From seed to seedling: Exploring the early life-cycle stages in alien conifer species

In invasion ecology, ecological factors affecting the early life-cycle stages are crucial to unders- tand the establishment success/failure of alien plants. Conifers have been a study model of in- vasiveness research, but relatively few studies have addressed the ecological barriers from seed to seedling that limit their establishment and whether these barriers are site speci c.

The general aims of this thesis are to provide in- formation about the current invasion status of alien conifers in Europe; to screen which early life-cycle stage limits the establishment of alien species conifer in alien stages life-cycle early seed to seedling: Exploring the From conifers and to compare differences in natural enemy resistance among native and alien conifer species.

••• PhD-Thesis M. Amparo Carrillo-Gavilán

Supervisor: Dra. Montserrat Vilà Planella – July 2011 - Institut de Ciència i Tecnologia Ambientals - Universitat Autònoma de Barcelona (Spain) ••• PhD-Thesis M. Amparo Carrillo-Gavilán

M. Amparo Carrillo-Gavilán

From seed to seedling: Exploring the early life-cycle stages in alien conifer species

© M. Amparo Carrillo-Gavilán, 2011, Bellaterra (Barcelona, Spain) Illustrations: Elena Caballero Cover design: Pedro Peinado Page layout: Rafa Ferreiro From seed to seedling: Exploring the early life-cycle stages in alien conifer species

De semilla a plántula: Explorando los estadios iniciales del ciclo de vida en coníferas exóticas

PhD Thesis M. Amparo Carrillo-Gavilán

Bellaterra, July 2011

Supervisor Dra. Montserrat Vilà Planella Professor of Research Department of Integrative Ecology Doñana Biological Station -CSIC (EBD-CSIC) Sevilla-Spain

Tutor Dr. Jordi Martínez-Vilalta Senior Lecturer Universitat Autònoma de Barcelona (UAB) & Centre for Ecological Research and Forestry Applications (CREAF) Bellaterra, Barcelona-Spain

A mis padres, por estar ahí siempre

A Elizabeth y Juanko, por ser como sois

Plántula, buena plántula, que tras la borrasca te erguiste en desnudez y desaliento, sobre una gran alfombra de hojarasca que removía indiferente el viento (…) Y en esa verde punta que está brotando en ti de no sé dónde, hay algo que en silencio me pregunta o silenciosamente me responde.

Poema del Árbol (modificado) - Antonio Machado

Table of Contents

Agradecimientos 15

Chapter 1: General Introduction 19 Biological invasions 21 Propagule pressure 24 Determinants of invasiveness 24 Determinants of invasibility 25 Conifer species as the study system for this thesis 27 Objectives 28 Thesis Outline 29

Introducción General 31 Invasiones biológicas 33 Presión de propágulos 36 Determinantes del potencial invasor 36 Determinantes de la invasibilidad 37 Las coníferas como sistema de estudio para esta tesis 40 Objetivos 42 Esquema de la tesis 43

Chapter 2: Little evidence of invasion by alien conifers in Europe 45 Abstract 46 Introduction 47 Identity, distribution and invasion status of alien conifers in Europe 48 Scientific literature on conifer invasion in Europe 50 General overview of the alien conifers 51 Scientific perception of alien conifers in Europe 54 Topics of study involving alien conifers in Europe 56 Invasiveness of alien conifers 57 Invasibility of alien conifers 57 Impacts of alien conifers 58 Discussion 59 Conclusions 62 Acknowledgements 62

11 Chapter 3: Comparing seed removal of 16 pine species differing in invasiveness 65 Abstract 66 Introduction 67 Materials and methods 68 Study area 68 Study species 69 Seed removal experiment 71 Seed removal trials with seed remover exclusion 72 Identity of seed removers 73 Results 75 Seed removal experiment 75 Seed removal trials with seed remover exclusion 76 Identity of seed removers 78 Discussion 79 Acknowledgements 80

Chapter 4: Establishment constraints of an alien and a native conifer in different habitats 83 Abstract 84 Introduction 85 Materials and methods 86 Study area 86 Study species 87 Seed removal 88 Seed germination 89 Seedling survival and growth 89 Probability of transition from seed to seedling 90 Results 90 Seed removal 90 Seed germination 92 Seedling survival and growth 93 Transition from seed to seedling 96 Discussion 97 Acknowledgements 99

12 Chapter 5: Early resistance of alien and native pines against two native generalist insect herbivores: no support for the Natural Enemy Hypothesis 101 Abstract 102 Introduction 103 Materials and methods 105 Study species 105 Greenhouse experimental design 106 Chemical analyses 108 Cafeteria bioassays 109 Statistical analyses 109 Results 110 Seedling damage 110 Constitutive and induced chemical defenses 112 Consumption preferences in the cafeteria bioassays 115 Discussion 118 Acknowledgements 120

Chapter 6: General Discussion 123 Current invasion status of alien conifers in Europe 125 Early stages of the plant life-cycle limiting the 126 Differences in natural enemy resistance among native and alien conifers 128

Conclusions 131

Conclusiones 135

References 141

Annexes 161

13 From seed to seedling: Exploring the early life-cycle stages in alien conifer species

14 Agradecimientos

Agradecimientos

Aunque mi profesor no lo sepa, mis clases de Biología en el Instituto me marcaron mucho, siendo una de las horas más esperadas del día. Desde ese momento, muchas de las preguntas sobre la naturaleza encontraron respuesta y esto me causó una gran fascinación. El entusiasmo con el que nos empujaba a curiosear y aprender más sobre la naturaleza caló hondo en mí. Así, mi primer agradecimiento va hacia mi profesor, Don Javier, pues es una suerte dar con docentes tan motivados a que sus alumnos aprendan.

Esta tesis se puede comparar a una planta pues para su desarrollo necesitó de una serie de recursos y condiciones oportunas que la favorecieran, pasando por distintas etapas tanto de éxito como de fracaso. La semilla que marcó el inicio de esta tesis se produjo en un país más al norte, un país lleno de molinos, canales, vacas y holandeses. Un país donde hice mis primeras incursiones en el mundo científico y donde me transmitieron las ganas por seguir aprendiendo sobre la ecología de plantas. Tras mucha insistencia y varios intentos de dispersión, la llegada a Sevilla estuvo favorecida por un recurso en forma de oferta de tesis. Esta oferta sería el primer paso para asentar las bases de mi llegada a la Estación Biológica de Doñana (EBD). Así que, mi primer gran agradecimiento va dirigido a Montse Vilà, mi directora de tesis, que gracias por haber creído en mí desde el principio, por su apoyo y por sus consejos durante estos años ha sido posible el desarrollo de esta tesis que hoy tienes entre tus manos. Y porque sé que muchos han sido sus movimientos para conseguir financiación durante todo este tiempo. A ella, gracias por todo.

Así, en febrero de 2007, llegué con todos mis bártulos a Sevilla y con el entusiasmo por aprender y estudiar sobre las invasiones biológicas. Como lugar de asentamiento me dieron un espacio en una habitación con otros recién aterrizados doctorandos, esto es, varias especies exóticas interactuando unas con otras en un mismo lugar. Allí conocí a David Canal y sus papamoscas; al Airam y su ños; a Alicia y sus pichis; o a Louise y sus mates….en donde más tarde se nos unirían más especies exóticas (Ainara, Ana Montero, Luis el gallego, José Sarasola, Luna,…). Llegamos a ser 13 en dicha habitación y bautizamos el lugar como La comuna. ¡¡Eso si que era optimización del espacio y de los recursos!!

Una vez establecida, empezó la fase de germinación de la tesis y comenzaron los experimentos por aquí y allí. Agradecer al vivero de San Jerónimo de la Junta de Andalucía y a su personal el proporcionarnos un hueco para nuestras plántulas en

15 Agradecimientos su extenso invernadero. En especial, agradecer a Laura Plaza por haber hecho más acogedoras las visitas al vivero, haber ayudado en la gestión, siembra y cuidado de las semillas y plántulas de pino. ¡Qué pena que al final no pudimos salvarlos! También, gracias a Hadrien Lalägue, el franchute, por los buenos momentos compartidos clavando estacas o deliberando sobre como hacer nuestros experimentos. Por supuesto, agradecer a los miembros de mi Comité de Tesis, Pilar Castro, Jorge Castro y Josep Mª Espelta, el haber ido asesorándome durante estos años.

La tesis iba desarrollándose, pero para darle más amplitud tuvo que ponerse a viajar por otros lugares. Así, la primera expansión fue rumbo a Barcelona. Aquí tengo que agradecer a Josep Mª Espelta su acogida en el CREAF y su orientación para desarrollar uno de los capítulos de esta tesis. Y, a Jordi Martínez-Vilalta el ser mi tutor de tesis de una forma accidental. Sobre todo, agradecer a todas aquellas personas que me hicieron más fácil aún la estancia por allí: a Elvira Pérez, Jara Andreu y Mónica Hernanz por su ayuda, trabajo en el campo y por las buenas charlas post-trabajo en el Puig. A Francesc Panyella, mi compañero de piso en la Ciudad Condal, por la buena acogida y las escapadas para conocer más las tierras catalanas. Y, por supuesto, un gran agradecimiento a todo el personal del Parque Natural del Montseny por su buena disposición a que los experimentos fueran siempre para adelante con todos los medios posibles. En especial, gracias a Josep Argemí y Montse Amorós de Fontmartina, y al personal de Can Casades (nunca se me olvidará que las llaves estaban siempre debajo de una piedra).

El tiempo iba pasando, su desarrollo seguía ampliándose y así llegó la segunda expansión con destino, Pontevedra. Por la gran acogida, entusiasmo por enseñar y compartir tanto conocimientos como recursos, otro especial agradecimiento a varias especies nativas que me acogieron por tierras gallegas: Luis Sampedro, Rafael Zas y Xoaquín Moreira. ¡¡Esas comidas de grupo nunca se me olvidarán!! Pero en la puesta a punto y para el desarrollo de todos los experimentos, muchas otras especies nativas intervinieron. Agradecer a Sergio Piñeiro y Chus Villanueva por sus buenos cuidados a todos los bichos. También a Benito Santos, Aída García, Lucía Martínez, Diana Blanco, Aurita Pazos y Enrique Dios por su trabajo y gran dedicación en el laboratorio e invernadero, por las muchas horas que compartimos, por haber hecho el trabajo más ameno y porque sin ellos los experimentos nunca se hubieran acabado según lo previsto. Y, en esta expansión gallega también agradecer a: Luis, el gallego, el haberme mostrado un rinconcito de su tierra; a mi compañera de despacho en Lourizán, Cristina Lazcano, y a mis compañeros de charlas gallegas y de juergas, Raúl de la Mata y Ana Hernández, por los buenos momentos compartidos juntos y por las expediciones a

16 Agradecimientos otras lugares gallegos. Y, sin olvidar, un agradecimiento a los pimientos del padrón y al queso de tetilla, ¡qué hicieron más llevadera mi dieta vegetariana en Galicia!

La tesis siguió creciendo, viajando, participando en congresos, explorando otras opciones (con éxitos y fracasos), expandiéndose por aquí y por allí, pero el rincón más explotado ha sido siempre la EBD. Despachos y pasillos por los que se fue extendiendo, buscando consejos y más recursos para su desarrollo. A mis compañeros de despacho, agradecerles los momentos de charlas, de risas, de consejos y Eco-terapias: al José, a Ana Montero, a Esperanza, a Violeta, a Alicia y, a los recién llegados, Rocío, Juanpe y Pablo González, muchas gracias por todo vuestro apoyo. Agradecer a toda aquella gente de la EBD que, directa o indirectamente, ha contribuido a hacer esta tesis más amena dentro y fuera del edificio. Son muchos a los que me gustaría dedicarles unas palabras pero como no quiero dejarme a nadie atrás, aquí dejo un agradecimiento general: para todos los compañeros con los que todos los días comparto comedor, la mini-mesa, reuniones de precarios y/o charlas al sol; para aquellos investigadores a los que acudí buscando consejo experimental o estadístico; para los que siempre echaron una mano desde administración; para los de seguridad; y para las señoras de la limpieza. Gracias por estos 4 años.

Pero estas líneas tienen que continuar agradeciendo a aquellos que han contribuido al desarrollo no sólo profesional sino también personal que ha habido durante todos estos años. A todos los flamencos apasionados con los que he tenido la suerte de compartir 3 años, desde el Carlitos hasta toda la tribu y piña flamenca, y que han permitido un desarrollo personal lleno de grandes emociones. A Peter Buston, por los momentos de charlas compartidas y por darme la oportunidad de sumergirme para conocer otros mundos. A Ana Montero, un enorme agradecimiento por su amistad y compañerismo, por su apoyo despachil desde el principio y por esos primeros meses de ayuda en el campo ¡¡Cuántas veces nos reiremos de nuestro cuasi- congelamiento acampando para hacer el trampeo de roedores!!! Al Gregor, por llegar a ser tan buen amigo y confidente durante estos años. Por esos grandes momentos de charlas científicas, de descanso bajo el sol, por las escapadas a la playa o quedadas sevillanas. Al Deivid, por ser tan apañao, amigo de charlas y peleas científicas. Mi vivencia por Sevilla no hubiera sido la misma sin ti y sin la más nativa de Sevilla, la Lidia ¡¡Porque llegar al piso siempre ha sido como sentirse en casa!! A este sentimiento también fueron contribuyendo la presencia del gran Juanillo y sus exquisitas cenas; y la del Néstor y su risa. Gracias por vuestro apoyo y por estos años de convivencia.

17 Agradecimientos

A Alex Bertó y Anichi, mis queridos caboverdianos, porque una isla nos unió y espero que eso quede para siempre. Gracias por estar a mi lado siempre con una sonrisa en los buenos y no tan buenos momentos, por vuestra complicidad. Y, sin olvidar, un agradecimiento a tres amigas con las que volví a coincidir en Sevilla y que han ido apoyándome todo este tiempo: a Cristina Ruiz, por nuestros reencuentros, charlas y apoyo logístico; a Elena Caballero, por escucharme, apoyarme y aconsejarme siempre, y por sus ilustraciones tan bonitas que adornan esta tesis; y, a Azahara, gracias por las estupendas charlas, por la complicidad, las palmeras de chocolate anti-crisis y por tu sonrisa ¡¡Ojala te pueda meter en mi maleta el día que me vaya de esta ciudad!!

En estas palabras de agradecimiento no puedo olvidar hacer mención especial al rincón más grande de Andalucía, mi ciudad, Córdoba. Haber llegado hasta aquí es fruto de todo lo que comenzó por mi tierra, de los pasos que fui dando y de las personas con las que he tenido la suerte de ir encontrándome. Por que 11 somos 1 más agregados.

Y, termino con el más profundo de los agradecimientos, para las personas más cercanas genéticamente. A mis hermanos, Elizabeth y Juanko, por estar ahí y por sacarme siempre una sonrisa. Y, a mis padres, Antonia y Juan Carlos, porque sin ellos nunca se hubiera escrito esta tesis, porque nunca se han cansado de apoyarme en todo momento, de entenderme (aunque les costara comprender), de darme y facilitarme todos los recursos y capacidades que han tenido a su alcance, y por un largo etcétera. Simplemente, muchas gracias por todo. Os quiero.

Financiación

Esta tesis se ha financiado gracias a los proyectos “Estudio del seguimiento delos trabajos de control de especies exóticas invasoras en Andalucía” de EGMASA del Departamento de Medio Ambiente de la Junta de Andalucía, Consolider- Ingenio MONTES (CSD2008-00040), COMPROPIN (AGL2010-18724) y proyecto Intramural CSIC (200840I153) del Ministerio de Ciencia e Innovación.

18 Chapter 1

General Introduction

19 Chapter 1

“Como las especies de un mismo género tienen por lo común mucha semejanza en costumbres y constitución y siempre en estructura, la lucha, si entran en mutua competencia será, en general, más rigurosa entre ellas que entre especies de géneros distintos”

El origen de las especies - Charles Darwin

20 General Introduction

Biological invasions

In the last 200 years, expanding transport and intensification of commerce have influenced changes in species distributions worldwide. Many species have been introduced from one region to another where some have established and became invasive (di Castri 1989). Nowadays, biological invasions are recognized as one of the major threats to biodiversity conservation (Humphries et al. 1991; Vitousek et al. 1997). Although the earliest studies that mention biological invasions date back to mid 19th century, by De Candolle (1855) and Darwin (1859), it was not until the mid 20th century that Charles Elton (1958) established invasion ecology as a new discipline.

Alien species are those species that are introduced in a new range outside of their native origin, and where their presence has been intentionally or unintentionally caused by humans (Pyšek et al. 2004, see Box 1.1 for more definitions in the context of alien plants). Globally, the number of alien species is very large, where for example, about 10000 species in Europe (DAISIE 2009) and 50000 species in the United States (Pimentel et al. 2005) are considered to be alien.

Focusing on plants, invasive species are those alien plants that after being introduced have established and expanded into natural areas in a fast period of time (Pyšek et al. 2004). Many invasive plants cause economic and/or ecological impacts in the new range, and the adverse consequences of biological invasions vary a great deal (Vilà et al. 2010). Plant invaders can cause significant changes in biodiversity. They might displace native species by competition or by altering ecosystem properties (Lodge 1993; Levine et al. 2003) such as alterations of fire cycles (D’Antonio 2000) or modifications of the soil nitrogen cycling (Castro-Díez et al. 2009). For example, Rossiter et al. (2003) showed how the alien Andropogon gayanus (gamba grass) invading northern savannas of Australia have increased the fuel load altering the grass-fire cycle. Castro-Díez et al. (2009), which have compared two alien and two native trees in relation to litter decomposition, observe that litter decomposition was faster below the exotic than the native canopies

21 Chapter 1

Box 1.1. Standardized terminology for alien plants based on Pyšek et al. (2004) and Hulme (2011) Native plants: plant taxa that have originated in a given area without human involvement or that have arrived there without intentional or unintentional intervention of humans from an area in which they are native. Alien or exotic plants: plant taxa in a given area outside its natural past or present range whose presence there is due to intentional or unintentional human involvement, or which has arrived there without the help of people from an area in which they are alien. Introduced plants: plants outside its past or present natural range as a consequence of direct or indirect movement by humans. Casual alien plants: alien plants that may flourish and even reproduce occasionally outside cultivation in an area, but that eventually die out because they do not form self-replacing populations. Naturalized or established plants: alien plants that sustain self-replacing populations for at least 10 years without direct human intervention by recruitment from seed or ramets (tillers, tubers, bulbs, fragments, etc.). Invasive plants: a subset of naturalized plants that produce reproductive offspring, often in very large numbers, at considerable distances from the parent plants, and thus have the potential to spread over a large area. Usually causing significant harm to biological diversity, ecosystem functioning, socio-economic values and/or human health in the invaded region. The approximate scales of spread, and therefore to consider an alien species as invasive, would be more than 100 m in less than 50 years for taxa spreading by seeds, or more than 6 m in 3 years for taxa spreading by roots, rhizomes, stolons or creeping stems.

Nevertheless, not all alien plant species that arrive to a new range become invasive and cause impact. From 5789 alien terrestrial plants in Europe, around 5% are reported to cause ecological and/or economic impacts (Vilà et al. 2010). As a generalization of the proportion of introduced species that become invasive, Williamson (1993) proposes the tens rule. It predicts that 10% of imported species escape to become casual, 10% of casuals become naturalized and 10% of naturalized species become invasive (sensu Pyšek et al. 2004). However, this rule has to be taken with caution since the stages of invasion change spatially and temporally (Richardson & Pyšek 2006).

Alien species are subjected to different barriers from transport to spread in the introduced range (Fig. 1.1). Those barriers are: first, the geographical barriers from their native range to the introduced range such as intercontinental and/or

22 General Introduction

infra-continental; second, environmental barriers in the introduced range such as abiotic and biotic ecological factors; third, reproduction barriers mainly for those introduced species with highly specialized pollination or seed-dispersal systems which are less likely to be present in the new range (Richardson 2006); fourth, local and regional dispersal barriers; fifth, environmental barriers in human-modified or alien dominated vegetation; and, finally, environmental barriers in the recipient natural or seminatural vegetation (from Richardson et al. 2000).

)

)

HABITATS

(LOCAL)

IC AL PH (NATURAL DISTURBED HABITATS DISTURBED ( DISPERS GEOGRA REPRODUCTIVE ENVIRONMENTAL ENVIRONMENTAL ENVIRONMENTAL

Status of alien taxa casual naturalized

invasive

Fig. 1.1. Schematic representation of different barriers which alien plant species are subjected to from seed transport to the spread in the new introduced range. The status of alien species is represented by arrows on the bottom of the scheme, depending on their success to overcome these barriers (from Richardson & Pyšek 2006). See Box 1.1 for definitions of invasion status.

In general, once an alien plant species arrive to the introduced range, invasion success depends on propagule pressure (i.e., the number of individuals introduced and the number of introduction attempts) (Williamson 1996; Lockwood et al. 2005; Colautti et al. 2006), invasiveness (i.e., the intrinsic potential of an alien species to invade) and invasibility (i.e., the susceptibility of the recipient community to be invaded) (Lonsdale 1999), aspects that have tended to be studied isolated from each other.

23 Chapter 1 Propagule pressure

Differences in invasion success among regions and alien plant species might be related to time since introduction and the amount and frequency of propagules introduced. For many forestry tree species, propagule pressure refers to the spatial scale of the plantation (Bucharova & van Kleunen 2009; Simberloff et al. 2010). Because of their economic importance, many conifer species have been planted for forestry and ornamental purposes in areas outside of their natural range (Richardson et al. 1994), and some of these have escaped beyond the plantation area (Richardson 1998). For example, in the Southern Hemisphere, several alien conifers have been planted widely and very frequently (Richardson & Higgins 1998), therefore possibly the invasion success of many conifers in the Southern Hemisphere compared to the lower success in the Northern Hemisphere depends on differences on propagule pressure.

Determinants of invasiveness

Many studies have tried to find common life-history traits that determine a high invasiveness of plant species (reviewed in Pyšek & Richardson 2007). The most cited ones correspond to those comparing invasive and non-invasive conifer species (Rejmánek & Richardson 1996; Grotkopp et al. 2002; Richardson & Rejmánek 2004).

Pines (Pinus spp.) constitute a classic model in the study of traits associated with invasiveness. The genus Pinus contains approximately 111 species (Price et al. 1998), which mainly occur as native species in the Northern Hemisphere (Richardson et al. 1994) where forestry plantation has been the main pathway of introduction. Today, 12% pine species have been reported as invasive in at least one country (Richardson & Rejmánek 2004). Pines can grow well in a wide range of environmental conditions, are wind pollinated, well dispersed by seeds and isolated individuals can transform into colonies by selfing (Richardson & Higgins 1998). All these characteristics predispose pines to rapid migration and demographic increase. Rejmánek & Richardson (1996) proposed a discriminate function derived from attributes of invasive and non-invasive Pinus species as predictors of invasiveness in pines. According to this function, there are three main traits associated with invasiveness: small seed mass, short juvenile period and short interval between large seed crops. Fast relative early growth rate has also been considered important to predict Pinus invasion potential (Grotkopp et al. 2002). These traits have been extrapolated to other conifer species as a baseline for their invasiveness (Richardson & Rejmánek 2004). The seed size aspect of pine species will be investigated in this thesis.

24 General Introduction

Determinants of invasibility

As above mentioned, the arrival of alien species to a new range is controlled by different environmental and biotic barriers against their establishment (Fig. 1.1). Most of these barriers act at early-life cycle stages that are critical in the invasion process (Carey 1996), encompassing the stages from seed dispersal to seedling survival (Fig. 1.2). Therefore, survival at these stages is a key factor for the fitness of alien plants and an important determinant for invasion success.

Seedling Seed survival and EXPANSION Seed rain germination growth STAGE

Seed removers Pathogens Herbivores

Environmental conditions

(e.g., temperature, humidity, etc)

Fig. 1.2. Simplification scheme of different early-life cycle stages for an alien plant species. Horizontal and dotted arrows show processes and factors, respectively, that might influence each stage.

Concerning abiotic barriers, Rejmánek (1989) suggests that plant communities in mesic environments are more invasible than communities in extreme terrestrial environments. In this vein, germination and seedling survival of many alien species would be limited by environmental stress (e.g., high temperatures and drought in xeric environments). Overall, it is generally assumed that more heavily disturbed communities are more prone to invasion (Hobbs & Huenneke 1992; Londsale 1999) hence, open and disturbed communities are invaded more often than undisturbed forests. For example, Rejmánek (1996) observed how in undisturbed tropical forest the number of alien plant species is very low and do not spread beyond canopy gaps.

Regarding biotic barriers, differences in invasion success among regions might also depend on differences in the interaction with resident species in the new range 25 Chapter 1 (Richardson & Higgins 1998). There are two main, and sometimes contradictory hypotheses on the role of biotic interactions on invasions: the Natural Enemies Hypothesis and the Biotic Resistance Hypothesis. On one hand, the Natural Enemies Hypothesis predicts that alien plant species are successful invaders in the introduced range due to a lack of natural specialist enemies, which were left behind in their native range (Maron & Vilà 2001). Several studies have shown how introduced plant populations in a new area are less damaged by native phytophages and pathogens than populations in the native range (DeWalt et al. 2004; Vilà et al. 2004; Rogers & Siemann 2005). The NEH also contends that aliens undergo less pressure from natural enemies than do related native plants coexisting in the area. Nevertheless, since generalist herbivores are common in most communities, it is unclear whether the level of damage inflicted by them against alien plant species is lower than coexisting native plants (Chun et al. 2010).

On the other hand, the Biotic Resistance Hypothesis states that native generalist herbivores are important barriers to invasion in the recipient community (Elton 1958). On average, native herbivores reduce the performance of alien plants by a third from seed dispersal to seedling survival (Maron & Vilà 2001). In this direction, post-dispersal seed predation might be an important filter by decreasing the number of seeds available for alien plant establishment (Vilà & Gimeno 2003). Although consumption of seeds by seed predators is mainly influenced by the abundance of seeds and herbivores, it also depends on seed traits of alien species (Hamilton et al. 2005). In this thesis we will explore if there is a relationship between seed size and seed predation in native and alien conifers.

