydon, 1991). Here, we used a neighbourhood approach to analyze seaweed species competition by relating the performance of individuals of the focal species to the identity and density of neighbouring species. Using plots with similar size but different density will change the overall biomass of each plot. Thus, to avoid confounding effect among overall plate biomass and density (Underwood, 1997) we changed plate size to achieve different densities but same number of individuals per plate (25 individuals per plate). Thus, monospecific assemblages consisted of 25 individuals of the same species and the one in the center was considered the focal individual, whereas mixed assemblages had 1 focal individual of one species in the center, surrounded by other 24 individuals of the other species. For example, the low density treatment with S. muticum as the focal species had 1 individual of this species and 24 of C. humilis per 500 cm2 (Fig. 4.1). The area of individuals’ distribution was inferior to the PVC area because a 2 cm border was left intact to allow for plates deployment in the field. A total of 64 assemblage plates were built (n = 8). Fig. 4. 1. Example of a mixed assemblage plate with Sargassum muticum as the focal species. It shows 1 individual of S. muticum surrounded by 24 individuals of Cystoseira humilis. Chapter 4 108 Algal length, recorded at the time of transplanting, was 14.96 ± 0.70 cm for S. muticum individuals and 8.46 ± 0.40 cm for C. humilis individuals (mean ± SE, n = 32). Biomass of both species was estimated following Åberg (1990), from length of primary lateral and perimeter of each alga individual. We used data of dry weight and size from 40 individuals previously collected from each species to construct a regression model that allowed our estimations: S. muticum: DW = 0.4092 + 0.006V (R2 = 0.98, P < 0.001) C. humilis: DW = 0.2636 + 0.007V (R2 = 0.82, P < 0.001) where DW is dry weight and V is the specimen volume (length x diameter2). After the laboratory experimental manipulation, synthetic assemblages were transported to Praia de Moledo where they remained for 6 months in intertidal rock pools. Assemblage plates were randomly placed and screwed to the bottom of the rock pools. We measured short-term changes in biomass to estimate growth during two sampling events, over a 6month period. For biochemical analysis of C and N tissue content, we collected approximately 2-5 cm of non-reproductive tissue from the focal species at each assemblage plate at the end of experimental period. Algal material was dried at 60ºC for 48 h and total C and N determined using a Carlo Erba CHNS-O Elemental Analyser (Model EA1108). The percent tissue C and percent tissue N were standardised to algal dry weight (g). 4.3.3. Nutrients over a tidal cycle To assess the variation of nutrient availability in intertidal rock pools, seawater samples were collected in August from three rock pools over a tidal cycle (5h) in Praia de Moledo. The first water sample was taken as soon as the tide was low enough and water exchange was no longer observed. Then, duplicate water samples were taken every 30 minutes until water exchange was again observed and the tide was high. Water samples were filtered in situ using portable microfiber filters (Fisherbrand® MF 300) and acid cleaned 50-ml plastic syringes. Immediately after collection, samples were placed on ice, returned to the laboratory in darkness, and frozen at -20 ºC until analysis (less than three months after) of nitrate + nitrite, ammonium and orthophosphate. Chapter 4 109 4.3.4. Nutrient uptake experiment in the laboratory Collection of samples and preincubation Vegetative individuals of Sargassum muticum and Cystoseira humilis were collected in the rocky intertidal of Viana do Castelo and Sines, respectively, in April 2012. Collection was made three days before assaying their uptake kinetics. In the laboratory, the material was pre-incubated in 2 L Erlenmeyer flasks filled with artificial seawater (Red sea salt) enriched with Von Stosch's (VSE) medium (Ott, 1965), but the specific nutrient (N for ammonium and nitrate or P for the phosphate uptake experiments) to be assayed. Adequate mixing was assured by bubbling filtered air into the culture medium. Flasks were left in a walk-in culture chamber at 15 °C, with constant photon flux density (approx. 