Furthermore, the impact of native herbivores on alien plants is determined by the interaction between their ability to recognize the alien plant as a food resource (Carpenter & Cappuccino 2005), the effectiveness of alien plant defenses against herbivores in the new range (Stowe et al. 2000) and the ability of herbivores to overcome defenses of the alien plant (Rausher 2001). Therefore, the level of attack by herbivores can vary considerably between plant species, depending on their defensive strategies (Agrawal et al. 2005). In this thesis, this interaction will be assessed in several congener pine species.

26 General Introduction

Conifer species as the study system for this thesis

In invasion ecology, conifer species are a target species due to availability of information about planting history and detailed data on species traits (Rejmánek & Richardson 1996; Grotkopp et al. 2002; Richardson & Rejmánek 2004; Richardson 2006). More than 30 alien conifer species have been reported as invasive in at least one country (Richardson & Rejmánek 2004). Several studies have reported ecological impacts of alien conifers in the introduced range. For instance, in Australia, comparisons between alien pine plantations and native eucalypt forests reveal that the former habitat is less useful for various local wildlife species (Richardson et al. 1994), while the presence of pine species, such as Pinus nigra or P. ponderosa, on infertile montane soils in New Zealand increases soil phosphorus levels under the pines (Davis & Lang 1991).

Most of the concern regarding alien conifer invasions has been voiced in the Southern Hemisphere (reviewed in Richardson et al. 1994; Simberloff et al. 2010). In South Africa, fynbos, a shrubland vegetation type with a low representation of native trees, is heavily invaded by pines (Richardson & Higgins 1998), which cause local extinction of many indigenous plants (Richardson & Van Wilgen 1986). In Argentina, Pseudotsuga menziesii seems to have an imminent invasion in open Austrocedrus forests (Simberloff et al. 2002). In sharp contrast, in the Northern Hemisphere many alien conifers were introduced and widely planted in the past century, but few conifer invasions have been reported (reviewed by Mortenson & Mack 2006). In USA, alien conifers have failed to persist and spread beyond their plantation perimeters (Mortenson & Mack 2006). The same trends might be occurring in Europe, where Adamowski (2004) found no evidence of invasion by 49 alien conifer taxa planted in the Polish–Belarusian border. Similarly, in Spain, Montero et al. (2005), who examined 34 alien conifers introduced as experimental plantations in several regions of Spain during the 60s, revealed low establishment success.

In short, in the last two decades, along with forests being attributed a more multifunctional value, there have been conservation concerns about the invasive spread of conifers and the resulting economic and ecological impacts (Richardson et al. 1994; Binimelis et al. 2007). However, in Europe, in spite of the fact that several alien conifers have been widely planted, little is known about their current invasion status and the barriers to their invasion. Further research is needed to determine reasons for the failure of alien conifers to spread in Europe (Rejmánek & Richardson 2004). This thesis is an attempt to clear this gap.

27 Chapter 1

Objectives

In invasion ecology, predictions about which species have the highest potential to invade and which habitats might be the most invaded are essential to manage alien species (Byers et al. 2002). As mentioned in the previous sections, ecological factors affecting early-life cycle stages are crucial to understand the establishment success/ failure of alien plants. In spite of the fact that conifers have been a study model of invasiveness research and much seed biology has been involved in this analysis, relatively few studies have addressed the ecological barriers from seed to seedling and whether these barriers are site specific.

The general aims of this thesis are:

I) To provide information about the current invasion status of alien conifers in Europe.

II) To screen which early stage of the plant life-cycle limits the establishment of alien conifers.

III) To compare differences in natural enemy resistance among native and alien conifer species.

28 General Introduction

Thesis Outline

In Chapter 2, we conducted a literature review to assess the scientific evidence of alien conifer invasion in Europe. This review seeks to identify the most widely distributed alien conifers in Europe, their present invasion status, determinants of invasion and impacts.

In Chapter 3, we performed cafeteria seed removal field experiments to compare 16 pine species (5 native and 11 alien species to Spain) differing in seed mass. The study that took place in disturbed Mediterranean shrublands in Aznarcóllar (Sevilla, Spain) addressed whether seed mass influenced the likelihood of differential seed predation between native and alien pine species.

In Chapter 4, we examined differences in the establishment success of the alien conifer Douglas fir (Pseudotsuga menziesii) and the native Silver fir (Abies alba) in Montseny Natural Park (Barcelona, Spain). We explored early stages of the life-cycle (seed survival, seed germination, seedling survival and seedling growth) by transplant experiments in three mountainous habitats adjacent to Douglas-fir plantations: beech forests, holm-oak forests and heathlands.

In Chapter 5, we tested, in greenhouse conditions, if there were differences in herbivory resistance between 4 native and 3 alien pine species when exposed to two European native generalist herbivores. Within the context of the predictions of the Natural Enemies Hypothesis, we investigated constitutional and induced chemical defenses of seedlings.

And, in Chapter 6, the main results from the preceding four chapters are synthesized and discussed. Finally, conclusions of this thesis are exposed.

29 From seed to seedling: Exploring the early life-cycle stages in alien conifer species

30 Introducción General

31 32 Capítulo 1

Invasiones biológicas

En los últimos 200 años, como consecuencia de la intensificación del comercio se están produciendo cambios en la distribución de especies en todo el mundo. Muchas especies han sido introducidas desde su distribución natural a otra nueva región, algunas de éstas se han establecido y otras han llegado a convertirse en invasoras (di Castri 1989). Hoy en día, las invasiones biológicas están consideradas como una de las principales amenazas en la conservación de la biodiversidad (Humphries et al. 1991; Vitousek et al. 1997). Aunque los primeros estudios que mencionan las invasiones biológicas se remontan a mediados del siglo XIX (De Candolle 1855; Darwin 1859), no será hasta mediados del siglo XX cuando Charles Elton (1958) establezca la ecología de las invasiones como una nueva disciplina.

Las especies exóticas son aquellas especies que son introducidas en un nuevo lugar fuera de su lugar natural de distribución. Su presencia en la nueva área de distribución puede ser debida a procesos intencionados o no intencionados por el hombre (Pyšek et al. 2004, ver Caja 1.1 para más definiciones en el contexto de plantas exóticas). Globalmente, el número de especies consideradas como exóticas es muy grande. Por ejemplo, alrededor de 10000 especies son consideradas exóticas en Europa (DAISIE 2009) mientras que 50000 especies lo son en Estados Unidos (Pimentel et al. 2005).

Considerando las especies vegetales, se denomina planta invasora a aquella planta que una vez ha sido introducida en un nuevo lugar, ha sido capaz de establecerse y expandirse en áreas naturales en un período corto de tiempo (Pyšek et al. 2004). Muchas plantas invasoras causan impactos económicos y/o ecológicos en el nuevo lugar de introducción, variando enormemente las consecuencias adversas de las plantas invasoras (Vilà et al. 2010). Las plantas invasoras pueden causar cambios significativos en la biodiversidad pudiendo desplazar a las especies nativas bien por competencia o bien por alteración de las propiedades del ecosistema receptor (Lodge 1993; Levine et al. 2003), tales como alterando el ciclo natural del fuego en determinados ecosistemas (D’Antonio 2000) o modificando el ciclo del nitrógeno en el suelo (Castro-Díez et al. 2009). Por ejemplo, Rossiter et al. (2003) muestran como la invasión de la especie exótica Andropogon gayanus en las sabanas del norte de Australia incrementa la carga de combustible alterando el régimen de incendios de los pastos en dicha área. Castro-Díez et al. (2009) que comparan tasas de descomposición de las hojas bajo la cubierta de dos especies arbóreas exóticas

33 Introducción General y dos especies nativas, observaron que la descomposición fue más rápida bajo las especies exóticas que bajo de las nativas. Caja 1.1. Terminología estandarizada para plantas exóticas de acuerdo a Pyšek et al. (2004) y Hulme (2011) Planta nativa: taxón que se ha originado en una zona determinada sin intervención humana o que ha llegado desde su área de distribución natural sin intervención deliberada o involuntaria del ser humano. Planta exótica: taxón localizado en una zona determinada fuera de su área de distribución natural pasada o presente y cuya presencia se debe a la intervención deliberada o involuntaria del ser humano, o que ha llegado desde un área donde es exótica sin la ayuda de personas. Planta introducida: aquella planta fuera de su área de distribución natural pasada o presente y que ha llegado allí como consecuencia del movimiento directo o indirecto del ser humano. Planta subespontánea: planta exótica que puede florecer e incluso reproducirse ocasionalmente fuera de su área de cultivo, pero que finalmente se extingue por que no es capaz de formar poblaciones autosuficientes. Planta naturalizada o establecida: planta exótica capaz de mantener poblaciones autosuficientes por al menos 10 años sin la intervención directa del hombre como consecuencia del reclutamiento de semillas o partes de la planta (tallos, tubérculos, bulbos, fragmentos, etc.). Planta invasora: planta naturalizada que produce descendencia reproductiva, a menudo en grandes cantidades, a distancias considerables de la planta madre, y por lo tanto tienen el potencial de propagarse en una amplia área. Generalmente causan daños significativos sobre la diversidad biológica, funcionamiento del ecosistema, sobre los valores socio-económicos y/o sobre la salud humana en la región de invasión. Las escalas aproximativas de propagación, y por tanto de considerar una especie exótica como invasora, sería de más de 100 m en menos de 50 años para los taxones cuya forma de propagación sea por semillas, o de más de 6 m en 3 años para los taxones cuya propagación sea por raíces, rizomas, estolones o tallos rastreros.

Sin embargo, no todas las especies exóticas que llegan a una nueva área de distribución se convierten en invasoras y causan impacto. De 5789 especies exóticas de plantas terrestres que existen en Europa, poco más de un 5% han sido descritas como causantes de algún tipo de impacto ecológico y/o económico, respectivamente (Vilà et al. 2010). Como una generalización a la proporción de especies introducidas que se convierten en invasoras, Williamson (1993) propone la regla del diez. Esta regla predice que 10 % de las especies importadas escapa para convertirse en especies subespontáneas, 10% de las especies subespontáneas llegan a naturalizarse y 10% de las naturalizadas llegan a convertirse en invasoras (sensu Pyšek et al. 2004). Sin embargo, esta regla tiene que considerarse con precaución puesto que las etapas del proceso de invasión pueden variar tanto espacial como temporalmente (Richardson & Pyšek 2006). 34 Capítulo 1

Desde su transporte hasta la llegada a una nueva área de distribución, las especies exóticas están sometidas a diferentes barreras (Fig. 1.1). Estas barreras serían, primeramente, las barreras geográficas con las que se encuentran las especies desde su área de distribución natural hasta su nueva área de distribución, siendo tanto barreras intercontinentales y/o intracontinentales. Segundo, barreras ambientales en la nueva área de distribución debido tanto a factores ecológicos bióticos como abióticos. Tercero, barreras reproductivas que pueden encontrarse aquellas especies con sistemas de polinización o dispersión altamente especializados los cuales son menos probables de encontrar en la nueva área (Richardson 2006). Cuarto, barreras locales y regionales contra su dispersión. Quinto, barreras ambientales en zonas de vegetación perturbada o dominada por otras especies exóticas. Y, finalmente, barreras ambientales debidas a la vegetación natural o seminatural nativa de la zona donde se introducen (según Richardson et al. 2000).

URBADO)

AT NATURAL) AT T I AT PERT AT T I DISPERSION GEOGRAFICA AMBIENTAL(LOCAL) REPRODUCTIVA AMBIENTAL(HAB AMBIENTAL(HAB

Estado exótica

subespontánea naturalizada

invasora Fig. 1.1 Representación esquemática de las diferentes barreras con las que una especie exótica se enfrenta desde su transporte hasta su expansión en la nueva área de distribución. El estado de la especie exótica es representado por flechas en la parte baja del esquema, dependiendo de su éxito para superar dichas barreras (según Richardson & Pyšek 2006). Ir a Caja 1.1 para ver definiciones sobre el estado de invasión.

35 Introducción General

En general, una vez que las especies exóticas llegan a la nueva área de distribución, el éxito de la invasión depende tanto de la presión de propágulos, es decir, del número de individuos introducidos y del número de intentos de introducción (Williamson 1996; Lockwood et al. 2005; Colautti et al. 2006); como del potencial intrínseco de invasión de la especie exótica (invasiveness); así como de la susceptibilidad del ecosistema receptor ha ser invadido (invasibility) (Lonsdale 1999), éstos aspectos han sido frecuentemente estudiados de forma independiente.

Presión de propágulos

Las diferencias en el éxito de invasión entre regiones y entre especies exóticas puede estar relacionadas con el tiempo desde que dichas especies fueron introducidas y la cantidad y frecuencia de propágulos con la que se introducen. Para muchas especies arbóreas forestales, la presión de propágulos se refiere al tamaño de la plantación (Bucharova & van Kleunen 2009; Simberloff et al. 2010). Por razones económicas, muchas especies de coníferas han sido plantadas por motivo forestal u ornamental en áreas fuera de su área natural de distribución (Richardson et al. 1994), algunas de ellas han escapado más allá de la plantación (Richardson 1998). Por ejemplo, en el Hemisferio Sur, muchas coníferas exóticas han sido amplia y frecuentemente plantadas (Richardson & Higgins 1998). Posiblemente el éxito de invasión de muchas coníferas en el Hemisferio Sur comparado con el bajo éxito detectado en el Hemisferio Norte pueda ser debida a diferencias en la presión de propágulos.

Determinantes del potencial invasor

Muchos estudios han intentado encontrar características comunes que determinen el gran potencial de invasión de las especies vegetales (revisado en Pyšek & Richardson 2007). Probablemente, uno de los trabajos más importantes es aquel que compara especies de coníferas invasoras y no-invasoras (Rejmánek & Richardson 1996; Grotkopp et al. 2002; Richardson & Rejmánek 2004).

Las especies de pino (Pinus spp.) constituyen un modelo clásico en el estudio de las características asociadas al potencial invasor. El género Pinus contiene aproximadamente 111 especies (Price et al. 1998), las cuales son nativas principalmente del Hemisferio Norte (Richardson et al. 1994). Hoy en día, un 12% del total de las especies de pino han sido identificadas como invasoras en al menos un país (Richardson & Rejmánek 2004). Los pinos pueden crecer bien en un amplio rango de condiciones

36 Capítulo 1 ambientales, se dispersan bien por semillas e individuos aislados pueden dar lugar a nuevas colonias por autofecundación (Richardson & Higgins 1998). Todos estos caracteres predisponen a los pinos a una rápida colonización y un rápido incremento demográfico de sus poblaciones. Rejmánek & Richardson (1996) propusieron una función discriminatoria creada a partir de caracteres de las especies de pino invasoras y no-invasoras que permite predecir el potencial invasor de este género. De acuerdo a esta función, tres son los principales caracteres asociados a dicho potencial invasor: un pequeño tamaño de semilla, un período corto de tiempo de los individuos como juveniles y un intervalo corto de tiempo entre dos períodos consecutivos de alta producción de semillas. Una tasa de crecimiento rápido ha sido también considerado un carácter importante para predecir el potencial de invasión en especies de pino (Grotkopp et al. 2002). Posteriormente, estos caracteres han sido extrapolados a otras especies de coníferas como línea base para determinar su potencial de invasión (Richardson & Rejmánek 2004). De estos caracteres, el tamaño de semilla en las especies de pino será investigado en esta tesis.

Determinantes de la invasibilidad

Como se mencionó anteriormente, la llegada de una especie exótica a una nueva área de distribución está controlada tanto por barreras ambientales como por barreras bióticas que impiden su establecimiento (Fig. 1.1). Muchas de estas barreras actúan en los estadios iniciales del ciclo de vida de las plantas, siendo esta fase crítica en el proceso de invasión (Carey 1996), abarcando desde la dispersión de semillas hasta la supervivencia de plántulas (Fig. 1.2). Por tanto, la supervivencia en estos estadios es un factor clave para el establecimiento de las plantas exóticas siendo un factor determinante en el éxito de la invasión.

37 Introducción General

Supervivencia Germinación Lluvia de y crecimiento ETAPA DE semillas de semillas de plántulas EXPANSION

Depredadores Patógenos Herbívoros

Condiciones ambientales

(Ej., temperatura, humedad, etc) Fig. 1.2. Esquema simplificado de los estadios iniciales del ciclo de vida en las especies vegetales. Flechas horizontales y discontinuas indican procesos y factores, respectivamente, que pueden influir en cada etapa.

De acuerdo a las barreras abióticas, Rejmánek (1989) sugiere que las comunidades de plantas en climas templados son más susceptibles a ser invadidas que aquellas comunidades en climas extremos. En este sentido, la germinación y supervivencia de las plántulas de muchas especies exóticas estaría limitada por el estrés ambiental (Ej., alta temperatura y sequía en ambientes xéricos). En general, las comunidades más fuertemente perturbadas son las que se consideran más propensas a ser invadidas (Hobbs & Huenneke 1992; Londsale 1999) de ahí que, comunidades de baja cobertura vegetal o perturbadas son más invadidas que los bosques no perturbados. Por ejemplo, Rejmánek (1996) observa como en bosque tropical no perturbado el número de las especies exóticas era muy bajo y que las mismas no se expandían más allá de las zonas de claro o borde.

Con respecto a las barreras bióticas, las diferencias en el éxito de invasión entre regiones también puede depender de las diferencias en la interacción con las especies nativas en la nueva área de distribución (Richardson & Higgins 1998). Las principales, y a veces contradictorias, hipótesis sobre el papel de las interacciones bióticas en el proceso de invasión son: la Hipótesis de los Enemigos Naturales y la Hipótesis de la Resistencia Biótica. En un sentido, la Hipótesis de los Enemigos Naturales predice que las especies exóticas tienen éxito en la invasión en la nueva área de distribución debido a una falta de sus enemigos naturales especialistas, los cuales controlan sus poblaciones en su área de distribución natural (Maron & Vilà 2001).

38 Capítulo 1

Muchos estudios han mostrado como las poblaciones de plantas introducidas en una nueva área son menos dañadas por parte de los fitófagos y patógenos que sus poblaciones en su área de distribución nativa (DeWalt et al. 2004; Vilà et al. 2004; Rogers & Siemann 2005). Esta hipótesis también sostiene que las especies exóticas están sometidas a una menor presión por parte de los enemigos naturales que las especies nativas con las que coexisten en la nueva área de distribución. Sin embargo, debido a que los herbívoros generalistas son frecuentes en muchas comunidades, no está claro si el nivel de daño causado por dichos generalistas es siempre menor sobre las especies de plantas exóticas que sobre las nativas con las que coexisten (Chun et al. 2010).

Por otro lado, la Hipótesis de la Resistencia Biótica afirma que los herbívoros nativos generalistas son barreras importantes contra la invasión de las especies exóticas en la comunidad receptora (Elton 1958). En general, los herbívoros nativos reducen en un tercio el rendimiento de las plantas exóticas desde la dispersión de semillas hasta la supervivencia de las plántulas (Maron & Vilà 2001). En este sentido, la depredación de semillas después de la dispersión puede ser un filtro importante de disminución del número de semillas disponibles para el establecimiento de la especie exótica (Vilà & Gimeno 2003). Aunque, el consumo de semillas por los depredadores está principalmente influenciado por la abundancia tanto de las semillas como de los depredadores, también depende de las características de las semillas (Hamilton et al. 2005). En esta tesis exploraremos si existe relación entre el tamaño de semilla y la depredación entre especies de coníferas nativas y exóticas.

Por tanto, el impacto de los herbívoros nativos sobre las plantas exóticas vendrá determinado por la interacción entre la habilidad de los herbívoros de reconocer a la planta exótica como recurso alimenticio (Carpenter & Cappuccino 2005), la efectividad de defensa de la planta exótica contra los herbívoros en la nueva área de distribución (Stowe et al. 2000) y la habilidad de los herbívoros de superar las defensas producidas por la planta exótica (Rausher 2001). Por lo que, el nivel de ataque de los herbívoros puede variar considerablemente entre especies vegetales, dependiendo de sus estrategias de defensa (Agrawal et al. 2005). En esta tesis, la interacción entre estos aspectos será evaluado comparando varias especies de coníferas pertenecientes a un mismo género.

39 Introducción General Las coníferas como sistema de estudio para esta tesis

En la ecología de la invasión, las coníferas son una especie modelo utilizada debido a la disponibilidad de información detallada que existe tanto sobre su historia de plantación como acerca de sus caracteres biológicos (Rejmánek & Richardson 1996; Grotkopp et al. 2002; Richardson & Rejmánek 2004; Richardson 2006). Más de 30 especies exóticas de coníferas han sido identificadas como invasoras en al menos un país (Richardson & Rejmánek 2004). Muchos estudios han informado como las coníferas exóticas introducidas han causado impactos ecológicos en la nueva área de distribución. Por ejemplo, en Australia se comparó plantaciones de pino exótico con bosques nativos de eucalipto y se observó que las primeras acogen menos diversidad de fauna (Richardson et al. 1994). Mientras que en Nueva Zelanda la presencia de especies de pino, tales como Pinus nigra o P. ponderosa, en suelos infértiles de montaña incrementa los niveles de fósforo en dichos suelos (Davis & Lang 1991).

La mayoría de las preocupaciones acerca de las invasiones de coníferas exóticas proceden del Hemisferio Sur (revisado en Richardson et al. 1994; Simberloff et al. 2010). En Sudáfrica, el fynbos, un tipo de vegetación arbustiva con una baja representación de especies arbóreas nativas, está altamente invadido por pinos (Richardson & Higgins 1998), los cuales causan extinciones locales de muchas plantas nativas (Ej., Richardson & van Wilgen 1986). En Argentina, Pseudotsuga menziesii parece predecir una inminente invasión en áreas abiertas de bosque de Austrocedrus (Simberloff et al. 2002). En contraste, en el Hemisferio Norte muchas coníferas exóticas fueron introducidas y ampliamente plantadas durante el siglo pasado pero han sido detectados pocos casos de invasión (revisado por Mortenson & Mack 2006). A pesar de la heterogeneidad ambiental del Hemisferio Norte, las coníferas exóticas parecen tener menos éxito de invasión que en el Hemisferio Sur. En Estados Unidos, las coníferas exóticas se establecen y expanden poco más allá de su área de plantación (Mortenson & Mack 2006). El mismo patrón podría estar ocurriendo en Europa, Adamowski (2004) tampoco encontró evidencia de invasión de ninguna de las 49 coníferas exóticas plantadas en la frontera Polaco-Bielorusa. Similarmente, en España, Montero et al. (2005), quienes examinaron el estado de 34 coníferas exóticas introducidas como plantaciones experimentales en los años 60 en distintas regiones del territorio español, encuentran un bajo éxito de establecimiento de las mismas.

40 Capítulo 1

En resumen, en las últimas dos décadas, junto con el valor multifuncional que se les atribuye a los bosques, existe la preocupación sobre la invasión de coníferas y sus impactos económicos y ecológicos (Richardson et al. 1994; Binimelis et al. 2007). Sin embargo, en Europa, a pesar de que varias especies de coníferas exóticas han sido ampliamente plantadas, poco es sabido sobre el estado de invasión actual de las mismas y sobre las posibles barreras que están actuando contra su invasión. En este sentido es necesario determinar las razones del fallo por las cuales las coníferas exóticas no se expanden con éxito en Europa (Rejmánek & Richardson 2004). Esta tesis trata de contribuir a este hueco del conocimiento.

41 Introducción General

Objetivos

En ecología de las invasiones, las predicciones acerca de qué especies tienen mayor potencial invasor y qué hábitats pueden ser los más invadidos son esenciales para la gestión de las plantas exóticas. (Byers et al. 2002). Como se menciona en secciones anteriores, el estudio de los factores ecológicos que afectan los estadios iniciales del ciclo de vida es crucial para entender el éxito/fracaso de las plantas exóticas. A pesar que las coníferas han sido un modelo muy estudiado en lo referente a su capacidad de invasión, y mucha biología de la semilla ha sido considerada en este análisis, hay relativamente pocos estudios que hayan abordado las barreras ecológicas que influyen desde la fase de semilla hasta la de plántula, y si esas barreras son específicas de las localidades de introducción.

Los objetivos generales de esta tesis son:

I) Proporcionar información sobre el estado actual de invasión por coníferas exóticas en Europa.

II) Identificar qué etapa del ciclo de vida inicial de las plantas limitael establecimiento de las coníferas exóticas.

III) Comparar diferencias en la resistencia a los enemigos naturales entre especies de coníferas nativas y exóticas.

42 Capítulo 1

Esquema de la tesis

En el Capítulo 2, hemos realizado una revisión bibliográfica para evaluar la evidencia científica de la invasión de coníferas exóticas en Europa. Esta revisión tuvo por objetivo identificar las coníferas exóticas de mayor distribución en Europa, su estado actual de invasión, los factores determinantes de la invasión y los impactos de las mismas.

En el Capítulo 3, se realizaron experimentos de depredación de semillas en el campo para comparar 16 especies de pino (5 nativas y 11 especies exóticas en España) las cuales difieren en el tamaño de la semilla. El estudio se llevó a cabo en áreas perturbadas de matorral Mediterráneo en Aznalcóllar (Sevilla, España) donde se trató de determinar si el tamaño de la semilla influye en la probabilidad de depredación de semillas entre especies de pino exóticas y nativas.

En el Capítulo 4, se examinaron diferencias en el éxito de establecimiento entre una especie de conífera exótica, el abeto de Douglas (Pseudotsuga menziesii) y una conífera nativa, el abeto común (Abies alba) en el Parque Natural del Montseny (Barcelona, España). Se exploraron las primeras etapas del ciclo de vida de las especies (supervivencia y germinación de las semillas, así como supervivencia y crecimiento de plántulas) a través de experimentos llevados a cabo en tres tipos de hábitats cercanos a las plantaciones de abeto de Douglas: hayedo, encinar y brezal.

En el Capítulo 5, se testaron, en condiciones de invernadero, si existen diferencias en la resistencia a herbívoros comparando 4 especies nativas y 3 especies exóticas de pino exponiéndolas a dos herbívoros generalistas nativos de Europa. Dentro del contexto de las predicciones de la Hipótesis de los Enemigos Naturales, se investigaron las defensas químicas constitutivas e inducidas en plántulas.

Y, en el Capítulo 6, los resultados principales de los cuatro capítulos precedentes son sintetizados y discutidos. Finalmente, se exponen las conclusiones de esta tesis.

43 44 Chapter 2

Little evidence of invasion

by alien conifers

in Europe1

1Carrillo-Gavilán MA, Vilà M (2010) Little evidence of invasion by alien conifers in Europe. Diversity and Distributions 16:203–213.