150 μmol photons m−2 s−1) and photoperiod (12:12; Light:Dark) until the experiment. Uptake experiments Nutrient uptake kinetics was assessed by measuring the decrease of nutrient concentration at different time intervals in 250 ml Erlenmeyer flasks filled with 200 ml of medium set at different initial substrate concentrations as in Martínez and Rico (2004) and Martínez et al. (2012). Algal fronds of each species, 1 g fresh weight (FW), were incubated in 250 ml Erlenmeyer flasks: 12 flasks with increasing nutrient concentration plus 3 blanks (no seaweed) as controls. In total 30 flasks were used per nutrient experiment (15 per species). Flasks were randomly arranged into a multi-stirrer magnetic plate (IKA-WERKE). Irradiance was provided with fluorescent lamps (cool white F18W/840) to a final value of 425 μmol photons m−2 s−1 and the temperature was controlled inside the walk-in chamber. The nutrient depletion of the medium was determined by nutrient analysis of water samples. Water samples (10 ml) were taken before the addition of the algae and then at 15, 30, 60, 120, 180, 240 and 300 min. After the experiment, algal material was oven-dried at 60 °C to a constant weight (48 h), for determination of dry weight (DW). Analyses of NH4+, NO3− and HPO42were carried out using a Skalar Sanplus++ segmented flow autoanalyser, at CIIMAR, Porto. Blank control flasks registered minor concentration changes and thus no corrections were made to the algal uptake rates. Mean (± SE) differences between initial and final values (0-300 min) in controls were 3.30 (± 1.57) µM for ammonium, 1.17 (± 0.67) µM for nitrate and 0.34 (± 0.33) µM for phosphate experiment. Chapter 4 116 4.4.3. Nutrient uptake The native C. humilis showed the greatest uptake rates concerning the N sources (Fig. 4.5). Nitrate uptake increased linearly with substrate concentration for both species, except for S. muticum during the period 60-120 min when a saturation response was observed (Fig. 4.5a, b; Table 4.3). Nitrate was exhausted earlier (240 min) by C. humilis and at higher rates (compare axes in figures 4.5a, b), suggesting higher nitrate uptake potential of C. humilis than S. muticum. The ammonium uptake response was linear for both species (Fig. 4.5c, d; Table 4.3). Ammonium was totally consumed after 30 min by C. humilis and 60 minutes later by S. muticum. The maximal uptake rates were recorded for C. humilis showing higher ammonium uptake potential. Phosphate uptake rates were similar for both species (Figs. 4.5e, f). Phosphate uptake response for C. humilis followed a saturated uptake response at the time intervals 15-30 and 60-120min, and a linear uptake response at 30-60, 120-180 and 180-240 min (Table 4.3). At 240-300 min, no phosphate concentration remained in C. humilis experiment. S. muticum did not show a complete phosphate uptake over the experimental period (300 min), contrasting with C. humilis. Up to the time interval 120-180 min S. muticum showed similar, although lower, uptake response pattern as C. humilis. The uptake response at 180-240 min described a Michaelis–Menten saturation pattern and the last time interval showed a linear response (Fig 4.5e, f; Table 4.3). Chapter 4 117 Fig. 4. 5. Uptake rates of phosphate (a, b), nitrate (c, d) and ammonium (e, f) for C. humilis and S. muticum. Kinetics are shown as a function of the substrate concentration at the beginning of 7 different time intervals (as different symbols). X, values not included in the fit. Chapter 4 118 Table 4. 3. Estimations (standard error) of the uptake kinetic parameters for S. muticum and C. humilis using either linear or Michaelis–Menten (MM) functions. Vmax is expressed in µmol g DW−1 h−1 and Ks in µM. Interval Sargassum muticum Cystoseira humilis (min) Slope (linear) Vmax (MM) Ks (MM) n Slope (linear) Vmax (MM) Ks (MM) n Ammonium 0-15 1.08 (0.34) 9 2.66 (0.21) 12 15-30 1.40 (0.33) 9 1.54 (0.23) 9 30-60 0.32 (0.06) 8 - - - - 60-120 - - - - - - - - Nitrate 0-15 0.32 (0.08) 10 1.16 (0.06) 10 15-30 0.63 (0.05) 9 0.55 (0.07) 10 30-60 0.10 (0.04) 9 0.11 (0.04) 8 60-120 15.74 (2.18) 34.82 (8.08) 10 0.46 (0.02) 9 120-180 0.51 (0.03) 8 0.21 (0.03) 12 180-240 - - - - 1.08 (0.16) 6 240-300 0.08 (0.02) 6 - - - - Phosphate 0-15 - - - - - - - - 15-30 9.16 (2.24) 16.72 (7.50) 10 5.98 (1.57) 7.63 (4.23) 11 30-60 0.26 (0.03) 9 0.15 (0.01) 10 60-120 2.36 (0.31) 4.83 (1.73) 10 2.44 (0.96) 5.19 (4.84) 11 120-180 0.07 (0.01) 8 0.17 (0.03) 12 180-240 3.75 (1.31) 13.61 (9.38) 8 0.20 (0.03) 12 240-300 0.14 (0.02) 8 - - - - Chapter 4 119 4.5. Discussion Small-scale spatial distributions of individuals belonging to different species have important consequences to the dynamics of the whole community. In particular, sessile organisms compete for space and major resources as nutrients, mainly at the neighbourhood scale (Bonan, 1988; Connell, 1983; Naeem et al., 1999). Moreover, macroalgal assemblages in tidepools are very patchy systems at very small scales, where extremely small spatial scales of variation seem to be among the most important sources of heterogeneity (Archambault and Bourget, 1996). Thus, the growth of a focal organism should be