45 Chapter 2

Abstract

Conifers are invasive species in many parts of the world. There are many introduced conifers in Europe, but their status as alien species is poorly documented. We conducted a comprehensive literature review to ascertain the extent to which alien conifers can be considered invasive. We reviewed the historical record of alien conifer invasion in Europe (i.e. species with a native range outside the continental boundaries of Europe) by screening the DAISIE database and the ISI Web of Science. According to DAISIE, there are 54 alien conifer species in Europe. Pseudotsuga menziesii is the species recorded as naturalized in the most countries (12) and the UK is the country with the most naturalized species (12). Thirty-seven of these conifers have been studied, to some extent, in a total of 131 papers (212 records). Nevertheless, only a few papers have investigated aspects related to biological invasions. In fact, the species are not referred to as alien by the authors in more than half of the papers (66%). Twenty-five percent of the papers have investigated plant traits, 46% are about biotic and abiotic factors influencing tree performance and 29% deal with ecological and economic impacts. Most papers are related to entomology, dealing with natural enemies affecting the alien conifers. Scientists have not yet perceived alien conifers in Europe as problematic species. Moreover, the low introduction effort, long lag-time since plantation and phylogenetic closeness between alien and native conifers are possible reasons for their low expansion in Europe to date. From a management point of view, careful observations of sites with alien conifers is necessary to watch for new invasions. From a scientific perspective, thorough analyses of the extent that introduction, ratesof naturalization and biogeographical differences influence invasive spread between the two hemispheres will prove timely.

46 Little evidence of conifer invasions

INTRODUCTION

Many conifer species have been planted for forestry and ornamental purposes around the world, and some of these have escaped outside their area of plantation (Richardson 1998). More than 30 species have been reported as invasive in at least one country (Richardson & Rejmánek 2004). In recent decades, along with forests being attributed a more multifunctional value, there have been conservation concerns about the invasive spread of such conifers and the resulting economic and ecological impacts (Richardson et al. 1994; Binimelis et al. 2007). Introduced conifers can change vegetation life-form dominance, reduce structural diversity, increase ecosystem biomass, disrupt prevailing vegetation dynamics, modify nutrient cycling and alter hydrological regimes (Richardson et al. 1994; Richardson & Higgins 1998; Levine et al. 2003).

As with any introduced plant species, the invasion success of alien conifers depends on the intrinsic potential of the species to invade (i.e., invasiveness), the susceptibility of the recipient community to be invaded (i.e., invasibility) (Lonsdale 1999), and the numbers of individuals introduced and of introduction attempts (i.e., propagule pressure) (Williamson 1996; Richardson 2006). These aspects have tended to be studied in isolation. The most successful invasive conifer species are those with a short juvenile period, a short interval between large seed crops and a small seed mass (Rejmánek & Richardson 1996; Richardson & Rejmánek 2004). Such features are correlated with a fast relative growth rate (RGR) (Grotkopp et al. 2002). Most concern regarding alien conifer invasions has been voiced in the Southern Hemisphere (Richardson & Higgins 1998), especially South Africa, where at least eight species are recorded as invasive. Several pine species, notably Pinus halepensis, P. pinaster and P. radiata are highly invasive in the region of the country with Mediterranean-type climate. Fynbos, a shrubland vegetation type with a low representation of native trees is heavily invaded by pines. In South America, there is also evidence of significant invasion by conifers driven by a high demand for wood (Simberloff et al. 2002; Pauchard et al. 2004; Simberloff et al. 2010). Conifer invasions in South America have lagged behind those in Australia, New Zealand and South Africa, where many conifers were widely planted a century earlier or more, but are increasing rapidly (Richardson et al. 2008).

Conifer invasions are far less conspicuous in the Northern Hemisphere (Richardson & Rejmánek, 2004). Although many conifers were introduced and widely planted in Europe in the past century, few species have become invasive. For example, Adamowski (2004) found no evidence of invasion

47 Chapter 2 by 49 alien conifer taxa planted in the Polish-Belarusian border. Similarly, of nine alien conifers widely planted in different parts of the USA, only two species are establishing in a few sites (Mortenson & Mack 2006).

In this paper we review the scientific evidence of alien conifer invasion in Europe (i.e. we consider species with a native range entirely outside the continental boundaries of Europe). The main questions we address are: (1) What are the principal alien conifers in Europe? (2) Where have they been introduced and what is their invasion status? (3) Which aspects of these conifers have been studied? We discuss these results as they relate to the recognized problem of conifer invasion in the Southern Hemisphere and the limited evidence of the same in the Northern Hemisphere.

IDENTITY, DISTRIBUTION AND INVASION STATUS OF ALIEN CONIFERS IN EUROPE

A preliminary analysis was based on the database Delivering Alien Invasive Species Inventories for Europe (DAISIE, 2009; http://www.europe-aliens.org/index.jsp), the largest database of alien species in the world. DAISIE was funded by the European Commission (2005-2008) to create an inventory of alien species that threaten European terrestrial, freshwater and marine environments in order to understand the environmental, economic, social and other factors involved in invasion. The DAISIE database has collated information for fungi, plants, vertebrates, invertebrates, marine and inland aquatic organisms from up to 63 countries/regions (including islands) and 39 coastal and marine areas, including regions adjacent to Europe. Presence of alien species was geographically assigned to NUTS as finer scale resolution (e.g. UTM) is not available for most species and countries.

According to DAISIE, 29 conifer species (35%) are alien in a part of Europe but native to another part, and 54 conifer species (65%) that are alien to Europe - i.e., have originated outside Europe (Lambdon et al. 2008). We focused on the last group (alien conifers or alien to Europe, hereafter). Those conifers that are alien to Europe belong to 6 families and 24 genera. Pinaceae is the most represented family, with 30 species, followed by Cupressaceae, with 13 species (Fig. 2.1a). Most of these conifers are native to North America (56%) followed by temperate Asia (26%) (Fig. 2.1b).

48 Little evidence of conifer invasions

a) a)

7% 7%7% 4%4% 2%2%

Araucariaceae

AraucariaceaeCephalotaxaceae 24% CephalotaxaceaeCupressaceae Cupressaceae PinaceaePinaceae Taxaceae TaxodiaceaeTaxaceae Taxodiaceae

56%56% b)

b)

26%

N America 6%6% AsiaN Temperate America AustralasiaAsia Temperate 6% Australasia Hybrid Hybrid 4%4% AfricaAfrica 2% S AmericaS America

56%56%

Fig. 2.1. Percentage of alien conifers in Europe by family (a) and by origin (b) according to the DAISIE database (http://www.europe-aliens.org/index.jsp).

49 Chapter 2

On average (± SE), each alien conifer in Europe has been recorded in 4 ± 0.47 countries/regions, but most species have only been recorded in one country (Fig. 2.2). Pseudotsuga menziesii is the species with the widest geographical distribution (19 countries), followed by Picea sitchensis, Pinus strobus and Platycladus orientalis (all found in 11 countries). The countries with the highest number of alien conifers are France (34), Sweden (25), Denmark (21) and UK (19).

All speciesspecies 18 Established species species 16

14

12

10

8

6 Number of species

Number of species 4

2

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 NumberNumber of countries Fig. 2.2. Frequency distribution of range sizes of alien conifers in Europe according to the DAISIE database (http://www.europe-aliens.org/index.jsp).

DAISIE classifies invasion status as established (i.e., naturalized sensu Pyšeket al. 2004), non-established and unknown. The status of alien conifers in Europe is unknown for half of the species, while only 31% of species are considered established, primarily in only one country (Fig. 2.2). The country with the largest number of established conifer species is the UK (18), and the species recorded most often as established is P. menziesii (12).

SCIENTIFIC LITERATURE ON CONIFER INVASION IN EUROPE

As a first step, we searched in the ‘‘topic’’ function of the ISI Web of Science for published papers on each alien conifer in Europe that had been identified by DAISIE. Then, we used the following criteria in the search: (the scientific OR common name of the species) AND (conifer) AND (invasion OR exotic OR expansion) AND (Europe). We accessed the ISI Web from January to March 2009.

50 Little evidence of conifer invasions

We also included some references listed by Richardson & Rejmánek (2004) who conducted a global survey of conifer invasion. A total of approximately 3500 papers were retrieved, but we selected only those studies conducted in Europe. Moreover, we did not consider publications on local or regional flora descriptions, which did not specify ecological information on particular alien conifers (e.g. Crawley et al. 1996; Sádlo et al. 2007; Lambdon 2008; Lambdon et al. 2008; Bucharova & van Kleunen 2009). A total of 131 papers met our criteria (see Annex 1). From each paper, the following information was gathered: study species, country of study, spatial scale of the study, pathway of introduction of the species and topic of the study (see specific classification below) yielding a total of 212 records. We also screened the papersto determine whether the authors presented their study within a biological invasion context.

General overview of the alien conifers

Of the 54 conifer species mentioned in DAISIE as alien to Europe, 37 have been studied in Europe (Table 2.1). One hundred and nine papers (82%) focused on a single species, 11 (8%) on two and the remaining 11 papers on more than two species simultaneously. The species coverage was very uneven with five species accounting for more than half of the records (P. sitchensis -63; P. menziesii -25; P. strobus -16; P. radiata -15; and, Larix kaempferi and Cedrus atlantica -11). Pinaceae was the family with the highest number of records analyzed (185). The species with the widest geographical distribution of studies was P. menziesii (11 countries) followed by P. strobus (8); L. kaempferi, Abies grandis and C. atlantica were studied in five different countries. Forty-seven percent of the records were performed in the UK, followed by Spain (13%), the Czech Republic (8%) and Germany (5%) (Fig. 2.3). The spatial scales of the studies reviewed were grouped as local (i.e., plot), regional (i.e., administrative regions within a country), national (i.e., several areas within a country) or international (i.e., several countries). Most records had a local focus (55%), while 12% and 32% of the records were carried out at regional and national scales, respectively.

51 Chapter 2

Table 2.1. General information on alien conifers in Europe found from an analysis of papers listed in the ISI Web of Science (date of access up to March 2009). For references of data sources, see Annex 1. Study Topic (Invasiveness/ Genus Species Origin Country1,2 References2 invasibility/ impacts) Araucariaceae Araucaria araucana S America SPA/UK 1/1/0 32-34 Cupressaceae Chamaecyparis pisifera Asia temperate UK 1/0/0 32 Cryptomerica japonica Asia temperate UK 1/0/0 32 Cupressus arizonica N America FRA/ITA/SPA 0/2/1 12-27-34 Cupressus lusitanica N America SPA 0/1/0 34 Cupressus macrocarpa N America POR/SPA/UK 1/3/1 7-32-34-101-116 Juniperus chinensis Asia temperate UK 1/0/0 32 Juniperus virginiana N America SPA/UK 2/1/0 32-34-128 Platycladus orientalis Asia temperate CZ/UK 2/0/0 32-95 x Cuprocyparis leylandii N America UK 1/0/0 32-128 Pinaceae Abies balsamea N America SK 0/0/1 41 Abies concolor N America CZ/SE/SK/SWI 1/3/1 5-41-67-80-92 Abies grandis N America CZ/FRA/GER/ 2/1/4 32-41-67-87- SK/UK 114 Abies lasiocarpa Asia temperate UK 1/0/0 32 Abies procera N America CZ/DEN/IR/UK 2/1/1 32-67-79-121 Cedrus atlantica Africa FRA/ITA/SE/ 1/8/2 15-32-34-41- SPA/UK [45-47]-72-77- 109 Larix gmelinii Asia temperate FIN/UK 1/1/0 32-93 Larix kaempferi Asia temperate FIN/GER/IR/ 1/4/6 2-8-32-55- SPA/UK 56-93-102-121- 125-127-129 Larix laricina Asia temperate FIN/UK 1/1/0 32-93 Larix x marschlinsii HYB UK 1/1/0 32-114 Picea engelmannii N America CZ/UK 2/1/0 32-67-78 Picea mariana N America CZ 1/0/0 67 Picea pungens N America AUT/CZ/RUS/ 2/1/1 32-41-67-103 UK

52 Little evidence of conifer invasions

Study Genus Species Origin Country1,2 References2 Topic Picea sitchensis N America DEN/FRA/ 0/33/30 11-14-17-18-20- IR/UK 21-23-26-[29-31]- [35-39]-[42-44]- [49-52]-55-59- 60-62-71-73-74- 76-78-[81-85]- 87-88-90-91-[97- 100]-[104-106]- 108-[110-115]-118- 120-[122-127]-130 Pinus banksiana N America CZ/SPA/UK 2/1/0 32-34-67 Pinus ponderosa N America SPA/UK 2/1/0 32-33-34 Pinus radiata N America FRA/ITA/ 3/6/6 1-3-4-13-22-25- SPA/UK 32-34-63-68- 69-72-107-116- 117-119 Pinus rigida N America SPA/UK 1/1/0 32-34 Pinus strobus N America CZ*/GER/ 5/10/1 9-10-13-24-32- ITA/NET/ 34-48-61-66-72- POL/ POR/ 75-57*-58*-67*- SPA/UK 95* Pinus wallichiana Asia temperate UK 1/0/0 32 Pinus x rotundata HYB CZ 0/1/0 107 Pseudotsuga menziesii N America BGM/CZ/ 7/9/9 6-16-19-28-32- DEN/FIN/ 41-49-53-54- FRA/GER/ [64-67]-70-71- IR/POR/ [86-88]-94-95- SPA/SWI/ 96-114-121-129- UK 130 Tsuga canadensis N America GER/UK 1/1/0 32-40 Taxaceae Sequoia sempervirens N America SPA/UK 1/1/0 32-34 Taxodiaceae Metasequoia glyptostroboides Asia temperate UK 1/0/0 32 Sequoiadendron giganteum N America SPA/UK 1/1/0 32-34 Taxodium distichum N America SE 0/1/0 89

1Country- AU: Austria; BGM: Belgium; CZ: Czech Republic; DEN: Denmark; FIN: Finland, FRA: France, GER: Germany; IR: Republic of Ireland; ITA: Italy; NET: The Netherlands; POL: Poland; POR: Portugal; RUS: Russia; SE: Serbia; SK: Slovakia; SPA: Spain; SWI: Switzerland; UK: United Kingdom. 2In bold: the species was considered naturalized; * Pinus strobus is considered invasive in the Czech Republic.

53 Chapter 2

A large proportion of the reviewed literature (29%) did not specify the pathway of introduction of the conifer species. Of the remaining publications, the main introduction pathways were via silviculture (57%), as ornamentals (11%) or a combination of both purposes (3%).

120

100

80

60

40 Number of records Numbers of records 20

0 IR SE SK CZ UK ITA FIN SWI SPA FRA AUT NET DEN RUS GER POR BGM Europe CZ/POL FRA/ITA Country SPA/FRA Fig. 2.3 Geographical distribution of records Country(n=212) for studies of alien conifers in Europe found at the ISI Web of Science up to March 2009 (period for which analysis was done from 1971 to March 2009). See Table 2.1 for country identification.

Scientific perception of alien conifers in Europe

Many papers cited the region of origin of the conifer species but did not explicitly mention the invasion status of the species in the introduced region (e.g., Day 1984; Walsh et al. 1999; Varbergen et al. 2003; Straw et al. 2005; Barbagallo et al. 2005; Straw et al. 2006; Mason 2007). We deemed that the context of biological invasions existed if the authors used the standard terminology of “exotic“, “alien“, “non-native“, “introduced“, “aloctonous“, “invasive“ or “neophyte“ to describe the species. We also recorded if the authors mentioned other terminology common in the field of biological invasion, e.g., invasiveness, invasibility, range expansion or species spread. Only a third of the papers (34%) considered the species to be alien or studied an aspect related to invasion biology.Since the biological invasion literature has increased exponentially in the last decade in Europe (Pyšek et al. 2008), we related whether the perception (yes or no) of

54 Little evidence of conifer invasions alien conifers in the studies has increased though time by a binary logistic regression with publication year as the explanatory variable. We did not find a significant relationship between perception trend and publication year (X2 =9.384; r2= 0.07; P < 0.005; n = 131). This suggests that alien conifers in Europe have not yet piqued the interest of ecologists working on biological invasions.

It was only possible to identify the invasion status for 72 of the 212 records. Fifty-seven records (27%) corresponded to non-established conifers, eleven (5%) to established and four (2%) to invasive. All records on invasive conifers referred to P. strobus in the Czech Republic (Hanzélyová 1998; Pyšek et al. 2002; Krivánek et al. 2006; Hadincová et al. 2008). Most cases of establishment were found in the UK and referred to A. grandis, Abies procera, Cupressus macrocarpa, L. kaempferi, P. radiata, P. strobus and P. menziesii (Crook 1997). P. menziesii is also considered established in the Czech Republic (Pyšek et al. 2002; Krivánek et al. 2006), Spain (Broncano et al. 2005) and Switzerland (Wittenberg 2005) (Table 2.1).

55 Chapter 2

Topics of study involving alien conifers in Europe

Although most papers did not explicitly focus on topics related to biological invasions, we attempted to classify all papers with regard to the following aspects of invasion biology: (a) invasiveness: examination of plant performance and demography without any specific association to site characteristics; (b) invasibility: exploration of the biotic and abiotic factors in the study site that determine the success or performance of the species; and (c) impacts: analysis of ecological patterns (e.g., species richness) or processes (e.g., nutrient cycling) between areas where the species occurs or between species. Any mention of health or economic impacts was also recorded (Fig. 2.4).

3%3%

34%

Invasiveness Invasiveness Invasibility-bioticInvasibility-biotic factors factors Invasibility-abioticInvasibility-abiotic factors factors Ecological impactsEcological impacts 52% Others impacts Other impacts

2%

9%9%

Fig. 2.4. Percentage of papers (n=131) for studies of alien conifers in Europe found at the ISI Web of Science up to March 2009 classified by the topic of study.

56 Little evidence of conifer invasions

Invasiveness of alien conifers

Only 12 papers (9%), constituting 52 records, examined plant performance. These records were conducted in five countries (UK -32; Czech Republic -13; Spain -4; Germany -2; and, Switzerland -1). All papers were published in the last decade and the study species were always considered as alien species by authors. They examined, for example, tree growth rates (Willoughby et al. 2007; Mason 2007) and seed production (Krivánek et al. 2006). A few papers have classified the conifers’ invasion status according to their population growth and rate of spread (Crook 1997; Lavery & Mead 1998; Pyšek et al. 2002; Chytrý et al. 2008). The majority of these studies focused on P. menziesii (e.g., Crook 1997; Kowarik 2005), P. strobus (e.g., Krivánek et al. 2006; Hadincová et al. 2008) and P. radiata (e.g., Lavery & Mead 1998; Lombardero et al. 2008).

Invasibility to alien conifers

The biotic and abiotic characteristics of the study site where conifers were introduced have been the main area of interest (Fig. 2.4) with seventy-one papers (54%) accounting for 96 records. Most papers focused on conifer species as the host tree of natural enemies (94%). Curiously, a large proportion of these studies described the entomology of the natural enemies (e.g., Parry 1979: Evans 1985; Masutti & Battisti 1990; Wilson & Day 1995; Alonso-Zarazaga & Goldarazena 2005; Bajo et al. 2008) rather than focusing on the damage they inflicted on the conifers (e.g., Nichols 1987; Pulkkinen 1989; Watt et al. 1992; Day et al. 1999; Fabre et al. 2004). A remarkable number of publications deal with the life cycle of insects, mainly Diptera, Hymenoptera and Coleoptera. The insects most studied were those which cause major damage to economically important alien conifers (Battisti 2006), such as Elatobium abietinum and Thaumetopoea pytiocampa (12 and 4 papers, respectively), and P. sitchensis was the most studied host conifer (32 records). The majority of studies were carried out in the UK (35 records).

Alternatively, the other 6% of papers about abiotic factors focused mainly on the influence of soil properties (e.g., Pedersen& Bille-Handse, 1995; Hanzélyová 1998) and climatic conditions (e.g., Parry 1980; Cuadros & Francia 1999) on tree performance.

57 Chapter 2 Impacts of alien conifers

Forty-eight papers (37%) with 64 records were classified as studies on impacts. P. sitchensis and P. menziesii were the most studied species (30 and 9 records, respectively), and the UK and the Republic of Ireland were the countries with the most records (32 and 9, respectively).

Many studies explored the effects of the conifers on biodiversity, focusing on, for example, the avifauna (e.g., Tellería 1983; Carrascal & Tellería 1990), rodents (Fernandez et al. 1994), spiders (Oxbrough et al. 2006), weevils (Parry et al. 1990), squirrels (Gurnell et al. 2004), deer (Staines et al. 1985) and wild boar (Irizar et al. 2004). Other studies include Arbea & Jordana (1988), who examined the effect of reforestation on the edaphic collembola community composition; Amezaga & Rodríguez (1998), who studied the niche relationships among four sympatric bark beetles with respect to conifer species (P. radiata) and swarming time; and Amezaga (1996), who showed the sustainability of Monterrey pine for the pine shoot beetle (Tomicus piniperda L.). Overall, despite major efforts in studying insects attacking alien conifers, this research does not provide support for the natural enemy hypothesis which states that alien species in the introduced range are less damaged by pathogens and herbivores than in the native range (Maron & Vilà 2001).

A few studies have focused on ecosystem processes affected by alien conifers. For example, Harriman & Morrison (1982) studied the combined and individual effects of acid precipitation and coniferous afforestation on stream ecology.

In summary, although we classified these studies as dealing with “impacts”, the conifers were not of invasion concern in most cases (only 27%), and the study simply documented differences in ecological patterns and processes between conifer species or between forested stands. Only two papers dealt with economic impacts: the effect of bark stripping on timber production (Welch & Scotch 1998) and the effect of forest management practices on the age structure and composition of forests (Mason 2007). Finally, one paper analysed the allergenic impact of cypress pollen on human health (Charpin et al. 2005).

58 Little evidence of conifer invasions

DISCUSSION

Despite the fact that there are many studies on alien conifers in Europe, few ecological aspects related to biological invasions have been investigated. The invasion status of these alien conifers is not mentioned in most cases suggesting either that (1) scientists have not perceived them to be of conservation concern, despite evidence that many of these species are invasive in the Southern Hemisphere (Richardson & Higgins 1998), or (2) such conifers are not expanding as notably as in other parts of the world (Mortenson & Mack 2006; Simberloff et al. 2010).

Studies have concentrated on species of greater economic relevance such as P. sitchensis and P. menziesii. Additionally, the lack of terminology related to biological invasions is partly due to the fact that the main conservation interest has been alien conifers as hosts of harmful insects rather than the conifer itself (Pyšek et al. 2004). Moreover, the published information is likely biased to those countries with a longer tradition in the study of forestry species (e.g., the UK), and possibly substantially more evidence could be found in the grey literature not listed in the ISI Web of Science. We only found information for 37 of the 54 alien conifers in Europe listed in DAISIE. Among them, ISI papers only mention 7 species as naturalized and one species (P. strobus in the Czech Republic) as invasive. In contrast, Rejmánek & Richardson (2004), whose review also considered other scientific papers, the grey literature and unpublished sources, listed 18 species as naturalized and 9 as invasive. For example, naturalization of P. strobus has also been documented in Poland and Austria according to non-ISI papers (Adamowski 2004; Essl 2007).

There are several explanations for the reducer invasion status of alien conifers in Europe as compared to the Southern Hemisphere. These explanations could be classified first based on the historical and socioeconomic features which have determined its introduction and second, the ecologically-based causes which have limited their establishment and spread.

First, one well-grounded generalization is that the probability of invasion increases with introduction effort (i.e., propagule pressure) and time since introduction (Rejmánek 2000). In trees, the probability of becoming invasive is primarily determined by the introduction history of the species, such as time since introduction and the spatial scale of the plantations (Pyšek & Jarošik 2005; Bucharova & Van Kleunen 2009; Simberloff et al. 2010).

59 Chapter 2

The majority of alien conifers in Europe were introduced during the last century. The earliest record in Europe seems to be from 1800 for P. strobus in the Czech Republic (Hanzélyová 1998; Pyšek et al. 2002; Krivánek et al. 2006; Hadincová et al. 2008). In contrast, pine introduction in the Southern Hemisphere started during the 17th century (Richardson & Higgins 1998) and the scale of these introductions has increased dramatically in the last century, particularly in South America (Pauchard et al. 2004; Richardson et al. 2008). Time lags are important determinants of the invasions process (Crook 2005).

In a recent review of alien conifers introduced in the Southern Hemisphere, Simberloff et al. (2010) suggested that the later introduction of large conifer plantations in South America, compared to other regions of the Hemisphere such as South Africa, might lead to an increase of invasions into the region in the near future. Kowarik (1995) determined the lag-time for 184 woody species and has suggested that approximately 150 years elapsed before species began to escape from cultivation. Both time since introduction and introduction effort are probably smaller in Europe compared to the Southern Hemisphere regions. As an example, 970.000 and 18.500 ha of alien conifers have been planted in the UK (Forestry Commission, 2003) and in the Czech Republic (Krivánek et al. 2006) in the last century, respectively. For the same time period, 2.05 and 1.5 million ha of alien conifers have been planted in Chile and in Brazil, respectively (Simberloff et al. 2010).

The smaller scale of alien conifer plantations might be due to some non- exclusive reasons. In the last few decades, Europe has changed forest management policies towards a more conservative approach that emphasizes awareness of the use of native species for planting (Quine et al. 2004). European foresters have long had native fast growing tree species, most of them conifers, for fibre production and restoration services while, for example, foresters in South America needed to introduce alien conifer species due to the lack of commercial native trees for planting. In South America, plantations of alien conifers have probably been exacerbated by foreign forestry investment and the potential of wood exports to the Northern Hemisphere (Nuñez & Pauchard 2009). Moreover, the area of Europe is much less than that of South America and the scale of plantations has therefore likely been proportional.

60 Little evidence of conifer invasions

Second, invasion differences might also depend on similarities between native and alien species. Empty niches prone for conifer exploitation are probably less common in the Northern Hemisphere than in the Southern Hemisphere. For example, South African fynbos has no native tree species and therefore is prone to invasion by alien fire-adapted trees like pines (Richardson & Brown 1986). Increased attention is being paid to the effects of phylogenetic relationships between the alien and native flora on invasion success. The naturalization hypothesis proposes that novel genera with native representatives should be less successful than genera lacking them in the native flora (Strauss et al. 2006). Phylogenetic proximity between alien and native trees can result in higher colonization rates of natural enemies on alien trees (Goßner et al. 2009). Most alien conifers in Europe have a temperate North American and Asian origin and have European congeners (e.g., the genus Pinus). Thus, once introduced, these conifers might recruit pathogens and phytophages from closely related species which in turn would decrease their capacity to spread in the new range.

The same argument was put forth by Mortenson & Mack (2006) who observed that some alien conifers planted in large extensions in North America are heavily attacked by pathogens in the introduced areas where they coexist with native conifers. Our review showed that a high proportion of published papers focused on the natural enemies associated to the alien conifers. This suggests that, in Europe, alien conifers might be strongly controlled by phytophages and pathogens, and this may be one reason why invasion events are less common in Europe.

Nevertheless, recent research has noted the importance of mycorrhizal symbiosis in explaining invasion success in conifers (Pringle et al. 2009). It may be easier for alien conifers to get soil mutualists in areas where there are native Pinaceae (Collier & Bidartondo 2009; Nuñez et al. 2009), which might be beneficial for their establishment. All else being equal, the success of invasion might depend on the balance between the role of natural enemies and mutualists. These two opposing biotic forces, in general, have been explored in isolation and deserve further exploration.

61 Chapter 2

CONCLUSIONS

Our review illustrates a gap in the understanding of alien conifer invasion in Europe. From the management point of view, we recommend the careful observation of sites where alien conifers are present to watch for new invasions (Richardson & Rejmánek 2004). From the scientific perspective, research on alien conifers should have a biogeographical focus. For example, a global analysis about the extent of introduction, rates of naturalization and regional differences influencing invasive spread between the two hemispheres is needed. Experimentally, parallel experiments following standard protocols on the biotic and environmental mechanisms controlling seed and seedling performance in different regions and ecosystems would shed light onto differences in ecosystem resistance to alien conifers.

Acknowledgements

We thank the DAISIE consortia for plant species data compilation and harmonization, and D Stouffer, DM Richardson, MA Nuñez and A Pauchard for comments on an early version of this manuscript. Research was partially funded by the Spanish Ministerio de Ciencia e Innovación project Consolider-Ingenio MONTES (CSD2008-00040).

62 63 64 Chapter 3

Comparing seed removal o f 16

pines species

differing in invasiveness2

2Carrillo-Gavilán MA, Hadrien L, Vilà M (2010) Comparing seed removal of 16 pines species differing in invasiveness. Biological Invasions 12:2233-224.

65 Chapter 3

Abstract

Small seed mass is regarded as a robust trait related to invasion success, especially in pines. However, few studies have explored whether invasiveness related to small seed size is also associated to low levels of seed predation in the recipient community. We conducted field cafeteria seed removal experiments comparing 16 Pinus species that differ in seed mass to test if seed removal might impose biotic resistance to Pinus spp. and if there are differences between species related to seed mass. Seeds were removed rapidly and in high proportion. In the Mediterranean shrublands, where the experiments were conducted, rodents and ants were the main seed removers. Mean seed survival time was significantly different between species. However, smaller seeds were not the most predated. Our study suggests that, in pine species with high invasiveness, the potential higher seed removal of small seeds can be counterbalanced by larger seed crops.

66 Comparing seed removal of pine species

INTRODUCTION

Invasion success depends mainly on invasiveness (i.e., the intrinsic potential of an alien species to invade), invasibility (i.e., the susceptibility of the recipient community to be invaded) (Lonsdale 1999), and propagule pressure (i.e., the number of individuals introduced and the number of introduction attempts) (Williamson 1996; Lockwood et al. 2005; Colautti et al. 2006), aspects that have tended to be studied in isolation.

Many studies have tried to find life-history traits that determine a high invasive potential (reviewed in Pyšek & Richardson 2007). The most cited studies on the plant traits that determine invasiveness likely correspond to those comparing invasive and non-invasive pine species (Rejmánek & Richardson 1996; Grotkopp et al. 2002). The genus Pinus contains approximately 111 species (Price et al. 1998) and 18% are reported to be invasive in at least one country (Richardson & Rejmánek 2004). Three traits characterise pine invader species (Rejmánek & Richardson 1996): a short juvenile period, a short interval between large seed crops and a small seed mass that is correlated with a fast relative growth rate (RGR) (Grotkopp et al. 2002). Small seed mass seems to be a robust trait related to invasion success at several spatial scales (Hamilton et al. 2005). It may play a significant role in facilitating species arrival to a new range since small seed mass increases successful dispersal, prolongs persistence in the soil (Thompson et al. 1993) and potentially escape from seed removers (Richardson et al. 2000).

Even for native species, however, there is no consensus as to whether small- seeded species escape more often from seed removal than those with large seeds, with the former consequently having a higher establishment success. This depends on a life-history trade-off between seed mass and seed production, viz. small seeds are produced in greater quantities than large seeds (Jakobsson & Eriksson 2000; Henery & Westoby 2001; Moles & Westoby 2004). Moreover, there is a preference for seed size according to predator type (Ordóñez & Retana 2004). For instance, in herbaceous areas in the UK, there is higher predation by rodents on larger (seed size >1mg) than on smaller seeds (Hulme 1998) while ants were the main predators on very small seeds (Andersen 1987; Nepstad et al. 1991). In degraded ecosystems, the absence of large mammals can lead to higher predation on smaller than on larger-seeded species (Hang-Hau 1997; Dirzo & Mendoza 2007).

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Native seed predators act as barriers to invasion by decreasing the number of seeds available for establishment in the new range (Vilà & Gimeno 2003). Nevertheless, few studies have investigated the link between seed size and post-dispersal seed predation in invasive species. In Argentina, Nuñez et al. (2008) have found that seeds of invasive conifers, two of them pine species, are more consumed than native tree species probably because seeds are larger and more attractive to native rodent and bird seed predators. Large seed predation thus determined lower alien seedling emergence.

We investigated the likely invasion success of 16 Pinus species differing in seed size in disturbed Mediterranean shrublands by conducting field cafeteria seed removal experiments to answer the following questions: 1) Is seed removal imposing a putative biotic resistance to Pinus spp. establishment? 2) Are there differences in seed removal between species? 3) Are differences in seed removal explained by seed size? and finally 4) Who are the main seed removers? Our hypothesis is that Pinus spp. shows significant differences in seed removal in relation with seed size. We expected small-seeded species to show lower seed removal and to have larger seed-survival times than large- seeded species. Given that many alien pine species have been experimentally planted in disturbed Mediterranean areas, this study is also an attempt to improve invasion risk analysis for introduced tree species.

MATERIALS AND METHODS

Study area

The study area was located in disturbed sclerophylous shrublands in Madroñalejo - Aznalcóllar hills (Spain) (37º34’N and 6º20’W) at an average altitude of 292 m a.s.l. (Fig. 3.1). The climate is Mediterranean with warm, dry summers and cool, wet winters. According to the nearest meteorological station in El Campillo (Huelva), the mean annual temperature is 18ºC. The mean minimum and maximum temperatures are reached in January (5.4ºC) and in July (34.8ºC), respectively. The mean annual precipitation is 679 mm. Geomorphologically, the area is dominated by ardoise schists rich in iron oxides (Lamy J.B., personal communication) and quartzite outcropping at the soil surface. Soils are red, with sandy-loam texture, very poor in organic matter and a very acidic pH (Navarro et al. 1998).

68 Comparing seed removal of pine species

a b

Fig. 3.1a and b. Study area dominated by Mediterranean shrubs (photo 3.1b by H. Lalagüe). This area has been highly disturbed by forestry management (e.g. clearing, afforestation) and by consecutive fires. The last wildfire was in summer 2004. Prior to the fire, the vegetation was composed of scattered holmQuercus oaks( ilex sub. ballota), cork oak (Q. suber) and reforested Pinus pinea with a shrubland understorey (Navarro et al. 1998). Two years after the wildfire the area was mainly reforested by Q. suber and Olea europaea sylvestris. Currently, the vegetation is dominated by Cistus salvifolius (15.4 % plant cover) and C. ladanifer (13 %) followed by Genista tridendata (8 %), Ulex eirocladus (4.25 %) and G. triacanthos (4%).

Study species

Sixteen Pinus species, which are broadly commercialized and planted worldwide, were used for analysis (Table 3.1). Of these, nine species have been planted in Spain (Montero et al. 2005) but only two (P. radiata and P. ponderosa) have escaped from cultivation but are not considered naturalized in Spain (Sanz-Elorza et al. 2004). Only P. pinea is native in the study area.

Seed mass of these species differ in two orders of magnitude ranging from 4 mg (P. contorta) to 829 mg (P. sabiniana). To guarantee a high seed quality, seeds were purchased in the following nurseries: Intersemillas (Spain), Sheffield´s Seed Co. (EE.UU) and Les Semences Du Puy (France) (http://www.intersemillas.es, http://www. sheffields.com/ and http://www.semencesdupuy.com/, respectively).

For each species, 20 seeds were randomly chosen to measure fresh seed mass, length and width. Coat strength was also estimated following Rogerdson’s (1998) protocol with a Chatillon Universal Force Tester (Amtek/Chatillon, Largo, Florida, USA).

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Given that seed mass is strongly positively correlated with strength (r2 = 0.95; P < 0.001), seed length (r2 = 0.84; P < 0.001) and seed width (r2 = 0.77; P < 0.001), seed mass was used as the seed size parameter for our analysis.

Table 3.1. Mean (± SE) characteristics of the Pinus species assayed. Native species distribution based on Montero et al. (2005) and http://www.arbolesornamentales.com.

Seed traits Mass Strength Length Width Native Abbrs Pinus spp. (g) (Newton) (cm) (cm) distribution Northern USA B2 P. banksiana 0.004 ± 0.0001 4.2 ± 0.32 0.37 ± 0.005 0.20 ± 0.003 and Canada CT2 P. contorta 0.004 ± 0.0002 13.5 ± 0.96 0.43 ± 0.012 0.23 ± 0.01 Western North- America RE2 P. resinosa 0.009 ± 0.0003 13.3 ± 0.41 0.43 ± 0.006 0.25 ± 0.004 USA and Canada S1 P. sylvestris 0.009 ± 0.0004 9.5 ± 0.7 0.47 ± 0.009 0.26 ± 0.004 Europe and Asia PA2 P. patula 0.009 ± 0.001 11.9 ± 1.38 0.52 ± 0.121 0.26 ± 0.075 Central America M2 P. muricata 0.012 ± 0.0005 11.4 ± 1.06 0.54 ± 0.009 0.30 ± 0.007 Western USA and Northern Mexico H1 P. halepensis 0.017 ± 0.0009 13.5 ± 0.63 0.59 ± 0.013 0.34 ± 0.072 Mediterranean Basin R2 P. radiata 0.03 ± 0.001 19.2 ± 1.20 0.69 ± 0.015 0.40 ± 0.008 Californian Coast PO2 P. ponderosa 0.049 ± 0.002 62.1 ± 3.22 0.71 ± 0.015 0.47 ± 0.007 From Southern Canada until Mexico P1 P. pinaster 0.053 ± 0.002 60.5 ± 2.16 0.74 ± 0.022 0.48 ± 0.02 Western Mediterranean and Atlantic area of France and Portugal PT P. palustris 0.077 ± 0.003 27.6 ± 1.36 0.98 ± 0.037 0.57 ± 0.014 North America RO2 P. roxburghii 0.087 ± 0.004 40.8 ± 2.06 1.09 ± 0.028 0.57 ± 0.014 Afganistan, Butan, India, Nepal and Pakistan C1 P. canariensis 0.112 ± 0.052 69.4 ± 3.61 1.25 ± 0.032 0.62 ± 0.013 Canary Islands CO P. coulteri 0.347 ± 0.155 161.2 ± 7.45 1.45 ± 0.021 0.82 ± 0.027 California, Mexico PI1 P. pinea 0.62 ± 0.030 509 ± 22.64 1.61 ± 0.026 0.86 ± 0.022 Mediterranean Basin SA2 P. sabiniana 0.829 ± 0.021 489.8 ± 26.36 2.09 ± 0.02 0.96 ± 0.014 California

1 Native species in Spain; only P. pinea occurs in the study area. 2 Alien species planted in Spain.

70 Comparing seed removal of pine species

Seed removal experiment

In 2007, we placed four 100 x 100 m plots, 0.2-3 km apart. In each plot, we placed ten to fourteen 50 m long transects. In each transect, we randomly placed seeds on the ground surface keeping a distance of approximately 3 m between each other. Each seed was glued to a piece of nylon fishing line that was tied to a wire stake (Schupp 1988). Glue was odourless and tasteless.

In order to consider time variation in seed removal, we repeated the experiment twice during the seasons when pine seed dispersal might occur (Lanner 1998; Tapias et al. 2004). In Summer, we used 14 species (n = 80 x 14 = 1120 seeds). In Autumn, we used 16 species (n = 80 x 16 = 1280 seeds). For commercial reasons, we could not use the same number of species in both seasons.

In Summer 2007, seeds were checked for removal after 1, 3, 7, 10, 16, 24, 31, 38, 52 and 59 days. In Autumn 2007, we introduced two more species: P. pinea and P. contorta and seeds were checked after 1, 2, 3, 7, 13, 21, 34 and 41 days. We considered either missing seeds or the presence of seed coat remnants as evidence of seed removal. We finished the experiment when, after two consecutive visits, the seed removal curves did not vary. After that day, all seeds were removed from the field.

We conducted two analyses. First, differences in seed removal (i.e., percentage of predated seeds at the end of the experiment) between species, seed mass class (smaller versus larger seeded species) and range origin (native versus alien) for each season were analyzed using SAS macro GLIMMIX (Littell et al. 1996) with binomial error and a logit link function. Species (n = 14 in Summer and n= 16 in Autumn), seed mass class and range origin were considered as a fixed factors and plot (n = 4) as a random factor. The interaction between each fixed factor and plot was considered as a random effect. Second, Kaplan-Meier estimations of the mean survival time of seeds were calculated. Differences of mean survival time among species for each season were compared with the Gehan-Wilcoxon test (Pyke & Thompson 1986). Seasonal comparisons were not conducted because the time duration of experiments was different. The software SPSS 14.0 was used for survival analysis (SPSS Inc., Chicago, Illinois, USA). Additionally, simple regression analysis between seed mass and seed removal and mean survival time were conducted for each season

71 Chapter 3

Seed removal trials with seed remover exclusion

To ascertain whether species differences in seed removal could be related to seed mass and to discriminate between seed remover guilds, three species with the lowest seed mass, P. halepensis, P. patula and P. contorta, and three with the highest seed mass, P. sabiniana, P. pinea and P. coulteri (Table 3.1), were chosen in March 2008 to compare the effectiveness of seed removal by birds, ants and rodents.

The device allowing seed access only for birds (i.e., preventing ants and rodents) was composed of a plastic curved tray suspended at a height of 15-20 cm above the ground on a wood stick. This system has been successfully tested with small birds by placing the device at 30 cm height (Castro et al. 1999). In the study area, however, there are a large number of red-legged partridge (Alectoris rufa) that are good walkers but fly with moderation. Therefore, we reduced the height to improve the potential access for seed consumption after testing it in wildfowl hatchery. A plastic cup (9 cm high and 6.5 cm diameter) was positioned against the bottom of the tray in order to prevent rodent access. The bottom of the tray was painted with a Teflon® emulsion (i.e., polytetrafluorethylene) to make the surface of the tray sticky for climbing ants (Hulme 1997; Fedriani et al. 2004). The tray was filled with soil and the seed was placed in the centre of the tray (Fig. 3.2a).

b)

a) c)

d)

Fig. 3.2. The four excluding experimental devices used to estimate seed removal by each remover taxon. a: bird access, b: ant access, c: rodent access, d: open access to all.

72 Comparing seed removal of pine species

The device for ants consisted of a transparent plastic tube 8 cm in length and 1.2 cm diameter in which the seed was placed inside. The tube was fixed to the soil with U-shaped wire (Castro et al. 1999; Deveny & Fox 2006; Jacob et al. 2006; Fig. 3.2b). The diameter was the minimal size that could be used to allow the removal of P. sabiniana (Table 3.1). The treatment revealing rodent activity was made with an opaque plastic tube 20 cm in length and 3.5 cm diameter. The tube was also fixed to the soil with U-shaped wire. To prevent seed displacement by ants, the seeds were glued to 3×3cm pieces of fine plastic mesh and placed inside the tube (Ordóñez & Retana 2004; Fig. 3.2c). Finally, the last treatment consisted of small Petri dishes 4.5 cm in diameter and 0.5 cm high fixed to the soil with a wire. This system allows all animals to remove seeds (Fig. 3.2d, open treatment hereafter).

The four exclusion treatments were randomly placed along 40 parallel transects 70m long, at approximately 1.5 m intervals. Each transect was approximately 2 m away from the next. There were 80 replicates for each treatment and species. Thus there were a total of 1920 seeds displayed in the field (6 species × 4 treatments × 80 replicates = 1920 seeds). Seed removal was monitored at 2, 6 and 13 days after seed placement.

For this experiment, we also conducted two analyses. Firstly, seed removal at the end of the experiment was compared between exclusion treatments, between species and for seed mass class (smaller versus larger seeded species) using SAS PROC GENMOD´s with binomial distribution and a logit link function (v9.1, SAS Institute Inc) with species, treatment and seed mass class as fixed factors. Secondly, a survival analysis was also computed following the same procedure as in the previous seed- removal experiments. Mean survival time was compared among treatments, among species within each treatment and between seed mass class.

Identity of seed removers

Rodent trapping was conducted in all four plots at the end of each experiment (Summer, Autumn and Spring). Fifteen Sherman traps (H. B. Sherman Traps, Inc., Tallahassee, Florida) were distributed in each experimental plot at dawn keeping a minimum distance of 6 m between them. Traps were surveyed at sunset the next day. Inside each trap, we placed a piece of bread with chocolate spreading. The protocol was repeated during 3 consecutive days (Fedriani 2005). Simple indexes of relative abundance of rodents were calculated for each plot and season as: the number of captured individuals / trapping effort (number traps x trapped nights) (Fig. 3.3a).

73 Chapter 3

Differences in the relative abundance of rodents between seasons were analysed by using SAS macro GLIMMIX (Littell et al. 1996) with binomial error and a logit link function. Season was considered as the fixed factor and plot as a random factor.

Pitfall traps were used to identify ant species foraging on the ground only in spring 2008 (Fig. 3.3b). Pitfall traps were 6.5 cm-diameter, 9 cm-deep plastic cup partially filled with water, ethanol and soap (Romero & Jaffe 1989). Fifteen traps were in place for 48 h in parallel with the rodent trapping. Ants were then sorted in the laboratory to the species level. Finally, common birds of the study area that might remove seeds were identified in spring 2008 according to visual identification and song.

a b

Fig 3.3a and b Examples of rodent (Apodemus sylvaticus) and ants (Crematogaster spp.) as pine seed removers in the study area

74 Comparing seed removal of pine species

RESULTS

Seed removal experiment

In both seasons, mean seed removal was very intense for all species at 90.53% in Summer and 93.03% in Autumn (Fig. 3.4a). There were no significant differences in seed removal neither between species (F13, 39 = 1.60; P = 0.13 in Summer and F15, 45

= 0.66; P = 0.81 in Autumn) nor between seed mass classes (F1,3 = 0.78; P = 0.44 in

Summer and F1,3 = 0.24; P = 0.65 in Autumn) or between origin range (F1.3 = 0.13; P =

0.74 in Summer and F1,3 = 1.48; P = 0.31 in Autumn) in any season. a) 105

100100

9595

9090

8585

Seed removal (%) Seed removal 8080 Seed removal (%) 7575

7070 B CT RE S PA M H R POPO P PT RORO CC COCO PI SA b) 25

20

15

10 Survival time (%) Survival time (days) time Survival 55

00 B CT RE SS PA M H R PO P PT RO C CO PIPI SA Pinus spp Pinus spp Fig. 3.4 Mean (± SE) seed removal (a) and mean (± SE) survival time (b) of pine species in Summer (grey bars) and in Autumn (white bars). Species are ordered from small to large seed mass. See Table 3.1 for species identification.

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There was a small significant negative relationship between seed removal and seed mass in Autumn (r2 = 0.37; P = 0.013) but not in Summer (r2 = 0.107; P = 0.254) (Fig. 3.5). However, this correlation was mostly driven by P. sabiniana, the species with the largest seeds. The relationship disappeared when this species was removed from the analysis (r2 = 0.017; P = 0.662 in Summer and r2 = 0.042; P = 0.462 in Autumn.)

100100

95

90 r2 = 0.36

85 r2 = 0.10 80 Seed removal (%) removal Seed Seed removal (%) 7575

7070 0 0.10,1 0.20,2 0.30,3 0.40,4 0.50,5 0.60,6 0.70,7 0.80,8 0.90,9 SeedSeed mass mass (%) (g) Fig. 3.5. Relationships between seed mass and percentage of mean seed removal of pine species in Summer (black points and continuous line) and in Autumn (white points and dotted line).

Mean survival time was shorter in Autumn (5.24 days ± 0.31) than in Summer (13.11 ± 0.55). There were significant differences in mean survival time between 2 species for both seasons (Gehan-Wilcoxon test; X 13= 48.36, P < 0.001 in Summer 2 and Gehan-Wilcoxon test; X 15= 83.88, P < 0.001 in Autumn) (Fig. 3.4b). The shortest and longest mean survival times were for P. radiata (8 ± 1.3) and P. patula (19.9 ± 2.5) in Summer and for P. ponderosa (2.5 ± 0.76) and P. sabiniana (9.41 ± 1.67) in Autumn, respectively., Mean survival time, however was not related to seed mass in any season (r2 = 0.001; P = 0.924 in Summer and r2 = 0.201; P = 0.082 in Autumn).

Seed removal trials with seed remover exclusion

Overall, 63.73% of seeds were removed after 13 days. There were significant differences 2 in seed removal between exclusion treatments (X 3 = 928.7, P < 0.005). The highest quantity of seeds removed (94.10%) occurred in the open treatment (Fig. 3.6a) followed

76 Comparing seed removal of pine species by ants (79.87%) and rodents (72.03%). The bird treatment had the lowest seed removal (8.94%). a) a 100 ab ab ab a ab a a a 90 b ab a a 80 ac bcd 70 d

60 b

b 50

40 Seed removal (%) Seed removal (%) 30 a 20 b 10 b b b b 0 CT PA H CO PI SA CT PA H CO PI SA CT PA H CO PI SA CT PA H CO PI SA CT All CO CT Ants CO CT Birds CO CT Rodents CO b) All Ants Birds Rodents 14 b ab ab a a a c b 12 a a b c

b 10 bc c bc d b b d c bc d 8 b c

6 Survival time (days) Survival time (%) 4

2

0 CT PA H CO PI SA CT PA H CO PI SA CT PA H CO PI SA CT PA H CO PI SA CT CO CT CO CT CO CT CO All All Ants Ants Birds Birds Rodents Rodents

Fig. 3.6. Percentage of seed removal (a) and mean survival time (b) of pine species in different exclusion treatments. Lower-case letters represent statistically significant differences (P ≤ 0.05) between species within treatments. Species are ordered from small to large seed mass. See Table 3.1 for species identification.

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On average, there were significant differences in seed removal between species 2 (X 5 = 56.11, P < 0.005). The most predated species were P. contorta, P. halepensis and P. patula (72.26%, 70.86% and 70%, respectively). P. coulteri and P. sabiniana had an intermediate seed removal percentage (63.30% and 55.22%, respectively), and the least removed was P. pinea (50.69%). However, the treatment by species interaction 2 was significant (X 23 = 1075.29, P < 0.05) indicating that these differences between species were not consistent across all exclusion treatments. In general, there were also 2 significant differences between seed mass classes X( 1 = 46.40, P < 0.005): species with smaller seeds being more predated than larger seeded-species (69.60% and 2 55.88%, respectively) and this trend was observed for all treatments (open: X 1= 5.03, 2 2 2 P < 0.05; ants: X 1=18.84, P < 0.0001; birds: X 1= 12.93, P < 0.005; rodents: X 1= 40.84, P < 0.0001) (Fig.3.6a).

Mean survival time was significantly different between exclusion treatments 2 (Gehan-Wilcoxon test; X 3= 649.6, P < 0.0001) and between species (Gehan- 2 Wilcoxon test; X 5= 186.76, P < 0.0001). Seeds in the open treatment showed the shortest survival time (8.46 ± 0.21) followed by the seeds removed by ants (9.28 ± 0.2), rodents (10.14d ± 0.19) and, finally birds (12.67 ± 0.08) (Fig. 3.6b).

Within each treatment, there were significant differences in mean survival time 2 2 between species (open: X 5= 49.53, P < 0.0001; ants: X 5= 37.19, P < 0.0001; birds: 2 2 X 5= 32.4, P < 0.0001; rodents: X 5= 114.12, P < 0.0001) (Fig. 3.6b). On average, mean survival time for smaller seeded-species (9.14 days ± 0.15) was significantly shorter than for larger seeded-species (11.03 days ±0.12) (Breslow test: X2= 88.87, P < 0.0001).

Identity of seed removers

The rodents captured in the area were Mus spretus and Apodemus sylvaticus (Muridae). In Summer and in Autumn, M. spretus was captured in a higher proportion than A. sylvaticus. There were significant differences in the relative abundance of rodents between seasons (F2, 6 = 6.62; P = 0.03) being lowest in Spring (Table 3.2).

78 Comparing seed removal of pine species

Table 3.2 Relative abundance index (number of individuals) of rodents in the study area. The relative abundance was calculated as the number of captured individuals / trapping effort, where trapping effort is the number traps x trapped nights.

Summer 2007 Autumn 2007 Spring 2008 Mus spretus 14.4 (26) 15.5 (28) 1.1 (2) Apodemus sylvaticus 8.3 (15) 3.8 (7) 1.1 (2) Total 22.7 (41) 19.4 (35) 2.2 (4)

We captured 9 different ant species of which Tetramorium impurum, Pheidole pallidula and T. caespitum are granivorous. The granivorous birds were predominantly: the red-legged partridge (Alectoris rufa, Phasianidae); Carduelis carduelis, C. cannabina, Fringilla coelebs, Miliaria calandra and Serinus serinus (Fringillidae); Parus major (Paridae); and Galerida cristata (Alaudidae).