related to the size and identity of the neighbouring organisms. We identified in this study an effect of the identity of the neighbours in the native Cystoseira humilis growth and nutritional state whereas the invasive Sargassum muticum focal species showed a density-dependent effect unrelated to the identity of the neighbour species. The former grew at a higher rate showing N accumulation if surrounded by S. muticum, whereas the later showed growth and N accumulation at low density conditions. In addition, C. humilis presented greater N uptake rates comparing to S. muticum suggesting better competitive potential to exploit high N transient pulses. In marine coastal systems, aggregations of predominantly single species of macroalgae, both at the intertidal and at the subtidal, have been reported worldwide (Moore and Seed, 1985). Nevertheless, marine macroalgae are susceptible to competition for space, nutrients and light because those are often in short supply in coastal ecosystems (reviewed in Carpenter, 1990) and vary at small scales. Our results showed a negative density-dependent effect on growth for S. muticum, independently of neighbour identity. These results were supported by biochemical analyses that indicated potential growth limitation by nitrogen availability (Corzo and Niell, 1991), i.e. a C/N ratio above 15 (see Hanisak, 1983), in monospecific assemblages and high density assemblages of S. muticum. Negative effects of density have already been previously registered for S. muticum and other intertidal macroalgae (e.g. Arenas et al., 2002; Viejo and Åberg, 2001). Our study suggests lower accumulation of N at higher densities, limiting growth. Other studies have also registered a positive correlation between individual length and density for fucoids and Laminaria spp. (Schiel, 1985). Also, because the present experiment was focused on post-recruitment competition, we exclusive used thalli of similar size (≈ 15 cm). However, natural stands of S. muticum include different adult and juvenile stages, i.e. mixture of sizes and biomass (Baer and Stengel, 2010; Strong and Dring, 2011). Thus, it has been suggested that high abnormal densities of S. muticum might led to high rates of frond erosion (Strong and Dring, 2011), possibly decreasing biomass as observed in the Chapter 4 120 present study. Furthermore, our results showed that density-dependent competition on S. muticum focal species can occur between individuals of the same species or among individuals of different species. Thus, our results suggested that density of neighbourhood assemblages was the key mechanism affecting biomass production of S. muticum focal species. Contrasting, the present study also revealed that the density of macroalgal assemblages did not represent a negative effect for C. humilis biomass production. Our results are in agreement with a previous study which tested for competition in mixed canopies of S. muticum and Saccharina latissima at subtidal communities in Strangford Lough (Strong and Dring, 2011). A strong intra-specific competition has been reported for S. muticum in high densities whereas no density-dependence was observed for the perennial Saccharina latissima (Strong and Dring, 2011). As mentioned previously, dense monospecific stands of C. humilis are common along the south-west coast of Portugal (Engelen and Santos, 2009), suggesting that density is not a problem for the species. Moreover, C. humilis focal species showed a greater increased in biomass when in mixed assemblages with S. muticum than in monospecific assemblages. These results may either suggest that S. muticum neighbours display a positive interaction with C. humilis when in mixed assemblages or that C. humilis presents greater intraspecific competition. Competition between closely related species, e.g. Family: Sargassaceae, is a deterministic factor in natural selection (Darwin, 1875). Positive interactions of neighbouring species on focal species may occur when the presence of one species ameliorates harsh environmental conditions, e.g. by reducing thermal, nutrient, predator or light stress (e.g. Bertness et al., 1999; Brawley and Johnson, 1991). Generally, the presence of algal canopy in rocky intertidal assemblages is known to reduce thermal stress at high tidal heights, whereas at the low intertidal border canopy effects are negative or neutral (Bertness et al., 1999; Bruno et al., 2003). Canopy effects in tidepools are expected to be similar to low intertidal border, as thermal stress is not so evident. Thus, thermal stress reduction may not be related to the positive interaction observed in the present study. Predation and competition are, however, both strong structuring forces in macroalgal assemblage dynamics (Edwards and Connell, 2012; Keane and Crawley, 2002). Nutrient limitation, in particular, may be one of the most important limitation factors for intertidal communities and tide pools due to tides, i.e. lack of nutrient renovation. In the study area, nitrate and phosphate in tide pools showed a linear decrease in concentration over a tidal cycle. Macroalgal productivity is sustained by the acquisition and utilisation of nutrients, particularly nitrogen and phosphorus. Thus, subsidy rates are a key determinant of the intensity of species interactions (Menge et al., 2003), together with net uptake rates capacity. In terms of nutrients, fast-growing opportunistic species show a high nutrient Chapter 4 121 uptake potential, while slow growing perennials have lower uptake rates and large nutrient store capacity (Lobban and Harrison, 1997; Martínez et al., 2012). Most macroalgae register higher uptake rates for the N form NH4+ than for NO3-, in accordance with our results (Phillips and Hurd, 2003, 2004; Rees, 2003). Surprisingly, we found remarkably higher nutrient uptake rates for the native perennial C. humilis, with values comparable to nutrient uptake potential previously registered for fast-growing opportunistic species (Martínez et al., 2012). A possible explanation may be related to the nutrient availability from the alga recipient region. C. humilis individuals were collected from Sines (SW Portugal) while S. muticum comes from Viana do Castelo (NW Portugal), with average (±SE) NO3concentration in 2011 of 1.52 ±0.46 µM and 4.31 ±0.56 µM, for Sines and Viana do Castelo, respectively (unpublished data). Generally, nutrient-limited algae show higher uptake rates (Lobban and Harrison, 1997), and thus our results might be explained by the fact that C. humilis may be nutrient limited in SW Portugal. Observational field results may reinforce this theory as C. humilis showed a maximum length of 44.8 cm in NW Portugal, while the maximum value recorded from SW Portugal was 18.8 cm (pers. obs). S. muticum, on the other hand, showed N-uptake rates comparable to summerannual species (Martínez et al., 2012). The fact that S. muticum did not demonstrate opportunistic characteristics in uptake rates is in agreement with a recent study which stated that at a population level, growth of S. muticum when invading intertidal rock pool habitats follow K-selected traits (Engelen and Santos, 2009). Also, because light can be absorbed and/or scattered by the macroalgae themselves, both direct and indirect competition with each other may occur (Edwards and Connell, 2012). Indirect effect mediated by grazing can also be an explanation for the observed positive interaction for C. humilis in mixed assemblages. Reduced palatability of C. humilis has been demonstrated with water-borne cues from grazed S. muticum suggesting induced defences (Yun et al., 2012). The same, however, has not been observed for S. muticum (Yun et al., 2012). Because induced defences may save metabolic costs, the greater performance of C. humilis focal species with S. muticum neighbours may be explained by the higher allocation of resources to growth instead of allocated to defence (Agrawal, 2005). Results on nutrient uptake rates obtained in the laboratory may explain the positive effects of S. muticum in the growth of C. humilis in the field. Reduced uptake rates of S. muticum may positively influence the nutrients available for C. humilis by decreasing the conditions of nutrient limitation for C. humilis individuals. Our results are, however, based on nutrient concentrations far higher than those found in the natural environment. These results can then be used to predict a species response to episodic high nutrient pulses such as those following an upwelling or discharge event (Phillips and Hurd, 2004). Chapter 4 122 Upwelling off the northern Portuguese coast typically occurs from late spring to late autumn (Lemos and Pires, 2004), thus coinciding with the study period. Many invasive species are, in fact, weak competitors and its establishment and spread is facilitated by disturbance (Bando, 2006). Nonetheless, the outcome of competitive interactions varies considerably with the age or size at which species interact (Olson and Lubchenco, 1990). Our experiment was a short-term experiment and described the postrecruitment competitive capacity of the native C. humilis and the invasive S. muticum. In conclusion, our study suggests that the growth of the native C. humilis is enhanced by the presence of S. muticum as a neighbour, although it remains unclear if this pattern is due to a greater intravs interspecific competition in C. humilis or due to facilitation by the invader. 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