DISCUSSION

Native fauna was very efficient in removingPinus spp. seeds (> 90%), exhibiting similar values to other post-dispersal seed removal experiments for pine species (Vander Wall 1992; Castro et al.1999; Borchert et al. 2003). Although our study should be considered a field cafeteria experiment, our study showed evidence that the native fauna can be an important component of biotic resistance to pine establishment (Elton 1958; Maron & Vilà 2001, Vilà & Gimeno 2003, Nuñez et al. 2008). In our study area, the main seed removers were ants and rodents, which, together with birds, are also considered the main post-dispersal seed removers of the native plant species (Herrera 1984; Hulme & Hunt 1999; Alcántara et al. 2000; Andrieu & Debussche 2007).

There were significant differences in seed survival between pine species. Contrary to expectation, however, the trend was not towards lower seed removal of small-seeded species or towards differences between alien and native pines. Instead, species with smaller seeds tended to be faster removed. Small-seeded species produce seeds in larger quantities than larger-seeded species (Jakobsson & Eriksson 2000; Henery & Westoby 2001) and this can counterbalance the effects of seed loss (Moles et al. 2003; Moles & Westoby 2004). Our finding suggests that, even if pine invasiveness is associated with small seed mass, their seed crop is what might determine invasion success. This is in accordance with the importance that propagule pressure has on invasion success (Lockwood et al. 2005; Colautti et al. 2006). Many pine species that escape from plantations and become invasive are often those that have been cultivated the most widely and for the longest time (Krivánek et al. 2006).

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Our negative relationship between seed removal and seed mass increases the percentage of publications where smaller seeded species were removed more than larger ones (e.x., Moles et al. 2003; Myster 2003; Dirzo & Mendoza 2007). There might be several non-exclusive reasons for our results. Firstly, the optimal foraging theory (Krebs 1978) explains that large seeds are consumed more because of their greater energy content and they provide a higher energy intake per travel unit by seed consumers. Moreover, large seeded species have usually adapted to seed removal by seed caching animals (Vander Wall et al. 2006).

Nevertheless, smaller seeds are eaten by a larger guild of predators, and, moreover, the absence of large mammals in disturbed ecosystems can lead to higher removal of smaller rather than larger-seeded species (Hang-Hau 1997; Dirzo & Mendoza 2007). The negative relationship between seed removal and seed mass seems to be related to rodent abundance (Ordóñez & Retana 2004; Fedriani 2005) since rodent abundance was significantly lower in Spring than in Summer and Autumn. Secondly, an animal does not remove a seed only using the mass criterion. All physical (Rodgerson 1998) and chemical traits come into play (Janzen 1969). However, this is beyond the scope of this work. Thirdly, seed removal is only a step towards seedling establishment. Seedlings from larger seeds have higher probability of survival from seedling emergence to adulthood (Moles & Westoby 2006) and perform better during early development than smaller seeds (Moles & Westoby 2004). These explanations highlight the need to incorporate the elements of species invasiveness with those of habitat susceptibility to invasion (Richardson 2006) to reduce the idiosyncrasies found in the importance of species traits for invasion (Kolar & Lodge 2002; Muth & Pigliucci 2006).

Acknowledgements

We thank E Caballero, F Herce, E Manzano, A Montero and L van Oudenhove for field assistance; C Gómez and J Oliveras for seed strength measurements; C Alonso and JM Fedriani for statistical advice; D Stouffer and two anonymous reviewers for comments on an early version of this manuscript. The Consejería de Medio Ambiente (Junta de Andalucía) and EGMASA provided working facilities. This study has been partially funded by the EU integrated project ALARM (http://www.alarmproject.net/alarm) and the CONSOLIDER-INGENIO 2010 project MONTES (CSD2008-00040).

80 81 82 Chapter 4

Establishment constraints of an

alien and a native conifer in

different habitats3

3Carrillo-Gavilán MA, Espelta JM, Vilà M. Submitted to Biological Invasions.

83 Chapter 4

Abstract

Alien plants are subjected to different biotic and environmental barriers that limit their establishment success in the introduced range. Pseudotsuga menziesii (Douglas fir) is considered one of the most invasive forestry conifers of the world. However, little is known about the ecological filters that constrain plant establishment at early life-cycle stages and differences in habitat invasibility to this species. We conducted field experiments to compare the establishment potential (i.e., seed removal, seed germination, seedling survival and growth) of Douglas fir in beech forests, holm-oak forests and heathlands; and compared it with the taxonomically close native conifer Abies alba (Silver fir). Douglas fir seeds were more removed than Silver fir in holm-oak and in heathlands. In all habitats, seed germination was significantly higher for Douglas fir compared to that of Silver fir and, seedling mortality was extremely high inboth species due to soil disturbance by wild boars and water stress. Douglas fir mortality was only lower than Silver fir in beech forests. However, species did not differ in seedling growth. Overall, the probability of invasion success of Douglas fir decreased along the sequential stages of plant establishment in all habitats. Only high seed production and seed germination rates of Douglas fir would predict its high invasive capacity but these advantages are counterbalanced by high seedling mortality. Results showed a mismatch between invasibility and this species’ degree of invasion. Therefore, future research focused in mid life-cycle stages is recommended to elucidate which processes are favoring its establishment success.

84 Establishment constraints of conifer in different habitats

INTRODUCTION

Alien plants are subjected to different biotic and environmental barriers that limit their establishment success in the introduced range (Lonsdale 1999). Many of these barriers have an effect on early life-cycle stages. For example, by limiting the supply of seeds due to seed consumption by native fauna (McCay & McCay 2009), or also, by controlling seedling establishment either caused by unfavorable weather conditions (Lambrinos 2002) or by competition with native plants (Dietz et al. 1999). As in native plants, these ecological barriers vary spatially and temporally (Schupp 1988; Vilà & Lloret 2000; Traveset et al. 2003). For instance, post-dispersal seed removal varies among habitats, which can be more intense in areas far from alien conifer plantations than in others close to them (Nuñez et al. 2008).

There are studies that have compared differences in the establishment traits between alien and native congeners in the introduced range (Shafroth et al.1995; Lambrinos 2002; Ferreras & Galetto 2010) to better understand the invasion potential of the alien species (i.e., invasivenesss), and to determine differences in habitat susceptibility to be invaded (i.e., invasibility) (Pyšek & Richardson 2007). For this purpose, experiments are the best approaches to disentangle the mechanisms of invasion (Sol et al. 2008). Furthermore, these experiments need to link the causal relationships of invasion, from seed dispersal to seedling establishment, in the introduced range (Vilà et al. 2006; Vilà & D’Antonio 1998) and compare them to that of native species.

Several conifer species are invasive in many regions of the world, mainly in the Southern Hemisphere (Richardson & Rejmánek 2004). There has been a fair amount of research on seed and seedling traits concerning conifer invasiveness (Grotkopp et al. 2002; Richardson & Rejmánek 2004), but differences in the ecological barriers that constrain their establishment have been less explored with the exception of herbivory resistance to native fauna (Lombardero et al. 2008; Nuñez et al. 2008; Carrillo-Gavilán et al. 2010). Pseudotsuga menziesii Mirb. Franco, the Douglas fir, is considered one of the most invasive forestry conifer species of the world (Richardson & Rejmánek 2004). Douglas fir is reported as invasive in areas close to plantations in New Zealand (Kay 1994), South Africa (Richardson & Higgins 1998), Argentina and Chile (Simberloff et al. 2010). In Europe, it is naturalized in several countries (Carrillo-Gavilán & Vilà 2010). According to previous studies (Richardson & Bond 1991; Sarasola et al. 2006), the most suitable habitats to be invaded by this sort of alien conifers would be grasslands and shrublands followed by open forests and, finally, by closed forests. 85 Chapter 4

To assess how Douglas fir establishment is modulated by its interaction with the host community, we conducted field manipulation experiments during the early life-cycle stages (i.e., post-dispersal seed removal, seed survival, seed germination, seedling survival and growth) and compared them with a taxonomically close native conifer Abies alba (Silver fir) as a baseline species. Our main objectives were: a) to know which habitats, close to Douglas fir plantations, have the highest invasibility, b) to assess the main ecological barriers limiting plant establishment, c) to determine how different these patterns are compared to a taxonomically close, coexisting native conifer species. Given the current wide distribution of Douglas fir plantations in Europe, this study can contribute to habitat risk assessment to invasion when this species is introduced for afforestation.

MATERIALS AND METHODS

Study area

The study area was located in the Montseny Natural Park in Barcelona-Spain (latitude 41º42’-41º52’N, longitude 2º16’-2º33’E), a 40,000 ha mountainous area declared a UNESCO Biosphere Reserve in 1978. Montseny encompasses a wide climatic gradient and, consequently, a large phytogeographic range from a Mediterranean to a subalpine climate (Peñuelas & Boada 2003). At medium and high altitudes (800- 1700m a.s.l.), the mean annual rainfall is 1148 mm and mean annual temperature is 8.7ºC (means for the 2007/10 period, Can Lleonart meteorological station). This area hosts the most extensive southern European distribution areas of Fagus sylvatica (beech) forests and the western distribution of Calluna vulgaris heathlands (Bolòs & Vigo 1990). Also, Montane Quercus ilex (holm-oak) forests occur at the lowest altitudinal ranges (< 800m a.s.l.), with a mean annual rainfall of 943 mm and mean annual temperature of 11.7ºC (means for the 2007/10 period, Fontmartina meteorological station).

Experiments were conducted in the above-mentioned habitat types: beech forests, holm-oak forests, and heathlands (n = 4 sites/habitat)(Fig.4.1). We chose these habitat types because of their close proximity to Douglas fir plantations and Silver fir forests, and therefore their vulnerability to be invaded or colonized, respectively. Sites were 0.2-3 km apart. These habitats differ in vegetation structure and light availability and seed remover community (Torre & Arrizabalaga 2009). Diameter at breast height of trees was significantly bigger in beech than in holm-oak forests (16.7 ± 0.8 and 9.4 ± 0.36 cm, respectively; t-test: t=9.3, df =627, p < 0.0001). Heathlands

86 Establishment constraints of conifer in different habitats were dominated by shrubby shaped Juniperus communis, whereas beech and holm- oak forests had a low understory cover. Light availability (PAR, μmol s-1 m-2) at soil 2 surface was significantly different among habitats (Kruskal-Wallis test; X 2 = 118.7, p < 0.0001), being much higher in heathlands (928± 28) compared to holm-oak (51± 11) and beech forests (53.8 ± 13). The main seed remover in all three habitats is Apodemus sylvaticus, followed by A. flavicollis. Other less frequent seed removers are Myodes glareolus (beech and holm-oak forests), Mus spretus (holm-oak), Microtus agrestis (heathlands) and Glis glis (beech) in the Montseny (I. Torre, personal communication).

a b c

Fig. 4.1. Photographs from the study areas (a) beech forests, (b) holm-oak forests and (c) heathlands. Study species

Pseudotsuga menziesii var. menziesii (Mirb.) Franco (Pinaceae) (Douglas fir, hereafter) is a conifer of Northwestern American origin and an invasive species in many parts of the world where it has propagated from plantations for wood production (Richardson & Rejmánek 2004). The first published records of Douglas fir introductions in Europe were in the Czech Republic (1842), Germany (1900), Denmark and the U.K (1940) (Carrillo-Gavilán & Vilà 2010). In Montseny Natural Park, it was introduced during the 1950s and throughout the following decades, when marginal croplands were transformed to tree plantations of mainly fast growing alien conifers (Boada 2000). Douglas fir was one of the most planted species in Montseny with 230 scattered plantations (Boada & Broncano 2003) with a total area of only 250 ha. At the present, Douglas fir is naturalizing into adjacent heathlands but not into beech or holm-oak forests, which are also adjoined to plantations (Broncano et al. 2005). Therefore, observational surveys suggest that invasibility might be low in these forest types.

Abies alba P. Mill (Silver fir, hereafter) is a native conifer in Montseny where it represents the southernmost European distribution site for this species. We chose Silver fir as a pair-wise baseline species due to its close taxonomic relationship to Douglas fir, since both species belong to the Pinaceae family. For this reason, they have similar

87 Chapter 4 dispersal mechanisms, reproductive phenology and climatic requirements (López- González 2002). In Montseny, there are only three natural populations of Silver fir occupying a total of 28.5 ha surrounded by beech forests (Boada & Broncano 2003), where recruitment of seedlings is observed in tree fall gaps and areas with bare soil (Carrillo-Gavilán, personal observation).

To conduct our study, we purchased seeds from the Intersemillas nursery, Spain (http://www.intersemillas.es) because field collection attempts in previous years provided few viable seeds. However, this was of little concern since our focus was to compare the potential for seedling establishment between the two species under a range of different ecological conditions rather than documenting actual local rates of establishment.

For each species, 20 seeds were randomly chosen to measure fresh seed mass. Coat strength was also estimated following Rogerdson’s (1998) protocol with a Chatillon Universal Force Tester (Amtek/Chatillon, Largo, Florida, USA). The two species differ in seed mass (t-test: t=12.04, df =38, p < 0.0001) and seed coat thickness (t-test: t=12.18, df =38, p < 0.0001). Seed mass is larger in Silver fir (0.03 ± 0.001 g) than in Douglas fir (0.01 ± 0.0006 g). Seed coat strength is also greater in Silver fir (16.7 ± 0.97 g) than in Douglas fir (4.1 ± 0.33 g).

Seed removal

To test differences in post-dispersal seed removal (seed removal, hereafter) in Autumn 2008, we delimited a 100 x 100 m plot at each site where we placed five 50 m long transects. In each transect, we placed 20 seeds of each species randomly chosen from a pool of seeds. Each seed was glued to a piece of nylon fishing line, tied to a wire stake (Schupp 1988) and placed on the ground surface keeping a distance of approximately 3 m between each other. A total of 1200 seeds (50 seeds x 2 species x 3 habitats x 4 sites) were surveyed for seed removal after 2, 6, 13, 20 and 27 days. As conifer seeds area barely secondarily dispersed (Ordóñez & Retana 2004), we considered either missing seeds or the presence of seed coat remnants as evidence of seed removal. We finished the experiment when after two consecutive visits the seed removal curves did not vary. After that day all remaining seeds were removed from the field.

We conducted two analyses. First, we tested for differences in percentage of removed seeds at the end of the experiment among habitats and between species using a Generalized Mixed Lineal analysis on SAS macro GLIMMIX (Littell et al. 2006)

88 Establishment constraints of conifer in different habitats applying a binomial error and a logit link function. Habitat (n = 3), species (n = 2) and habitat x species interaction were considered as fixed effects. Site and site x habitat interaction were considered as random factors.

Second, Kaplan-Meier estimations of mean seed survival time were calculated, where differences among habitats and between species were compared with the Gehan-Wilcoxon test (Pyke & Thompson 1986). The software SPSS 13.0 was used for survival analysis (SPSS Inc., Chicago, Illinois, USA).

Seed germination

In November 2008, in each site we buried 10 metallic mesh bags per species at 2 cm depth. Each bag contained 10 seeds per species. Bags rested flat, so that all seeds were in direct contact with the soil. In total 2400 seeds were buried (10 seeds x 10 bags x 2 species x 3 habitats x 4 sites). In July 2009, 240 days after sowing, we retrieved the bags, and seeds were checked in the laboratory. They were considered to have germinated when the radicle and/or the cotyledon had emerged from the seed coat.

Differences in the percentage of germinated seeds in each bag among habitats and between species were analyzed using SAS macro GLIMMIX applying a binomial error and logit link function. Habitat (n = 3) and species (n = 2) were considered as fixed factors while site n( = 4) and site x habitat interaction were considered as random factors.

Seedling survival and growth

In February 2009, we transplanted one year old seedlings grown at a nursery (Centre de Jardineria Sils, www.jardineriasils.com, Girona, Spain) to each site, and thirty seedlings per species were randomly planted 3 m apart from each other along transects. In total, 720 seedlings were planted (30 seedlings x 2 species x 3 habitats x 4 sites) in this way. Seedlings were examined one month after being planted and those that had died from transplant shock were excluded from the experiment. Seedling survival was recorded every month until June 2010. Plant height was measured before summer drought (July 2009 and June 2010, i.e. five and sixteen months after transplanting, respectively) and after the first autumn rains (October 2009, i.e. eight months after transplanting).

Seedling survival time was calculated through Kaplan-Meier estimations and differences in mean survival time among habitats and between species were compared with the Gehan-Wilcoxon test (Pyke & Thompson 1986). 89 Chapter 4

Differences in seedling relative growth rate (i.e., relative increase in height of surviving seedlings between two consecutive monitoring times) were also analyzed with the macro GLIMMIX with the same fixed and random factors as mentioned in the seed germination analysis above, but applying a normal error and identity link function instead. Due to high seedling mortality in the holm-oak stands at the beginning of the experiment (see results), we performed comparisons of seedling growth between beech forests and heathlands at five months after transplantation, and only between species within beech forests at ten months and at the end of the experiment.

Probability of transition from seed to seedling

In order to link the different stages of the early life-cycle, we estimated the probability of a seed becoming an established seedling, for each habitat type, in both conifer species (i.e., probability of recruitment). The model assumed that the probability of a seed achieving recruitment can be estimated as the product of the previous elemental transition probabilities (Herrera et al. 1994; Vilà & D’Antonio 1998). We assumed that this model is independent of propagule pressure (i.e., seed production); the sequential recruitment stages considered were seed survival, seed germination and seedling survival.

RESULTS

Seed removal

Overall, seed removal was not very intense in all three habitats. At the end of the experiment, of the 1200 seeds deployed per species, only 20.5 % were removed. There were significant differences between species1, (F 9 = 6.56, P <

0.05) but not among habitats (F1, 6 = 0.84, P = 0.47). There was a significant habitat x species interaction (F2, 9 = 4.15, P < 0.05) indicating that Douglas fir seeds were removed more frequently than Silver fir seeds in holm-oak and heathlands, whereas in beech forests this difference was not significant (Fig. 4.2a).

2 There were differences in seed survival time among habitats (X 2 = 16.48, P < 0.0005) and between species (X2 = 9.19, P < 0.005), being significantly longer in beech forests, followed by holm-oak forests and heathlands. Also, seed survival time of Silver fir was longer than Douglas fir (Fig. 4.2b).

90 Establishment constraints of conifer in different habitats

Referringa) to differences within habitats, seed survivalAbies alba time was significantly shorter in Douglas fir than in Silver fir in heathlandsX ( 2 = 8.17,Pseudotsuga P < 0.005) menziesii and in holm-oak forests (X2 = 9.38, P < 0.005,) but not in beech forests (X2 = 1.07, P = 0.3) (Fig. 4.2b and 4.3).

a) 40 Abies alba * * Pseudotsuga menziesii

30

20

10 b) (%) removal Seed 10 Seed removal (%)

0 b) 30 ** ** 2525

20

1515

1010 Survival time (days) time Survival Survival time (days) 55 c) 0 c) 0 100 90 * * 80 70 60 50 40 30 Seed germination (%) Seed germination (%) germination Seed 20 10 0 BeechBeech Holm-oakHolm-oak Heatlands Habitat Fig. 4.2. a) Post-dispersal seed removal, b) seed survival time and c) seed germination (mean + s.e.) for Abies alba and Pseudotsuga menziesii in three habitat types. Asterisks indicate significant differences between species (* P <0.05, ** P <0.005) within habitats. 91 Chapter 4

Seed germination

There were significant differences in germination rates between species (F1, 8 = 19.05,

P < 0.005), yet none were found among habitats (F2, 6 = 0.22, P = 0.80). Also, habitat x species interaction was not significant 2,(F 8 = 0.07, P =0.93). In all habitats, the percentage of seed germination was significantly higher in Douglas fir than in Silver fir a) Beech (Fig. 4.2c). a) 100

80

60 AbiesAbies alba alba PseudotsugaPseudotsuga menziesii menziesii

40 Seed survival (%) survival Seed

Seed survival (%) 20

b) Holm-oak 0

b) 100

8080

6060 Abies alba alba Pseudotsuga menziesii menziesii 4040 Seed survival (%) survival Seed 2020 Seed survival (%)

c) Heathlands0 c) 100

80

60

40 AbiesAbies alba alba PseudotsugaPseudotsuga menziesii menziesii Seed survival (%) survival Seed 20 Seed survival (%)

0 0 5 1010 1515 2020 25 30 TimeTime (days) (days) Fig. 4.3. Seed survival curves (mean ± s.e.) of Abies alba and Pseudotsuga menziesii in a) beech forests, b) holm-oak forests and c) heathlands. 92 Establishment constraints of conifer in different habitats

Seedling survival and growth

Seedling survival was very low in holm-oak forests due to soil disturbance by native fauna. In this habitat, wild boars pull up more than 50 % of seedlings three months after transplantation, and by the end of the experiment all seedlings had died.

In heathlands and beech forests drought stress seemed to be the main barrier against seedling survival. In heathlands, only 4% of seedlings survived after the summer, while, in beech forests, seedling survival was less than 40% (Fig. 4.4).

Overall, at the end of the experiment, seedling survival time was significantly 2 different among habitats (X 2 = 319.8, P < 0.0001). Survival time was longer in beech forests and shorter in holm-oak forests (Fig. 4.5). There were differences between species (X2 = 25.4, P < 0.0001) with Silver fir seedlings having longer survival time than Douglas fir. Likewise, these differences between species were also observed within habitats (beech: X2 = 8.77, P < 0.005; holm-oak: X2 = 11.6, P < 0.0005; and heathlands: X2 = 6.05, P < 0.05) (Fig. 4.5).

93 a) Beech Chapter 4

a) 100100 AbiesAbies albaalba 9090 PseudotsugaPseudotsuga menziesii 8080 7070 6060 5050 4040 3030 Seedling survival (%) survival Seedling

Seedling survival (%) 2020200 10 b) Holm-oak 0

100100 b) Abies albaalba 90 Pseudotsuga menziesii menziesii 80 70 60 50 40 30 Seedling survival (%) survival Seedling 20 Seedling survival (%) 10 c) Heathlands 0

100100 AbiesAbies alba c) 90 PseudotsugaPseudotsuga menziesii 80 70 60 50 40 30 Seedling survival (%) survival Seedling 20 10 Seedling survival (%) 0 00 60 120 180180 240 300 360 420420 480480 TimeTime (days) (days) Fig. 4.4. Seedling survival curves of Abies alba and Pseudotsuga menziesii in a) beech forests, b) holm-oak forests and c) heathlands.

94 Establishment constraints of conifer in different habitats

Due to high seedling mortality, seedling relative growth rate could not be analyzed in holm-oak forests. Overall, five months after transplantation, relative growth rate was very low in both species (Douglas fir: 0.009 ± 0.01; and Silver fir:

0.022 ± 0.01 cm). There were no significant differences between species (F1, 468 = 0.75,

P = 0.38) or between beech forests and heathlands (F1, 6 = 0.4, P = 0.55). In beech forests, relative growth rate was not significantly different between species, both eight months after transplantation (F1, 234 = 0.68, P = 0.41) and at the end of the experiment

(F1, 237 = 2.36, P = 0.12).

AbiesAbies alba 300 PseudotsugaPseudotsuga menziesii menziesii **

250

200 * 150 **

100

Survival time (days) 100 Survival time (days) time Survival

50

00 Beech Holm-oakHolm-oak HeatlandsHeatlands HabitatHabitat

Fig. 4.5. Seedling survival time (mean + s.e.) of Abies alba and Pseudotsuga menziesii in three habitat types. Asterisks indicate significant differences between species (* P <0.05, ** P <0.005) within habitats.

95 Chapter 4

Transition from seed to seedling

Overall, the probability of one seed becoming an established seedling was very low for both conifer species (Fig. 4.6). In beech forests, approximately 4% of Douglas fir seeds, and 5% of Silver fir seeds of initial seed sets become established seedlings (Fig. 4.6a). In heathlands, Douglas fir showed a higher probability of survival than Silver fir (0.4% and 0%, respectively)a) Beech due to a higher germination rate, while in holm-oak forests, survival was null for both species. a) 1 AbiesAbies alba alba PseudotsugaPseudotsuga menziesii 0,80.8 menziesii

0,60.6

0,40.4

Probability of survival Probability 0,20.2 b) Holm-oak Probability of survival

0 b) 1 AbiesAbies alba alba PseudotsugaPseudotsuga menziesii 0,80.8 menziesii

0,60.6

0,40.4

Probability of survival Probability 0,20.2

c)Probability of survival Heathlands

0 c) 1 1 AbiesAbies alba alba Pseudotsuga menziesii 0,80.8 Pseudotsuga menziesii 0,60.6

0,40.4

0,20.2 Probability of survival Probability

Probability of survival 0 Post-Post SeedSeed Seed SeedlingSeedling dispersaldispersal survivalsurvival Germinationgermination survivalsurvival

RecruitmentRecruitment stages stages Fig 4.6. Probability of survival along each establishment stage of Abies alba and Pseudotsuga menziesii in a) beech forests, b) holm-oak forests and c) heathlands. 96 Establishment constraints of conifer in different habitats

DISCUSSION

The probability of invasion success of Douglas fir decreased along the sequential stages of establishment in all habitats, showing the best performance in beech forests. The higher germination capacity of Douglas fir seemed to predict a better establishment success compared to Silver fir. However, pressure by native fauna and drought stress acted as the main ecological filters limiting seedling establishment of both conifer species.

In Douglas fir, seed removal was not very intense in any habitat compared to its native range (Huggard & Arsenault 2009) or with other conifer species (Borchert et al. 2003; Peters et al. 2004; Carrillo-Gavilán et al. 2010). Seed removal was similar among habitats despite of the fact that the three habitat types analyzed were located at different altitudes. These results do not match with previous studies, which have observed that small mammal abundance decreases with altitudinal gradients (Torre & Arrizabalaga 2009). The low seed removal rates might be due to seed satiation (Stowe et al. 2000) as the experiment coincided with seed dispersal of many native plant species. Moreover, seed removers might have a higher preference either towards holm-oak acorns (Gómez 2004; Espelta et al. 2009) or beech nuts, which are larger than fir seeds.

Douglas fir germination rates were high and did not differ between habitats despite large differences in vegetation structure and light availability. This is in accordance with other field studies where Douglas fir seedling emergence was very homogeneous under different microsites (Dunne & Parker 1999). In spite of this, our seed germination rates should be considered with caution. Since they come from a nursery, our seed sample might be causing an overestimation of this parameter attributable to a higher quality of seeds. Nonetheless, seed germination was higher in Douglas than in Silver fir. Silver fir’s sensitivity to other environmental stresses such as drought (Gradecki & Poštenjak 2001), might be the cause of its low germination rates. In this context, a recent review by Pyšek & Richardson (2007) reveals that alien species are generally reported to germinate earlier, better and in a wider range of conditions than native species.

Since many seedlings disappeared in holm-oak forests due to soil disturbance by wild boars, native fauna seemed to be an important factor controlling the establishment of seedlings in this forest type. In contrast, in beech forests and in heathlands, seedling mortality due to summer drought stress (Alpert et al. 2000) was a prominent ecological

97 filter for both native and alien plant species as has been reported in other Mediterranean regions (Dunne & Parker 1999; Rey & Alcántara 2000; Domènech & Vilà 2006).

The higher seedling survival in beech forests compared to other habitats might suggest a better establishment success of Douglas fir in this habitat. However, Douglas fir is considered a pioneer species that prospers in partial shade, but once established it requires strong light availability (Hermann & Lavender 1999). As with most alien conifers, non-disturbed forests are resistant communities to invasion (Richardson & Bond 1991; Sarasola et al. 2006). Simberloff et al. (2002) observed that Douglas fir did not establish itself into closed forest plantations. Therefore, we would expect invasibility to decrease in beech forests in subsequent stages (Caccia & Ballaré 1998), except in natural gaps (Spies & Franklin 1989) or if forests are intensively managed. If disturbance occurs and resources fluctuate (Davis et al. 2000) this species might naturalize, even if currently there is no evidence of seedling establishment in this forest type.

Conifer traits such as a short juvenile period, a short interval between large seed crops, a small seed mass (Rejmánek & Richardson 1996; Richardson & Rejmánek 2004) and fast relative growth rate (Grotkopp et al. 2002) are correlated with high invasion success. Differences in seedling survival time between conifer species might be explained by the seed-seedling conflict (Shupp 1995), for which small seeded species have better dispersal capacity but low seedling survival due to lower investment in reserves than large seeded species. Small seeded species such as Douglas fir tend to produce large seed crops to counterbalance the low survival probability from seed to adulthood, which would be expect due to small seed mass (Moles & Westoby 2004). According to Greene & Johnson (1999), seed production in trees is inversely proportional to seed mass. In Douglas fir, data describing seed production in U.K. reveal that good crops are produced every 4-6 years (Wilan 1985). Hermann & Lavender (1999) estimate that seed production is about 2.2 kg/ ha, which varies widely among years; and, Gashwiler (1967) estimates annual mean seed dispersion around 3x105 seeds/ha of Douglas fir in the northwest of the United States. While seed production of Silver fir seems to be lower than of Douglas fir,it highly varies among years. For example, seed production strongly varied in two consecutive years from 59.6 to 118.8 seeds/m2 in the French Alps (Sagnard et al. 2007).

98 Establishment constraints of conifer in different habitats

To sum up, soil disturbance by native fauna in holm-oak forests and drought stress in heathlands were the most i mportant factors controlling seedling establishment. Douglas fir performed better in beech forests because seed survival, seed germination and seedling survival were higher than in the other habitats. In Douglas fir, higher seed production and seed germination rates compared to the native Silver fir, are the only early-life cycle traits that appear to be different. Currently the only invaded habitat is the heathlands (Broncano et al. 2005). This indicates that there is a mismatch between the invasibility observed in previous studies and the degree of invasion suggested by our results. This mismatch might predict that Douglas fir establishment observed in heathlands is a consequence of processes and/or mechanisms that act at mid life-cycle stages from seedling to adulthood where establishment success is favored. Therefore, prevention of Douglas fir establishment should focus in reducing propagule pressure, independently of habitat type at early life-cycle stages. Yet, we highlight the need for future research at mid life-cycle stages to determine what are the main drivers of invasiveness of Douglas fir and the invasiblity of different habitats.

Acknowledgments

We are very grateful to E Pérez-García, J Andreu, V Meco, and M Hernanz for field assistance, and to all personnel at Fontmartina and Can Casades in the Montseny Natural Park for their assistance during field work. Research was partially funded by the Spanish Ministerio de Ciencia e Innovación projects: Consolider-Ingenio MONTES (CSD2008-00040), COMPROPIN (AGL2010-18724), and CSIC (project 200840I153).

99 100 Chapter 5

Early resistance of alien and natives pines against two native generalist insect herbivores: no support for the Natural Enemy Hypothesis4

4Carrillo-Gavilán MA, Moreira X, Zas R, Vilà M, Sampedro L. In review in Functional Ecology.

101 Chapter 5

Abstract

The Natural Enemy Hypothesis (NEH) predicts that alien plant species might receive less pressure from natural enemies than do related coexisting native plants. However, most studies to date are based on pairs of native and alien species and the results remain inconclusive. The level of attack by native generalist herbivores can vary considerably between plant species, depending on defensive traits and strategies. Plant defenses include preformed constitutive and induced defenses that are activated as plastic responses to herbivore attack. However, the efficacy of induced defenses could be altered when alien species entering an area are exposed to new enemies. We tested the NEH for several closely related alien and native pines to Europe by examining early anti-herbivore resistance to damage by two generalist native insect herbivores (Hylobius abietis and Thaumetopoea pityocampa); the differences in constitutive and inducible chemical defenses (i.e., diterpenes in stems, total phenolics, and condensed tannins in needles); and whether consumption preferences shift after defenses have been induced. We did not find alien to be less damaged than native pines. Constitutive concentration of total phenolics and condensed tannins differed between the alien and native pines but diterpenes did not. Concentrations of diterpenes in the stem increased after herbivory by H. abietis in four pines but without significant differences between alien and native pine species. We found no significant quantitative changes in chemical defenses after herbivory by T. pytiocampa. In cafeteria bioassays, H. abietis consumed more the twigs from alien than those from native species irrespective to previous exposition to the insect. T. pityocampa preferred needles from native pine but only when the plants were previously exposed to this insect. Overall, we found no support for the NEH in alien pines to Europe. This suggests that alien pines, in regions where they coexist with native congeners, may be controlled by native generalist herbivores, this being one reason that invasion by alien pines is not frequent in Europe.

102 Resistance alien and natives pines to herbivores

INTRODUCTION

The success of alien plants in their new range depends partially on the interaction with the native biota, particularly with herbivores that could damage establishing seedlings (Maron & Vilà 2001; Levine et al. 2004). The impact of native herbivores on alien plants is determined by the interplay between their ability to recognize the alien plant as a food resource (Carpenter & Cappuccino 2005), the effectiveness of the alien plant defenses against herbivores in the new range (Stowe et al. 2000) and the ability of herbivores to overcome plant defenses of the alien (Rausher 2001). All these mechanisms might influence the success or failure of an alien plant to invade a new area.

The Natural Enemies Hypothesis (NEH) predicts that alien plant species are successful invaders in the new range due to a lack of natural enemies, which were left behind in their native range (Maron & Vilà 2001). Several studies have shown how introduced plant populations in a new area are less damaged by native phytophages and pathogens than are populations in the native range (DeWalt et al. 2004; Vilà, Maron & Marco 2004; Rogers & Siemann 2005). The NEH also contends that aliens undergo less pressure from natural enemies than do related native plants coexisting in the area. Nevertheless, there are contradictory results on whether native herbivores cause more damage to native than to alien plant species (Chun et al. 2010). The level of damage inflicted against alien species and its influence on plant fitness is not always lower than against coexisting native plants, even when the species are taxonomically related (Agrawal & Kotanen 2003). A recent meta-analysis comparing pair-wise native and alien trees noted that the majority of native insect species that colonize alien trees are generalist herbivores (Bertheau et al. 2010). In addition, reduced resistance of alien plants leading to rapid colonization by native generalist herbivores was found to occur when congener native plant species live in the introduced range (e.g. Goßner et al. 2009). However, most studies have compared alien-native species pairs, and, as suggested by Chun et al. (2010), more than two species should be considered for testing related alien and native species.

The level of attack by herbivores can vary considerably between plant species, depending on their defensive strategies (Agrawal et al. 2005). Plant chemical defenses include constitutive defenses, which are permanently expressed irrespective of the plant exposure to natural enemies, and induced defenses enhance the basal defense capacity as a result of plastic responses to natural-enemy attack (Karban & Myers 1989). Induced strategies, rather than constitutive, are expected to be favored when herbivore

103 Chapter 5 pressure varies spatially and temporally, and initial attacks are reliable cues of further attacks (Karban 2010). Closely related species may differ in their defensive strategy, depending on the biotic and abiotic environment where they have evolved (Orians & Ward 2010). Production of effective induced defenses requires the plant to quickly recognize the herbivore damage as a potential risk, signaling the danger across tissues and subsequently triggering defenses (Heil 2010). There is emerging evidence that induced responses to herbivory can show high specificity regarding both the damaging herbivore (i.e., specificity of elicitation) and the efficacy of the induced responses (i.e., specificity of the effect) (Agrawal 2010). The efficacy of the induced defensive strategies could thus be notably altered when alien plants are exposed to new enemies in the colonized area. Native plants due to a common evolutionary history with the native generalist enemies have evolved to a greater inducible defense against them than alien plants (Joshi & Vrieling 2005). Recognition mismatch by the alien plant, for example, could hinder induced responses and, consequently, reduce the fitness of the alien plant. Although constitutive as well as induced defenses can contribute to effective resistance to herbivory, the interplay between the two defensive strategies has been rarely considered when explaining invasiveness of alien species (Orians & Ward 2010).

Pines (Pinus spp.) constitute a classic model in the study of traits associated with plant-invasion potential (i.e., invasiveness) (Richardson 2006). Pine invasiveness has been associated with small seed mass, short juvenile period, short interval between large seed crops and fast relative early growth rate (Rejmànek & Richardson 1996; Grotkopp et al. 2002). However, there is pronounced geographical variation in invasion success within species. For example, the North American Pinus radiata is invasive in many regions of the Southern Hemisphere (Lavery & Mead 1998), while in Europe little evidence of invasion has been reported (Carrillo-Gavilán & Vilà 2010). Differences in invasiveness among biogeographical regions have been suggested to be related to different propagule pressure (Pyšek & Jarošik 2005) but might also depend on differences in the interaction with natural enemies (Richardson & Higgins 1998). For example, in Europe, P. radiata is attacked by many generalist pests that also damage native pines species (Lombardero et al. 2008), while in the Southern Hemisphere native pines —and hence their associated specialist and generalist herbivores— are absent. In addition, as survival at early stages is key for future fitness of pine trees, early resistance of pine seedlings is a likely determinant for invasiveness.

In this study, we investigated the early resistance to generalist herbivores in seedlings of alien and native pine species to Europe within the context of the predictions of the NEH. The specific objectives were: 1) to compare the damage caused by two

104 Resistance alien and natives pines to herbivores generalist native insect herbivores in alien and native pine species, 2) to evaluate quantitative differences between alien and native pine species in constitutive chemical defenses and in those expressed after exposition to herbivory by those insects, and 3) to assess whether feeding preferences on alien and native pines shift after defenses have been induced. We expected that: 1) generalist herbivores would cause lower damage to alien than to native pine species as predicted by the NEH; 2) alien plants exposed to generalist native herbivores would induce weaker defenses than would native pines because a lack of common evolutionary history; and finally 3) herbivory by native herbivores would decrease further herbivore consumption more in native than in alien pines.

MATERIALS AND METHODS

To test the above mentioned objectives, we measured the whole-plant damage caused by two insect generalist herbivores to living seedlings; evaluated the differences in constitutive and inducible chemical defenses after the treatments of exposition to herbivores, and assessed through cafeteria bioassays whether feeding preferences of both insects shifted after plant defenses have been induced.

Study species

Nine pine species belonging to the Pinus clade (Eckert & Hall 2006), which are broadly planted worldwide, were used for this study: Pinus canariensis C.Sm. (CAN), P. halepensis Mill. (HAL), P. pinea L. (PIN), P. pinaster Ait. (PTR), and P. sylvestris L. (SYL) are native to Europe; while P. coulteri D. Don (COU), P. radiata D. Don (RAD), P. roxburghii Sarg. (ROX), and P. sabiniana Douglas ex D.Don (SAB) are alien to Europe (DAISIE 2009). Seeds were purchased in the following nurseries: Intersemillas, Spain (http://www.intersemillas.es); Sheffield’s Seed Co., USA (http://www.sheffields.com); and Les Semences Du Puy, France (http://www.semencesdupuy.com ).

The assayed herbivores were the large pine weevil, Hylobius abietis L. (Coleoptera: Curculionidae) and the pine processionary caterpillar, Thaumetopoea pityocampa Dennis & Schiff (Lepidoptera: Thaumetopoeidae) (Hylobius and Thaumetopoea respectively, hereafter). These two insect species are naturally distributed in Europe and eastern Asia, where they feed on several pine species. Adults of Hylobius feed on the phloem and bark of young conifer seedlings, leading to high seedling mortality and major growth losses. This weevil is the most harmful pest of conifer plantations in Central and Northern Europe (Nordlander et al. 2003).

105 Chapter 5

Caterpillars of Thaumetopoea intensively defoliate several pine species in the Mediterranean basin, causing serious economic and ecological impact (Hódar et al. 2002). Adult weevils were trapped in the field following the method described by Moreira et al. (2008a), stored in culture chambers at 15ºC and fed fresh twigs for maximum two weeks before the experiment begun. Entire Thaumetopoea caterpillar nests were collected directly from infested trees, labeled, and maintained as above.

Greenhouse experimental design

We conducted a two-factorial greenhouse experiment with pine species and induction of plant defensive responses by exposing living seedlings to two insect herbivores as the main factors (herbivore-induction treatments, hereafter). The experiment followed a randomized split-plot design replicated in 10 blocks, with herbivore-induction (two levels: control and herbivory by Hylobius and by Thaumetopoea) as the whole factor and pine species (9 levels) as the split factor. However, due to lack of enough pine seedlings, only six pine species (CAN, PTR, SYL, COU, RAD, and ROX ) were induced with Thaumetopoea. In total, we grew 240 pine seedlings, corresponding to 10 blocks x 2 treatments (control and herbivory by Hylobius) x 9 species, plus 10 blocks x 1 treatment (herbivory by Thaumetopoea) x 6 species.

In October 2008, at the Forestry Research Centre of Lourizán (Pontevedra, Spain), pine seeds were individually sown in 2-l pots filled with a mixture of perlite and peat (1:1 v:v), fertilized with 12 g of a slow-release fertilizer (Multicote® N:P:K 15:15:15), and covered with a 1-2 cm layer of sterilized sand (Fig. 5.1). Pots were placed in a greenhouse with controlled light (12 h light) and temperature (10 to 25ºC at night and day, respectively) and daily irrigation (Fig. 5.1). Plants were supplemented monthly with a solution of micronutrients to avoid nutritional deficiencies.

a b

Fig. 5.1. Photographs of (a) emergence and (b) growth of pine seedlings at the greenhouse in Lourizán (Pontevedra, Spain). 106 Resistance alien and natives pines to herbivores

Before the herbivore-induction treatment began, weevils were kept individually without food for 48 h in labeled Petri dishes with a moist filter paper (15ºC, dark) and then weighed. Similarly, Thaumetopoea nests were carefully opened, 2nd-instar larvae randomly separated into groups of 10 caterpillars, and starved and weighed as above. In September 2009, we applied the herbivore-induction treatments (herbivory by Hylobius and by Thaumetopoea). In the herbivory by Hylobius, one pre-weighed weevil was placed on each pine seedling, allowed to feed for 5 days and then removed and weighed again. Damage inflicted by the weevil was evaluated independently in every 1/5 stem sections as the relative debarked area using a four- level scale (0 = undamaged; 1 = 1-25% damaged; 2 = 26-50% damaged; 3 = >50% damaged), and the sum of values for the 5 sections per seedling (i.e., 0-15 score) was considered to be the debarked area (Fig. 5.2a). For the herbivory by Thaumetopoea, as caterpillars are gregarious and move little from the place where they are deposited, we placed one pre-weighed group of 10 caterpillars in the top section and other in the bottom section of each living pine seedling. Due to the smaller size of P. sylvestris seedlings, only one group of 10 larvae was placed on this species. Caterpillars were allowed to feed on the needles for 6 days, and then removed, counted and weighed (Fig. 5.2b). Foliar damage was evaluated for the whole plant in a three-level scale: 0 = undamaged leaves, 1 = damaged leaves < 5, 2 = damaged leaves > 5 (i.e., 0-2 score). Each plant of the three herbivore-induction treatments was carefully covered with a nylon mesh to avoid either herbivore escape or interference among treatments.

a b

Fig. 5.2. Details of the assayed herbivores: (a) Hylobius abietis and (b) Thaumetopoea pytiocampa during the herbivore-induction treatment.

107 Chapter 5

One week after the application of herbivore-induction treatments, all pine seedlings were harvested and directly weighed to determine the total aboveground plant biomass. Just after harvesting, two basal 2.5 cm long stem sections, and two subsamples (1.5 g each) of needles randomly from the whole pool of needles of each plant were collected, labeled and immediately stored in ice-coolers to be used in the cafeteria bioassays.

Chemical analyses

Concentration of stem diterpenes was estimated gravimetrically following the procedure proposed by Sampedro et al. (2011b), and expressed as mg of non-volatile resin g-1 stem dry weight (d.w.). A 5-cm-long section of the low part of the stem of each plant was collected, weighed, immediately frozen, and stored at -30ºC. Briefly, stem material was cut in small sections (ca. 5 g fresh weight) and resin compounds were quantitatively extracted twice with n-hexane in an ultrasonic bath, after which the plant material was recovered by filtration (Whatman GFF), the solvent in the tubes was evaporated to dryness, and the mass of the non-volatile resin residue was determined gravimetrically with a precision scale. This gravimetric determination of non-volatile resin was well correlated with the concentration of the diterpenoids fraction as quantified by gas chromatography in previous trials (r = 0.921, P < 0.0001; Sampedro et al. 2010).

Total phenolic and condensed tannins in the leaves were extracted and analyzed as described by Sampedro et al. (2011b). A subsample of needles (approximately 2 g dried weight) was also weighed, oven-dried (45ºC to constant weight) and then manually ground in a mortar with liquid nitrogen. Briefly, needles were extracted with aqueous methanol (1:1 vol:vol) in an ultrasonic bath, and total phenolics in the methanolic extract were analyzed by the Folin-Ciocalteu method, with colorimetric determination in a microplate reader at 740 nm using tannic acid as standard. Then the methanol extract was mixed with acidified butanol and a ferric ammonium sulfate solution, while condensed tannins analyzed by the protocyanidine method, determined in a microplate reader at 550 nm against purified condensed tannins of the tree quebracho (Schinopsis balansae Engl.; Droguería Moderna, Vigo, Spain). The concentration of total phenolics and condensed tannins is thus expressed as mg of tannic acid equivalents and mg of quebracho condensed -tannin equivalents g-1 needle dried weight, respectively.

108 Resistance alien and natives pines to herbivores

Cafeteria bioassays

Independent cafeteria bioassays were conducted with Hylobius and with Thaumetopoea in six pine species: CAN, PTR, SYL, COU, RAD and ROX. In the Hylobius bioassay one weevil (starved and weighed as explained above) was allowed to feed on a 2.5-cm-long section of basal stem for 48 h in a Petri dish with a moist filter paper (dark, 18 ºC). The volume of tissue ingested by the weevil (mm3) was estimated after measuring the debarked area by the weevil with a millimeter grid, and the depth of the gnawing with an electronic gauge. We conducted two replicates per plant, except for P. sylvestris, for which only one replicate was made due to insufficient plant material.

In the Thaumetopoea bioassay, groups of ten 2nd- or 3rd-instar caterpillars (starved and weighed as above) were allowed to feed on fresh needles for 7 days in Petri dishes as described above. A parallel series of needles was immediately dried at 65ºC to constant weight to determine percentage of dry matter. The specific ingestion of needles (g needles d.w.) by the caterpillars was calculated as the dry food ingested. The percentage of caterpillar survival at the end of bioassays was also registered as an indicator of plant defenses. As in the Hylobius bioassay, two replicates were performed per seedling except for P. sylvestris, for which we included only one replicate. In total 330 Petri dishes were used in the bioassay (5 species x 3 treatments x 10 plants x 2 replicates, and 1 species (SYL) with 3 treatments x 10 plants x 1 replicate).

Statistical analyses

Differences among pine species in whole-plant damage to seedlings , measured as debarked area by Hylobius and foliar damage by Thaumetopoea, were analyzed in treated plants using a generalized linear model: Yjk = μ + SPj + Bk +Ejk, where μ is the general mean, SPj and Bk are the main effects of the species j (SPj = 1-9 and SPj = 1-6 in herbivory for Hylobius and by Thaumetopoea treatments, respectively) and block k (Bk = 1–10), respectively. This analysis was performed with the PROC-GLIMMIX procedure of the SAS System (Littell et al. 2006) with normal error and an identity link function. While, differences among pine species in constitutive chemical defenses were analyzed in control untreated plants using the same equation model as before but using PROC-MIXED procedure of the SAS System.

The effects of herbivore-induction treatments and pine species on chemical defenses were analyzed with the following mixed model: Cijk = μ + Ti+ SPj + Bk + Ti*SPj

+Ti*Bk + Eijk, where μ is the overall mean, Ti and SPj are the main effects of induction

109 Chapter 5 treatments i (Ti= 0, 1, 2), species j (SPj = 1-9 and SPj = 1-6 in herbivory for Hylobius and by Thaumetopoea treatments, respectively), TixSPj and BkxTi are the corresponding interactions and Eijk is the experimental error. While block k (Bk = 1 –10) and the BkxTi interaction were considered random factors in order to analyze the main factor T with the appropriate error terms (Littell et al. 2006). Because of the different number of pine species included in the Hylobius and Thaumetopoea herbivore-induction treatments, the effects of induction were analyzed independently for each treatment together with the control. Plant biomass was previously checked not to covariate with the response variables and thus was not included in the model.

The cafeteria bioassays were analyzed with an hierarchical mixed model in which we included the BkxTixSPj interaction as a random effect (split-split design), in order to account for the autocorrelation of the two Petri dishes used for each pine seedling (Littell et al. 2006), and to analyze the species factor with the appropriate degrees of freedom. Insect weight was included as a covariate in the analyses. Relationships between damage in cafeteria bioassays and chemical defenses were carried out across pine species (n = 6).

When there were significant differences between species, we analyzed whether native pine species differed from alien species by including the factor Range (alien or native) and nesting the factor Species within the factor Range in the mixed models. All analyses were performed with the PROC-MIXED procedure of the SAS System. When needed, normality was achieved by log-transforming the original variables. Equality of residual variance across treatments was tested in all cases, and residual heterogeneity variance models were used when significant deviations were found (Littellet al. 2006). Data are shown as least square means ± standard error of the mean (s.e.m.). When main effects or differences between species were significant, pair-wise differences were tested for significance using the LSMEANS statement.

RESULTS

Seedling damage

Seedling damage by Hylobius and Thaumetopoea after herbivore-induction treatments significantly differed among pine species (Hylobius: F8, 70 = 2.66, P < 0.05;

Thaumetopoea: F5, 45 = 3.4, P < 0.05) (Fig. 5.3). The relative resistance of the pine species to both herbivore species was not consistent. For example, the least damaged pine by Thaumetopoea was P. canariensis, while stems of this pine species showed

110 Resistance alien and natives pines to herbivores intermediate resistance to Hylobius (Fig. 5.3). Additionally, herbivore damage did not significantly differ between alien and native pine species (Hylobius: F1, 70 = 0.19,

P = 0.66; Thaumetopoea: F1, 45 = 1.19, P = 0.28). a) 7 7 a

66 ab

55 abc abc bcd bcd a cd a 44 cd

33 d

22 Debarkedarea (0-15 score)

Debarkerd area (0-15 score) Debarkerd area (0-15 11

00 CAN HAL PIN PTR SYL COU RAD ROX SAB native alien CAN HAL PIN PTR SYL COU RAD ROX SAB native alien b) 33

2.52,5 a ab 22 ab ac a bc a 1.51,5 c

11 Foliar damageFoliar (0-3 score)

0.50,5 Foliar damage (0-3 score)

00 CAN HAL PIN PTR SYL COU RAD ROX SAB native alien CAN HAL PIN PTR SYL COU RAD ROX SAB native alien Native Alien

Fig. 5.3. Seedling damage in European native (black bars) and alien (gray bars) pine species to (a) the pine weevil Hylobius abietis and (b) the pine processionary moth Thaumetopoea pytiocampa. Bars are LS means ± s.e.m. Different letters above columns indicate significant differences (P <0.05) among pine species or between Range (native vs. alien). Species abbreviations: Pinus canariensis (CAN), P. halepensis (HAL), P. pinea (PIN), P. pinaster (PTR) and P. sylvestris (SYL) and P. coulteri (COU).

111 Chapter 5

Constitutive and induced chemical defenses

Constitutive concentration of chemical defenses in undamaged seedlings varied significantly among pine species (diterpenes: F8, 71 = 18.88, P < 0.001; total phenolics:

F8, 72 =2.17, P = 0.04; condensed tannins: F8, 70 = 7.46, P < 0.001; Fig. 5.4). P. sabiniana and P. sylvestris were the species with the highest concentration of diterpenes in the stem, while P. radiata and P. roxburghii showed the highest concentration both of total phenolics and of condensed tannins in the needles. Alien pine species registered significantly greater constitutive concentrations of total phenolics and condensed tannins in the needles than did native pine species (total phenolics F1, 72 = 6.22, P <

0.05; condensed tannins F1, 70 = 4.80, P < 0.05, Figs. 5.4b and c). However, alien and native pines did not differ in their constitutive concentration of diterpenes in the stem

(F1, 71 = 1.79, P = 0.18).

After the exposure to the herbivore-induction treatments, the diterpenes concentration in the stem was significantly affected by Hylobius herbivory (Table 5.1a). Overall, the concentration of diterpenes in seedlings exposed to Hylobius was 25% greater than those in undamaged control seedlings (Fig. 5.4a). Moreover, the induction of stem diterpenes after herbivory by Hylobius differed among the pine species, as indicated by the significant species x treatment interaction (Table 5.1a, Fig. 5.4a).

112 Resistance alien and natives pines to herbivores

a) 80 Control Control Hylobius 70 Hylobius Thaumetopoea Thaumetopoea 60 * HYL THA 50 d.w) *** R ns. ns. -1 40 * *

30 (mg g

20 Diterpenes in the stem Diterpenes in the stem (mg g-1 d.w.) g-1 (mg stem the in Diterpenes

10

0 CANCAN HALHAL PINPIN PTR SYL COUCOU RAD ROXROX SABSAB nativenative alienalien b) Control 40 Control Hylobius 35 Hylobius d.w) Thaumetopoea -1 Thaumetopoea 30 HYL THA

25 R * *

20

15

Leaf total phenolics 10

(mg tannic acid equiv g 5 Leaf total phenolics (mg tannic acid equiv. g-1 d.w.) g-1 equiv. acid tannic (mg phenolics total Leaf

0 CANCAN HAL PINPIN PTRPTR SYL COUCOU RAD ROXROX SABSAB nativenative alienalien

c) Control 140 Hylobius Control

d.w) Thaumetopoea -1 120 Hylobius Thaumetopoea 100 HYL THA R ns. *

80

60

40 Leaf condensed tannins 20

(mg quebracho tannic equiv g 0 CAN HAL PIN PTR SYL COU RAD ROX SAB native alien Leaf condensed tannins (mg quebracho tannins equiv. g-1 d.w.) g-1 quebrachoequiv. condensed tannins (mg Leaf tannins CAN HAL PIN PTR SYL COU RAD ROX SAB native alien Native Alien Fig. 5.4. Effects of the herbivore-induction treatments (control plants; plants exposed to herbivory by Hylobius abietis [HYL]; plants exposed to herbivory by Thaumetopoea pytiocampa [THA] on the concentration of (a) diterpenes in the stem, (b) total phenolics and (c) condensed tannins in the needles. The species are grouped into native and alien species in Europe. Bars are LS means ± s.e.m. Asterisks indicate significant differences (* P<0.05, ** P <0.005, *** P <0.0001 and ns.= non significant) only between herbivory by Hylobius treatment with respect to control treatment in each species on the left side of solid line or between Range (R, native vs. alien) on the right side of solid line. See Fig. 5.3 legend for species abbreviations. 113 Chapter 5

Table 5.1. Results of the mixed models with species (SP) and herbivore-induction treatment (T) as fixed factors and concentration of diterpenes in the stem and of total phenolics and condensed tannins in the needles as dependent variables. Analyses were performed independently to compare control plants and a) pine seedlings induced with the large pine weevil Hylobius abietis and, b) pine seedlings induced with caterpillars of the pine processionary moth Thaumetopoea pityocampa. a. Nine pine species in control and after herbivore-induction by Hylobius Diterpenes Total Phenolics Condensed tannins Source df F P df F P df F P Block 9, 9 0.43 0.885 9, 9 1.76 0.206 9, 9 4.63 0.016 Treatment (T) 1, 9 12.55 <.005 1, 9 1.13 0.315 1, 9 0.74 0.411 Species (SP) 8, 137 13.33 <.0001 8, 139 3.55 <.0001 8, 137 11.43 <.0001 T*SP 8, 137 2.46 < 0.05 8, 139 1.13 0.346 8, 137 1.1 0.365 b. Six pine species in control and after herbivore-induction by Thaumetopoea Diterpenes Total Phenolics Condensed tannins Source df F P df F P df F P Block 9, 9 2.68 0.078 9, 9 2.66 0.080 9, 9 5.07 0.011 Treatment (T) 1, 9 0.11 0.75 1, 9 0.31 0.590 1, 9 0.1 0.764 Species (SP) 5, 84 32.01 <.0001 5, 85 2.28 < 0.05 5, 84 5.49 <.005 T*SP 5, 84 1.25 0.294 5, 85 0.52 0.759 5, 84 0.82 0.539

Diterpene concentrations in the stem of P. pinaster, P. sylvestris, P. coulteri, and P. roxburghii significantly increased after exposure to the weevil, while no significant changes were detected in the other pine species (Fig. 5.4a).We found no differences between alien and native pine species (F1, 137 = 1.21, P = 0.27). Diterpenes concentrations in the stem were not affected by Thaumetopoea (Table 5.1b). The concentrations of total phenolics and condensed tannins in the needles were not affected by either the Hylobius or Thaumetopoea treatments (Tables 5.1a and b; Figs. 5.4b and c).

114 Resistance alien and natives pines to herbivores

Consumption preferences in the cafeteria bioassays

Pine species differed significantly in the consumed volume of stem by Hylobius in the cafeteria bioassays (Table 5.2a). Pinus sylvestris was the least damaged while P. coulteri and P. canariensis were the most (Fig. 5.5a). We found a marginally significant negative relationship (at the species level) between the amount of tissue consumed by Hylobius in the cafeteria bioassays and the concentration of diterpenes in the stem (r2 = 0.46, P = 0.06, n = 6). Consumption by Hylobius was not significantly affected by the herbivore-induction treatment (Table 5.2a). Specifically, stems from seedlings previously exposed to herbivory by Hylobius were consumed as much as were those from control seedlings (Fig. 5.5a).However, alien and native pine species significantly differed in consumption (F1, 89 = 7.69, P < 0.005), with alien pines on average were consumed more than native pine species.

115 Chapter 5

3535 Control a) Control a Hylobius 3030 Hylobius ab )

3 2525 b a 2020 b b b

1515

of stem (mm c 1010 Consummed volume Consumed volume of stem (mm3) (mm3) stem Consumed of volume

55

00 CANCAN PTRPTR SYLSYL COUCOU RADRAD ROXROX nativenative alienalien b) Control 0.20,2 a a Thaumetopoea Control 0,18 Thaumetopoea 0.160,16 a 0,14 a a

0.120,12 b b b 0,1

0.080,08

0,06 (g needles d.w.)

0.040,04 Specific ingestion of needles (g needles d.w.) d.w.) needles (g needles of ingestion Specific

Specific ingestion ofr needles 0,02

00 c) CANCAN PTR SYL COUCOU RAD ROXROX nativenative alienalien 100100 Control Control 9090 a a Thaumetopoea Thaumetopoea 8080 b a b a 7070 bc 6060 c 5050

4040 Caterpillar survival (%) survival Caterpillar 3030

Caterpillar survival (%) 2020

1010

00 CANCAN PTRPTR SYLSYL COUCOU RADRAD ROXROX nativenative alienalien Native Alien Fig. 5.5. Results of the cafeteria bioassays showing the feeding activity of Hylobius abietis (a), and feeding activity (b) and performance (c) of Thaumetopoea pityocampa when feeding on pine tissues from control seedlings (white bars) and seedlings previously induced by these insects (black bars). Pine species are grouped from native to alien species in Europe. Bars are LS means ± s.e.m. Different letters indicate significant differences (P<0.05) between induced- plant with respect to control plants or between Range (native vs. alien). See Fig. 5.3 legend for species abbreviations. Insect weight was included as covariate in the models. 116 Resistance alien and natives pines to herbivores

Table 4.2. Results of the mixed model for the cafeteria bioassays with species (SP) and herbivore-induction treatment (T) as fixed factors. Analyses were carried out independently between control plants and (a) the consumed volume of stem by the large pine weevil Hylobius abietis and (b) the specific ingestion of needles and caterpillar survival of the pine processionary moth Thaumetopoea pityocampa as dependent variables. Nine pine species were tested with the weevil and six were tested with the moth, so that independent analyses were performed for each insect bioassay. a. Hylobius cafeteria bioassay Consumed volume of stem Source df F P Block 9, 9 4.98 0.012 Treatment (T) 1, 9 0.27 0.616 Species (SP) 5, 89 7.83 <.0001 T*SP 5, 89 0.52 0.758 Insect weight 1, 94 17.15 <.0001 b. Thaumetopoea cafeteria bioassay Specific ingestion of needles Caterpillar Survival Source df F P df F P Block 9, 9 4.51 0.017 9, 9 0.71 0.690 Treatment (T) 1, 9 0.01 0.938 1, 9 0.31 0.593 Species (SP) 5, 89 12.09 <.0001 5, 90 7.65 <.0001 T*SP 5, 90 1.42 0.225 5, 91 0.67 0.643 Insect weight 1, 64 1 0.320 1, 65 27.35 <.0001

Pine species also differed significantly in the ingestion of needles by Thaumetopoea caterpillars (Table 5.2b). Less consumed needles had higher concentrations of condensed tannins (r2 = 0.5, P < 0.05, n = 6). Specifically,P. pinaster and P. coulteri were the most consumed pine species (Fig. 5.5b). These two pine species also showed the greatest caterpillar survival (Fig. 5.5c). Specific ingestion of needles and caterpillar survival were not significantly affected by herbivore-induction treatment (Table 5.2b; Figs. 5.5b and c). Although, specific ingestion of needles was marginally

different between native and alien pine species (F1, 89 = 2.96, P = 0.08). And, in spite of that pine seedlings did not previously show induction by Thaumetopoea, specific ingestion of needles was lower in alien than native pines (significant Range*Treatment

interaction; F1, 89 = 4.96, P < 0.05; Fig. 5.5b).

117 Chapter 5

DISCUSSION

Given that alien pine seedlings were not less damaged than native pines when exposed to native generalist herbivores, our results do not support the predictions of the NEH. Our observations agree with a recent meta-analysis that found no differences in the damage caused by native herbivores with respect to coexisting alien and native species (Chun et al. 2010).

Hylobius has been documented to feed on several conifer species, but there are no available studies that simultaneously compare several alien and native pine species with regard to the preference of this insect. Under field conditions, Hylobius significantly feeds more on alien P. radiata than on native P. pinaster seedlings (Zas et al. 2006, 2008). However, studies evaluating the damage that other weevil species cause to alien and native pines showed opposite results. For example, observational field studies have shown that the bark beetleTomicus piniperda attacks P. pinaster more than P. radiata (Lombardero et al. 2008).

With respect to herbivory by Thaumetopoea, it reportedly feeds on several pine species across Europe but the results on the preference of this insect for different species are also contradictory. For example, in Greece, Avtzis (1986) found different resistance against this caterpillar among five pine species,P. radiata and P. pinea being the most and the least consumed, respectively. On the contrary, in Italy, Thaumetopoea caterpillars showed no preferences between different alien and native pine species (Tiberi et al.1999).

The constitutive concentration of chemical defenses differed among pine species but only the concentration of total phenolics and condensed tannins were greater in the needles of alien than of native pines. Detectable quantitative responses to the herbivore-induction treatments were caused only by Hylobius, which increased the diterpenes concentration in the stem of some species. Four of the nine assayed pine species, two native (P. pinaster and P. sylvestris) and two alien species (P. coulteri and P. roxburghii), showed a quantitative induction of diterpenes after herbivory by Hylobius. Contrary to our expectations, the response did not differ between alien and native species. Induced defensive responses in the native pines to the pine weevil have been reported elsewhere (Heijari et al. 2005; Sampedro et al. 2011a) but inducibility remains unknown.

118 Resistance alien and natives pines to herbivores

Induced responses in alien pines might be for several reasons. First, the coexistence of other insect herbivores belonging to the same feeding guild as Hylobius, such as Pissodes spp. in North America, might explain the induction of stem diterpenes by the weevil in the American origin P. coulteri. Secondly, the Resource Availability Hypothesis (Coley et al. 1985) assumes that fast-growing species will be less defended because they invest more in growth than do slow-growing species, which invest more in defenses under the same environmental conditions (Blumenthal 2006). According to pines’ classification from Rejmánek& Richardson (1996) and Grotkopp et al. (2002), only P. radiata is considered an invasive species and also fast-growing species in many parts of the world, while the other studied alien pines are mainly considered non-invasive species and slow-growing species. Thus, non-invasive species might be better defended than P. radiata, which appears to invest more in growth than in defenses.

By contrast, the Thaumetopoea treatment caused no significant defensive response and, therefore, no differences in inducibility between alien and native pines were detected. The lack of a significant induction to Thaumetopoea might be owed to several non-exclusive factors. First, our pine species might have not been damaged long enough by herbivores to provoke a defensive response (Gardner & Agrawal 2002; Underwood et al. 2005). Second, although it is known that measurable responses to Hylobius by native pines can be found within 48 h (Sampedro et al. 2011a), the response of pines to Thaumetopoea caterpillars might be slower, requiring a greater lag time between the damage and the induced response (Underwood et al. 2005). Furthermore, there is also a lag time from the translocation of the signal from damaged to undamaged leaves (Howe & Schaller 2008); the activation of a defensive response in undamaged tissues might require more time than the length of our experiment (6 d) for quantitative changes in secondary chemical metabolites to be detected. Finally, the induction of defenses might involve qualitative changes in secondary compounds rather than quantitative changes (Petrakis et al. 2005; Sampedro et al. 2010).

We expected that prior exposure of the seedlings to herbivores would influence the consumption of both generalist herbivores in the bioassays. Nevertheless, for comparisons of alien vs. native pines, in the Hylobius cafeteria bioassays, alien pine species were more consumed while in the Thaumetopoea cafeteria bioassays, native pines were more consumed. The lack of induced response after herbivore-induction

119 Chapter 5 treatment by Thaumetopoea did not agree with the consumption differences detected between alien and native pines when comparing Thaumetopoea-treated and control plants in the Thaumetopoea cafeteria bioassays. Thus, those differences might be explained by the higher content of constitutive total phenolics in alien compared to native pines. Cafeteria bioassays with insects were also performed with several pine and insect species. The results provide useful information but it should be noted that some other factors could affect the accuracy of the results. In other studies, cafeteria bioassays using Hylobius in P. radiata and P. pinaster have registered significant differences in consumption between species and induction treatments (Moreira et al. 2009; Zas et al. unpublished data). However, our results from bioassays performed with Hylobius and Thaumetopoea were not conclusive, since, although differences between some species were found, the herbivores showed little sensitivity to changes in the nutritional quality or other physical or chemical properties of the tissues offered, as reported in previous studies (Hódar et al. 2002). However, very little is known about the ability of pine trees to produce induced responses against those insects, as reflected by the present work.

Our study is the first to report early plant resistance and response to generalist herbivores in several alien and native woody species. Overall, our findings do not provide support for the NEH. Our results concerning chemical defenses suggest that alien pines in regions where they coexist with native congeners are not well defended, and might be controlled by native generalist herbivores. As hypothesized in previous studies, native herbivores might be one reason why there is a low incidence of invasion by alien pines in Europe (Carrillo-Gavilán & Vilà 2010).

Acknowledgements

We thank B Santos, D Blanco, A García, L Martínez, A Pazos and E Dios for greenhouse and laboratory assistances; S Piñeiro and M J Villanueva for field assistance and maintenance of insects; and P González, E Manzano, A Montero and A Montesinos for their comments on a previous draft of the manuscript. Research was partially founded by the Spanish Ministerio de Ciencia e Innovación, projects Consolider-Ingenio MONTES (CSD2008-00040), COMPROPIN (AGL2010-18724), and CSIC (project 200840I153).

120 121 122 Chapter 6

General Discussion

123 Chapter 6

“Todo lo que podemos hacer es tener siempre presente que todo ser orgánico está esforzándose por aumentar en razón geométrica, (…), tiene que luchar por la vida y sufrir gran destrucción. Cuando reflexionamos sobre esta lucha nos podemos consolar con la completa seguridad de que la guerra en la naturaleza no es incesante, que no se siente ningún miedo, que la muerte es generalmente rápida y que el vigoroso, el sano, el feliz, sobrevive y se multiplica”

El origen de las especies - Charles Darwin

124 General discussion

Current invasion status of alien conifers in Europe

Studies on invasions by alien conifers are biased to regions in the Southern Hemisphere (Richardson & Rejmánek 2004; Simberloff et al. 2010). We reviewed the historical records of alien conifer invasion in Europe (i.e., species with a native range outside the continental boundaries of Europe) by screening the DAISIE database and the ISI Web of Science. Our literature review (Chapter 2) showed that, according to DAISIE, there are 54 alien conifer species in Europe. Thirty-seven of these have been studied, to some extent, in a total of 131 papers (212 records). There are many studies on alien conifers in Europe but few ecological aspects related to biological invasions have been investigated. Not many studies consider alien conifers from the point of view of biological invasions. Not many studies consider alien conifers from the point of view of biological invasions. Pseudotsuga menziesii is the species recorded as naturalized in the majority of countries and the UK is the one with the most naturalized species. Seven species were defined as naturalized, while only one,Pinus strobus, is considered as an invasive species. Most studies have concentrated on species of greater economic relevance such as Picea sitchensis or Pseudotsuga menziesii and on the natural enemies that might damage these species, therefore reduce their wood production.

We propose that the little evidence of invasion by alien conifers in Europe, as compared to the Southern Hemisphere (Chapter 2), depends upon historical, socioeconomic aspects, and that it also has an ecological base. First, time since introduction and propagule pressure might play an important role on the reduced evidence of invasion by alien conifers in Europe. The majority of alien conifers in Europe were introduced later than in the Southern Hemisphere. Moreover, plantations in Europe have been done at a smaller scale than in other regions, basically because the extension of forested land is much less than in South American countries, and therefore the scale of plantations is likely to be proportional to its availability. Additionally, in Europe, forest management policies are changing towards a more conservative approach that recommends the use of native species for planting (Quine et al. 2004).

Regarding ecologically based causes, as recently pointed out by Essl et al. (2011), differences in alien conifer naturalization worldwide are a potential consequence of the biogeographic features in the recipient community. The presence of congeners in the introduced range will likely decrease the niche opportunity for alien establishment, because natural enemies, competitors and mutualisms might be shared between native and alien conifers. Thus, invasion differences might also depend on similarities between native and alien species, where phylogenetic proximity can result in higher

125 Chapter 6 colonization rates of natural enemies on alien trees (Goßner et al. 2009). Most alien conifers in Europe have a temperate North American and Asian origin and have European congeners (e.g., the genus Pinus). So, once introduced, these conifers might recruit natural enemies from closely related species, and be strongly controlled by them. This may be one reason why invasion events are less common in Europe than in the Southern Hemisphere. In the next sections we discuss the resistance of conifers to cope with natural enemies.

As suggested in Chapter 2, alien conifer plantations in Europe might be in a period of time lag between arrival and establishment. From a management point of view, since the early stages of invasion are most amenable to cost-effective management (Wittenberg & Cock 2001), it is recommended that in the future we act upon those species with invasive tendencies in other regions of the world to prevent possible invasion. For instance, revisiting as much sites as possible where alien conifers were planted long ago in Europe (Richardson & Rejmánek 2007). In this sense, Essl et al (2011) observe that climate, habitat diversity, introduction effort and biogeographic features (natural enemies, competitors and mutualisms) in the introduced range co- determine alien conifer naturalization across the globe. Thus, national management plants could in principle identify areas where alien conifers can be grown with high and low risk of invasion. Hence, it is necessary to implement special measures to manage the spread from alien conifer plantations in the areas with the highest risk of invasion (Rouget et al 2002).

Early stages of the plant life-cycle limiting the establishment of alien conifers

As we mentioned in the General Introduction, conifer traits such as short juvenile period, short interval between large seed crops, small seed mass (Rejmánek & Richardson 1996; Richardson & Rejmánek 2004) and fast relative growth rate (Grotkopp et al. 2002) are correlated with high invasion success. In this thesis, we focused on the early life-cycle stages of plant establishment (Fig. 1.2). Our prediction was that conifer species with small seed mass might be more successful in establishing themselves.

Comparison of seed removal rates, among different pine species differing in invasiveness, showed that in Mediterranean shrublands the native fauna was very efficient in removing seeds (Chapter 3). Although, seed survival time was significantly different between species, there were no significant differences in seed removal between small and large-seeded species. As suggested by Chapter 2, our study

126 General discussion showed evidence that the native fauna might play an important role as a component of biotic resistance to pine establishment (Elton 1958; Maron & Vilà 2001; Vilà & Gimeno 2003; Nuñez et al. 2008).

Nevertheless, seed removal is only one barrier towards seedling establishment. It is assumed that seedlings from larger seeds have a higher probability of survival from seedling emergence to adulthood (Moles & Westoby 2006), and also that they perform better during early development than smaller seeds (Moles & Westoby 2004). Our initial plan was to complete the sequential stages of the early life-cycle of the 16 assayed pine species in the Aznalcóllar shrublands. However, several unexpected events aborted the experiments. In the germination tests, we buried seed bags in the study area following the same methodology as in Chapter 4, but the native fauna removed a large proportion of seed bags, considerably reducing the sample size. Regarding seedling survival tests, seedlings of the assayed pine species were first placed to emerge in greenhouse conditions during one year before transplantation to the field, but a pathogen attack caused the death of a huge number of seedlings.

Fortunately, we found an alternative case study. We compared the early life- cycle stages between the alien Douglas fir and the native Silver fir in three habitat types (beech forests, holm-oak forests and heathlands) in the Montseny Natural Park. (Chapter 4). Therefore, in Chapter 4, we incorporated the elements of species invasiveness with those of habitat invasibility (Richardson 2006). Results showed that the probability of invasion success of both firs decreased along the life-cycle stages in all habitats. Only the higher germination capacity of Douglas fir seemed to predict a better establishment success compared to Silver fir. However, in general, the ecological processes that prevented the establishment of both conifers were the native fauna, in holm-oak forests, and drought stress, in heathlands and beech forests.

There is a mismatch between the current naturalization patterns of Douglas fir in Montseny and the habitat invasibility. Currently, Douglas fir establishment is only observed in heathlands, yet we found a very low invasibility for this habitat. Our suggestion is that low invasibility in heathlands is possibly counterbalanced by a higher propagule pressure in this habitat. This mismatch might predict that Douglas fir naturalization observed in heathlands is a consequence of processes and/or mechanisms that act at mid life-cycle stages from seedling to adulthood where establishment success is favored.

127 Chapter 6

It is typical to find that, differences in conifer species invasiveness are predicted by small seed mass, since small seeds are associated with several potentially important phenomena: larger numbers of seeds produced, best dispersion, high initial germinability, and a shorter chilling period needed to overcome dormancy (Richardson & Rejmánek 2004). Our findings from Chapter 3 & 4 suggested that it might be seed crop size and germination capacity and not seed size per se that might predict invasion success. Small-seeded species produce seeds in larger quantities than larger-seeded species (Jakobsson & Eriksson 2000; Henery & Westoby 2001) and this can compensate for the effects of seed loss or seedling mortality (Moles et al. 2003; Moles & Westoby 2004). This is in accordance with the importance that propagule pressure has on invasion success (Lockwood et al. 2005; Colautti et al. 2006). Many pine species that escape from plantations and become invasive are often those that have been cultivated the most widely and for the longest time (Krivánek et al. 2006), as discussed in Chapter 2. Additionally, alien species are generally reported to germinate earlier, better, and in a wider range of conditions than native species (review in Pyšek & Richardson 2007) as we showed in Chapter 4.

For future research we suggest to compare invasiveness of several closely related species (Kleunen et al. 2010), not only regarding early life-cycle stages, but also at mid life-cycle stages. In this approach it would be important to consider phylogenetic relationships among species, because accounting for phylogenetic information can change the invasion status from insignificant to significant (Sol et al. 2008).

Differences in natural enemy resistance among native and alien conifers

The Natural Enemies Hypothesis predicts that alien plant species are successful invaders in the new range due to a lack of natural enemies (Maron & Vilà 2001). The impact of native herbivores on alien plants is determined by the interaction between their ability to recognize the alien plant as a food resource (Carpenter & Cappuccino 2005), the effectiveness of the alien plant defenses against herbivores in the new range (Stowe et al. 2000), and the ability of herbivores to overcome alien plant’s defenses (Rausher 2001). Thus, the level of damage inflicted on alien species is not always lower than on coexisting native plants, even when the species are taxonomically related (Agrawal & Kotanen 2003). Moreover, a recent meta-analysis noted that the majority of native insect species that colonize alien trees are generalist herbivores (Bertheau et al. 2010). We expected that, as predicted by the Natural Enemies Hypothesis, generalist herbivores would cause lower damage to alien than to native pine species. But, alien

128 General discussion plants exposed to generalist native herbivores would induce weaker defenses than would native pines, due to a lack of a common evolutionary history. And also, herbivory by native herbivores would further decrease; more so in native than in alien pines.

Contrary to our expectations, generalist native herbivores did not caused more damage in native than in alien pines. There were mild differences in constitutive and induced chemical defenses caused by two native generalist herbivores (Hylobius abietis and Thaumetopoea pytiocampa on alien and native pine seedlings (Chapter 5). Constitutive concentration of total phenolics and condensed tannins was the only variable that was greater in needles of alien than of native pines. Detectable quantitative responses, after the herbivore-induction treatments, were caused only by H. abietis, which increased the diterpenes concentration in the stem of some species. Nevertheless, the response did not differ between alien and native species. T. pytiocampa caused no significant defensive response, and therefore, no differences in inducibility were detected between alien and native pines.

Our results did not support the assumption made by the Natural Enemy Hypothesis for which native herbivores would avoid alien plants due to a lack of recognition. Results from other studies are rather contradictory. Some show either no evidence that generalist herbivores avoided alien plants (Morrison & Hay 2011) or, by contrast, they show that generalist herbivores have greater impact on native species than on alien species (Schaffer et al. 2011). Our observations agree with a recent meta- analysis that found no differences in the damage caused by native herbivores with respect to coexisting alien and native species (Chun et al. 2010).

In relation to induced chemical defenses, our results suggest that alien pines, in regions where they coexist with native congeners, are not well defended and might be controlled by native generalist herbivores. This is in accordance with Lind & Parker (2011), who compared chemical deterrents in several invasive and co-occurring native plant species and found that alien species did not show higher deterrence to native generalist herbivores than native species.

Our experimental approach was greenhouse based and we did not explore the full predictions of the Natural Enemy Hypothesis. To have a complete picture of the importance of natural enemies on pine species we should incorporate spatial and temporal variation (Gurevitch et al. 2011). To confirm the observed patterns, it would be interesting to compare plant damage of species in field conditions. Further testing, considering other natural enemies (e.g., Tomicus piniperda) and analyzing chemical

129 Chapter 6 defenses in different parts of the plant (e.g., above vs. belowground) would largely complement the results of this thesis. Another venue of research would be to assess plant tolerance mechanisms such as growth rate after herbivore attack, to test if damage is compensated with a higher growth rate in alien than in native species.

130 Conclusions

131 132 Conclusions

I) Our literature review illustrates there is little evidence of invasion by alien conifers in Europe. From the 54 species considered as alien by the DAISIE database (www.europe-aliens.org/), only 37 of these conifers have been studied to some extent. Among them, ISI papers only mention 7 species as naturalized and one species (Pinus strobus in the Czech Republic) as invasive.

II) The low introduction effort of alien conifers, long lag-time from plantation to establishment, and the phylogenetic proximity between alien and native conifers are possible reasons for their low expansion in Europe to date. The taxonomic relatedness between alien and native species can result in high colonization rates of natural enemies on alien trees.

III) In Mediterranean shrublands, native fauna was very efficient in removing seeds of Pinus species, where rodents and ants were the main seed removers. Although mean seed survival time was significantly different between species, there were no significant differences in seed removal neither between small and large-seeded species nor between alien and exotic species in any season. This result suggests that in pine species with high invasiveness, due to small seed mass, large seed crops might counterbalance high seed removal rates.

IV) In the Montseny Natural Park, Douglas fir (Pseudotsuga menziesii) performed better in beech forests, where seed survival, seed germination and seedling survival were higher than in holm-oak forests and heathlands. Soil disturbance by native fauna in holm-oak forests, and drought stress in heathlands, where Douglas fir is considered naturalized, were the most important factors limiting seedling establishment.

V) There are no major differences in the performance of early life-cycle stages of Douglas fir and the coexisting native Silver firAbies ( alba). High germination rates of Douglas fir seeds is the only different stage that together with higher annual seed production rates would increase the probability of seedling establishment.

VI) There is a mismatch between the degree of invasion by Douglas fir in Montseny and our results on invasibility”. The low invasibility in heathlands is possibly counterbalanced by a higher propagule pressure in this habitat. This mismatch might be due to the fact that the observed naturalization of Douglas fir in heathlands is a consequence of processes and/or mechanisms that act at mid life-cycle stages, from seedling to adulthood, where establishment success might be favored.

133 Conclusions

VII) Contrary to predictions of the Natural Enemy Hypothesis, alien pine seedlings were not less damaged by native generalist herbivores than the native ones. Furthermore, the analysis of chemical defenses did not reveal major differences between alien and native pine species. This result suggests that alien pines present in regions where they coexist with native congeners are not well defended. Their invasion potential might be partially controlled by native generalist herbivores.

VIII) In summary, our results reinforce the hypothesis that native generalist enemies in the introduced range act as a barrier against the establishment of alien conifers (Biotic Resistance Hypothesis). This might be one reason why there is a low incidence of invasion by these aliens in Europe.

134 Conclusiones

135 136 Conclusiones

I) Nuestra revisión de la literatura científica ilustra poca evidencia de invasión por coníferas exóticas en Europa. De las 54 especies consideradas como exóticas en Europa por DAISIE (www.europe-aliens.org/), sólo 37 de estas especies han sido estudiadas en cierta medida. Entre ellas, las publicaciones del ISI sólo mencionan a 7 especies como naturalizadas y sólo a una especie (Pinus strobus en la República Checa) como invasora.

II) El bajo esfuerzo de introducción de las coníferas exóticas, el tiempo largo que debe pasar desde que las coníferas son plantadas hasta su establecimiento, y la proximidad filogenética entre especies exóticas y nativas de coníferas son razones posibles de la baja expansión que se observa actualmente en Europa. La proximidad filogenética entre las especies exóticas y nativas puede provocar unas tasasde colonización altas por parte de los enemigos naturales sobre las coníferas exóticas.

III) En áreas de matorral Mediterráneo, la fauna nativa fue muy eficiente en eliminar semillas de especies de Pinus, siendo roedores y hormigas los principales depredadores. Aunque los tiempos medios de supervivencia de las semillas fueron significativamente diferentes entre especies, no hubo diferencias en las tasas de depredación ni entre especies de semilla pequeña y de semilla grande ni comparando entre especies nativas y exóticas en ningún período muestreado. Este resultado sugiere que las especies de pino con alta capacidad de invasión, debido a un tamaño de semilla pequeño, las tasas altas de depredación de semillas podrían estar compensadas con una alta producción de semillas.

IV) En el Parque Natural del Montseny, el abeto de Douglas (Pseudotsuga menziesii) se estableció mejor en el hayedo debido a que la supervivencia y germinación de las semillas, así como la supervivencia de las plántulas tuvieron unas tasas más altas que en el encinar y en el brezal. La perturbación del suelo por parte de la fauna nativa en el encinar, y el estrés provocado por sequía en el brezal, donde el abeto de Douglas es considerado naturalizado, fueron los factores más importantes que limitaron el establecimiento de las plántulas.

V) No existen grandes diferencias entre los estadios iniciales del ciclo de vida del abeto de Douglas y del abeto común (Abies alba). Las tasas altas de germinación de las semillas del abeto de Douglas es el único estadio que junto con una alta producción de semillas, incrementaría la probabilidad de establecimiento de las plántulas.

137 Conclusiones

VI) Hay un desacople entre el grado de invasión del abeto de Douglas en el Montseny y los resultados de invasibilidad que hemos encontrado. La baja invasibilidad en el brezal es posiblemente contrarrestada con una alta presión de propágulos en dicho hábitat. Este desacople podría ser debido a que la naturalización que se observa del abeto de Douglas en el brezal es consecuencia de procesos y/o mecanismos que actúan en etapas intermedias de su ciclo de vida, desde la fase de plántula hasta la de adulto, en donde su éxito de establecimiento podría estar favorecido.

VII) En contra de las predicciones de la Hipótesis de los Enemigos Naturales, las plántulas de pinos exóticos no fueron menos dañadas por herbívoros generalistas nativos que las plántulas de las nativas. Además, los análisis de defensas químicas no mostraron grandes diferencias entre especies exóticas y nativas de pinos. Esto sugiere que los pinos exóticos presentes en regiones donde coexisten con especies congenéricas nativas no están bien defendidos. Su potencial de invasión podrían estar parcialmente controlado por herbívoros generalistas nativos.

VIII) En resumen, nuestros resultados refuerzan la hipótesis de que los enemigos generalistas nativos de la nueva área de introducción actúan como barreras contra el establecimiento de las coníferas exóticas (Hipótesis de la Resistencia Biótica). Esto podría ser una razón por la que existe poca evidencia de invasión de estas especies exóticas en Europa.

138 139 140 References

141 142 General references

Adamowski W (2004) Why don’t alien conifers invade the Bialowieza Forest? Weed Technology 18:1453-1456. Agrawal AA, Kotanen P (2003) Herbivores and the success of exotic plants: a phylogenetically controlled experiment. Ecology Letters 6:712–715. Agrawal AA, Kotanen P, Mitchell C, Power AG, Godsoe W, Klironomos J (2005) Enemy release? An experiment with congeneric plant pairs and diverse above- and belowground enemies. Ecology 86:2979-2989. Agrawal AA (2010) Current trends in the evolutionary ecology of plant defence. Functional Ecology 25:420-432. Alcántara JM, Rey P, Sánchez-Lafuente A, Valera F (2000) Early effects of rodent post-dispersal seed predation on the outcome of the plant-seed disperser interaction. Oikos 88: 362-370. Alonso-Zarazaga MA, Goldarazena A (2005) Presencia en el País Vasco de Rhyephenes humeralis (Coleoptera, Cuculionidae), plaga de Pinus radiata procedente de Chile. Boletín de la Sociedad Entomológica Aragonesa 36:143-146. Alpert P, Bone E, Holzapfel C (2000) Invasiveness, invasibility and the role of environmental stress in the spread of non-native plants. Perspectives in Plant Ecology, Evolution and Systematics 3:52-66. Amezaga I (1996) Monterrey pine (Pinus radiata D. Don) suitability for the pine shoot beetle (Tomicus piniperda L.) (Coleoptera: Scolytidae). Forest Ecology and Management 86:73-79. Amezaga I, Rodríguez MA (1998) Resource partitioning of four sympatric bark beetles depending on swarming dates and tree species. Forest Ecology and Management 109:127-135. Andersen AM (1987) Effects of seed predation by ants on seedling densities at a woodland site in SE Australia. Oikos 48:171-174. Andrieu E, Debussche M (2007) Diaspore removal and potential dispersers of the rare and protected Paeonia officinalis L. (Paeoniaceae) in a changing landscape. Botanical Journal of Linnean Society 154:13-25. Arbea JI, Jordana R (1988) Efecto de la repoblación con alerce (Larix kaempferi) en la zona norte de Navarra, sobre la estructura de la poblaciones de colembolos edáficos. Biología ambiental: actas del Congreso de Biología Ambiental (II Congreso Mundial Vasco), pp. 253-265. Pamplona, España. Avtzis N (1986) Development of Thaumetopoea pityocampa (Den. and Schiff.) in relation to food consumption. Forest Ecology and Management 15:65-68.

143 General references

Bajo J, Santamarıía O, Diez J (2008) Cultural characteristics and pathogenicity of Pestalotiopsis funerea on Cupressus arizonica. Forest Pathology 38:263–274. Barbagallo S, Binazzi A, Ortu S (2005) On the presence in Italy of the nearctic aphid Essigella californica (Essig) living on American pines. Redia 88:79-84. Battisti A (2006) Insect populations in relation to environmental change in forests of temperate Europe. Invasive forest insects, introduced forest trees, and altered ecosystems (ed TD Paine), pp. 127-140. Kluwer Academic Publishers, The Netherlands. Bertheau C, Brockerhoff EG, Roux-Morabito G, Lieutier F , Jactel H (2010) Novel insect-tree associations resulting from accidental and intentional biological invasions: a meta-analysis of effects on insect fitness. Ecology Letters 13:506–515. Binimelis R, Born W, Monterroso I, Rodríguez-Labajos B (2007) Socio economic impact and assessment of biological invasions. Biological Invasions (ed W Nentwig), pp. 331-347. Ecological Studies, Springer-Verlag Berlin Heidelberg, 193. Blumenthal DM (2006) Interactions between resource availability and enemy release in plant invasion. Ecology Letters 9:887-895. Boada M (2000) El Montseny. Fifty Years of Landscape Evolution. Publications de l’Abadia de Montserrat, Barcelona, Spain. Boada M, Broncano M (2003) Estudi sobre la distribució d’espècies forestals allòctones al Parc Natural del Montseny. Evidències de la naturalització de l’avet de Douglas (Pseudotsuga menziesii) al Parc Natural del Montseny. ICTA-Institut de Cièncie i Tecnologia Ambientals, Universitat Autònoma de Barcelona, Barcelona, Spain. Bolòs O, Vigo J (1990) Flora Dels Països Catalans. Barcino Press, Barcelona, Spain. Borchert M, Johnson M, Schreiner D, Vander Wall S (2003) Early post-fire seed dispersal, seedling establishment and seedling mortality of Pinus coulteri (D. Don) in central coastal California, USA. Plant Ecology 168:207–220. Broncano M, Vilà M, Boada M (2005) Evidence of Pseudotsuga menziesii naturalization in montane Mediterranean forests. Forest Ecology and Management 211:257-263. Bucharova A, van Kleunen M (2009) Introduction history and species characteristics partly explain naturalization success of North American woody species in Europe. Journal of Ecology 97:230-238. Byers JE, Reichard S, Randall JM, Parker IM, Smith CS, Lonsdale WM, Atkinson IAE, Seastedt TR, Williamson M, Chornesky E, Hays D (2002) Directing research to reduce impacts of non-indigenous species. Conservation Biology 16:630–640. 144 General references

Caccia FD, Ballaré CL (1998) Effects of tree cover, understory vegetation, and litter on regeneration of Douglas-fir Pseudotsuga( menziesii) in southwestern Argentina. Canadian Journal of Forest Research 28: 683–692. Carey JR (1996) The incipient Mediteranean fruti fly population in California: implications for invasion biology. Ecology 77:1690-1697. Carpenter D, Cappuccino N (2005) Herbivory, time since introduction and the invasiveness of exotic plants. Journal of Ecology 93:315–321. Carrascal L, Tellería J (1990) Impacto de las repoblaciones de Pinus radiata sobre la avifauna forestal del norte de España. Ardeola 37:247-266. Carrillo-Gavilán M A, Vilà M (2010) Little evidence of invasion by alien conifers in Europe. Diversity and Distributions 16:203–213. Carrillo-Gavilán MA, Hadrien L, Vilà M (2010) Comparing seed removal of 16 pine species differing in invasiveness. Biological Invasions 12:2233–224. Castro J, Gómez JM, García D, Zamora R, Hódar JA (1999) Seed predation and dispersal in relict Scots pine forests in southern Spain. Plant Ecology 145:115- 123 Castro-Díez P, González-Muñoz N, Alonso A, Gallardo A, Poorter L (2009). Effects of exotic invasive trees on nitrogen cycling: a case study in Central Spain. Biological Invasions 11:1973–1986. Charpin D, Calleja M, Lahoz C, Pichot C, Waisel Y (2005) Allergy to cypress pollen. Allergy 60:293–301. Chun YJ, van Kleunen M, Dawson W (2010) The role of enemy release, tolerance and resistance in plant invasions: linking damage to performance. Ecology Letters 13:937-946. Chytrý M, Maskell L, Pino J, Pyšek P, Vilà M, Font X, Smart S (2008) Habitat invasions by alien plants: a quantitative comparison among Mediterranean, subcontinental and oceanic regions of Europe. Journal of Applied Ecology 45:448–458. Colautti R, Grigorovich I, MacIsaac H (2006) Propagule pressure: a null model for biological invasions. Biological Invasions 8:1023-1037 Coley PD, Bryant JP, Chapin FS (1985) Resource availability and plant antiherbivore defense. Science 230:895-899. Collier FA, Bidartondo MI (2009) Waiting for fungi: the ectomycorrhizal invasion of lowland heathlands. Journal of Ecology 97:950-963. Crawley, MJ, Harvey, PH, Purvis, A (1996) Comparative ecology of the native and alien floras of the British Isles. Philosophical Transactions of the Royal Society of London Series B-Biological Sciences 351:1251-1259. Crook CS (1997) A (very) provisional checklist of Conifers in the British Isles. BSBI NEWS 75 (ed G Ellis), pp. 42-47. Lostock Hall, Preston, Lancaster, UK. 145 General references

Crook JA (2005) Lag times and exotic species: The ecology and management of biological invasions in slow-motion. Ecoscience 12:316-329. Cuadros S, Francia JR (1999) Caracterización del sitio de ensayo de especies forestales de Lanjarón, vertiente sur de Sierra Nevada. Aspectos climatológicos y fitoclimáticos. Investigación Agraria, Sistemas y Recursos Forestales 1:143-158. D’Antonio CM (2000) Fire, plant invasions and global change. Invasive species in a changing world (eds HA Mooney & RJ Hobbs), pp. 65-93.Island Press, Washington, USA. DAISIE (2009) Handbook of Alien Species in Europe. Springer, Dordrecht. Darwin CR (1859) On the origin of species by means of natural selection, or the preservation of favoured races in the struggle for life. John Murray, London, UK. Davis MA, Grime JP, Thompson K (2000) Fluctuating resources in plant communities: a general theory of invasibility. Journal of Ecology 88:528–534. Davis MR, Lang MH (1991) Increase nutrient availability in topsoils under conifers in the South Island high country. New Zealand Journal of Forestry Science 21:165-179. Day K (1984) The growth and decline of a population of the spruce aphid Elatobium abietinum during a three year study, and the changing pattern of fecundity, recruitment and alary polymorphism in a Northern Ireland Forest. Oecologia 64:118-124. Day KR, Armour HL, Henry CJ (1999) The performance of the green spruce aphid (Elatobium abietinum Walker) on provenances of Sitka spruce (Picea sitchensis (Bong.) Carr.). Colloques de l’INRA; Physiology and genetics of tree- phytophage interactions: International Symposium, pp.199-210. De Candolle AP (1855) Géographie Botanique Raisonnée. Paris, Masson, vol 2. Deveny AJ, Fox LR (2006) Indirect interactions between browsers and seed predators affect the seed bank dynamics of a chaparral shrub. Oecologia 150:69-77 DeWalt SJ, Denslow JS, Ickes K (2004) Natural enemy release facilitates habitat expansion of the invasive tropical shrub Clidemia hirta. Ecology 85:471–483. Di Castri F (1989) History of biological invasions with emphasis on the Old World. Biological invasions:a global perspective (eds J Drake, F di Castri, R Groves, F Kruger, H.A Mooney, M Rejmánek & M Williamson), pp. 1-30. Wiley, New York, New York, USA. Dietz H, Steinlein T, Ullmann I (1999) Establishment of the invasive perennial herb Bunias orientalis L.: an experimental approach. Acta Oecologia 20:621–632.

146 General references

Dirzo R, Mendoza E (2007) Size-related differential seed predation in a heavily defaunated Neotropical Rain Forest. Biotropica 39:355-362 Domènech R, Vilà M (2006) The role of successional stage, vegetation type and soil disturbance on Cortaderia selloana invasion. Journal of Vegation Science 17:591-598. Dunne JA, Parker VT (1999) Species-mediated soil moisture availability and patchy establishment of Pseudotsuga menziesii in chaparral. Oecologia 119:36-45. Eckert AJ, Hall BD (2006) Phylogeny, historical biogeography, and patterns of diversification forPinus (Pinaceae): Phylogenetic tests of fossil-based hypotheses. Molecular Phylogenetics and Evolution 40:166–182. Elton CS (1958) The ecology of invasions. Methuen, London, United Kingdom Espelta JM, Bonal R, Sánchez-Humanes B (2009) Pre-dispersal acorn predation in mixed oak forests: interspecific differences are driven by the interplay among seed phenology, seed size and predator size. Journal of Ecology 97:1416–1423. Essl F (2007) Verbreitung, Status und Vergesellschaftung von Pinus strobus in Österreich. Tuexenia 27:59-72. Essl F, Moser D, Dullinger S, Mang T, Hulme PE (2010) Selection for commercial forestry determines global patterns of alien conifer invasions. Diversity and Distributions 16:911-921 Essl F, Mang T, Dullinger S, Moser D, Hulme PE (2011) Macroecological drivers of alien conifer naturalizations Worldwide. Ecography 34:1-9. Evans HF (1985) Great spruce bark beetle, Dendroctonus micans: an exotic pest new to Britain. Antenna 9:117 -121. Fabre JP, Auger-Rozenberg MA, Chalon A, Boivin S, Roques A (2004) Competition between exotic and native insects for seed resources in trees of a Mediterranean forest ecosystem. Biological Invasions 6:11–22. Fedriani JM, Rey PJ, Garrido JL, Guitián J, Herrera CM, Medrano M, Sánchez- Lafuente AM, Cerdá X (2004) Geographical variation in the potential of mice to constrain an ant-seed dispersal mutualism. Oikos 105:181-191 Fedriani J (2005) Do frugivorous mice choose where or what to feed on? Journal of Mammalogy 86:576–586. Fernandez FA, Evans PR, Dunstone N (1994) Local variation in rodent communities of Sitka spruce plantations: the interplay of successional stage and site-specific habitat parameters. Ecography 17:305-313. Ferreras AE, Galetto L (2010) From seed production to seedling establishment: Important steps in an invasive process. Acta Oecologia 36:211-218.

147 General references

Forestry Commission (2003) Forestry Commission. National Inventory of Woodland and Trees – Great Britain, Forestry Commission, Edinburgh. Gardner SN , Agrawal AA (2002) Induced plant defence and the evolution of counter-defences in herbivores. Evolutionary Ecology Research 4:1131–1151. Gashwiler JS (1969) Seed fall of three conifers in west-central Oregon. Journal of Forest Science 15:290-295. Gatehouse JA (2002) Plant resistance towards insect herbivores: a dynamic interaction. New Phytologist 156:145–169. Gómez JM (2004) Bigger is not always better: Conflicting selective pressures on seed size in Quercus ilex. Evolution 58: 71–80. Goßner MM, Chao A, Bailey RI, Prinzing A (2009) Native fauna on exotic trees: phylogenetic conservatism and geographic contingency in two lineages of phytophages on two lineages of trees. The American Naturalist 173:599-614. Gradecki M, Poštenjak K (2001) Relation between laboratory and nursery germination of silver fir seed Abies( alba Mill.). Znanost u potrajnom gospodarenju hrvatskim šumama: znastvena knjiga. (eds S Matic, APB Krpan & J Gracan), pp. 191-200. Zagreb: Šumarski fakultet Sveucilišta u Zagrebu, Šumarski Institute, Croatia. Greene DF, Johnson EA (1999) Estimating the mean annual seed production of trees. Ecology 75:642-647. Grotkopp E, Rejmánek M, Rost TL (2002) Toward a causal explanation of plant invasiveness: seedling growth and life-history strategies of 29 pine (Pinus) species. The American Naturalist 159:396-419. Gurevitch J, Fox GA, Wardle GM, Inderjit, Taubs D (2011) Emergent insights from the synthesis of conceptual frameworks for biological invasions. Ecology Letters 14:407–418. Gurnell J, Lurz PW, Shirley MD, Cartmel S, Garson PJ, Magris L, Steeles J (2004) Monitoring red squirrels Sciurus vulgaris and grey squirrels Sciurus carolinensis in Britain. Mammal Review 34:54-74. Hadincová V, Münzbergová Z, Wild J, Šajtar L, Marešová, J (2008) Dispersal of invasive Pinus strobus in sandstone areas of Czech Republic. Plant Invasions: Human perception, ecological impacts and management (eds B Tokarsha-Guzik, JH Brock, G Brundu, L Child , CC Daehler & P Pyšek ), pp 117-132. Backhuys Publishers, Leyden, The Netherlands. Hamilton MA, Murray BR, Cadotte MW, Hose GC, Baker AC, Harris CJ, Licari D (2005) Life-history correlates of plant invasiveness at regional and continental scales. Ecology Letters 8:1066-1074.

148 General references

Hang-Hau C (1997) Tree seed predation on degraded hillsides in Hong Kong. Forest Ecology and Management 99:215-221 Hanzélyová D (1998) A comparative study of Pinus strobus L. and Pinus sylvestris L.: growth at different soil acidities and nutrient levels. Plant invasions: ecological mechanisms and human responses (eds U Starfinger, K Edwards, I Kowarik & M Williamson), pp. 185-194. Backhuys Publishers, Leiden, The Netherlands. Harriman R, Morrison BR (1982) Ecology of streams draining forested and non forested catchments in an area of central Scotland subject to acid precipitation. Hydrobiologia 88:251-263. Heijari J, Nerg AM, Kainulainen P, Viiri H, Vuorinen M, Holopainen JK (2005) Application of methyl jasmonate reduces growth but increases chemical defence and resistance against Hylobius abietis in Scots pine seedlings. Entomologia Experimentalis et Applicata 115:117–124. Heil M (2010) Plastic defence expression in plants. Evolutionary Ecology 24:555–569. Henery ML, Westoby M (2001) Seed mass and seed nutrient content as predictors of seed output variation between species. Oikos 92:479-490. Hermann RK, Lavender DP (1999) Pseudotsuga menziesii (Mirb.) Franco. Silvics of North America (eds R Burns & BH Honkala), Vol. 1, pp. 540–557. Conifers. USDA Forest Service. Washington D. C, USA. Herrera CM (1984) Seed dispersal and fitness determinants in wild rose: combined effects of hawthorn, birds, mice, and browsing ungulates. Oecologia 63:386-393 Herrera CM, Jordano P, López-Soria L, Amat JA (1994) Recruitment of a mass- fruiting, bird-dispersed tree: Bridging frugivore activity and seedling establishment. Ecological Monographs 6:315-344. Hobbs RJ, Huenneke LF (1992) Disturbance, diversity, and invasion: Implications for conservation. Conservation Biology 6:324-337. Hódar JA, Zamora R, Castro J (2002) Host utilisation by moth and larval survival of pine processionary caterpillar Thaumetopoea pityocampa in relation to food quality in three Pinus species. Ecological Entomology 27:292-301. Howe GA, Schaller A (2008) Direct defenses in plants and their induction by wounding and insect herbivores. Induced plant resistance to herbivory (ed A Schaller ), pp. 7-29. Springer, New York, USA.. Huggard DJ, Arsenault A (2009) Conifer seed predation in harvested and burned dry Douglas fir forests in sourthern British Columbia. Canadian Journal of Forest Research 39:1548–1556.

149 General references

Hulme PE (1997) Post-dispersal seed predation and the establishment of vertebrate dispersed plants in Mediterranean scrublands. Oecologia 111:91–98. Hulme PE (1998) Post-dispersal seed predation and seed bank persistence. Seed Science Research 8:513-519. Hulme PE, Hunt MK (1999) Rodent post-dispersal seed predation in deciduous woodland: predator response to absolute and relative abundance of prey. Journal of Animal Ecology 68: 417-428. Hulme PE (2011) Addressing the threat to biodiversity from botanic gardens. Trends in Ecology and Evolution 26: 168-174. Humphries SE, Groves RH, Mitchell DS (1991) Plant invasions of Australian ecosystems: a status review and management directions. Kowari 2:1–127. Irizar I, Laskurain N, Herrero J (2004) Wild boar frugivory in the atlantic Basque Country. Galemys 16:125-133. Jacob HS, Minkey DM, Gallagher RS, Borger CP (2006) Variation in post- dispersal weed seed predation in a crop field. Weed Science 54:148-155. Jakobsson A, Eirksson O (2000) A comparative study of seed number, seed size, seedling size and recruitment in grassland plants. Oikos 88:494-502. Janzen DH (1969) Seed-eaters versus seed size, number, toxicity and dispersal. Evolution 23:1-27 Joshi J, Vrieling K (2005) The enemy release and EICA hypothesis revisited: incorporating the fundamental difference between specialist and generalist herbivores. Ecology Letters 8:704-714. Karban R, Myers JH (1989) Induced plant responses to herbivory. Annual Review of Ecology and Systematics 20:331–348. Karban R (2010) The ecology and evolution of induced resistance against herbivores. Functional Ecology 25:339-347. Kay M (1994) Biological control for invasive tree species. New Zealand Forestry 39:35-37. Kolar CS, Lodge DM (2002) Ecological predictions and risk assessment for alien fishes in North America. Science 298:1233-1236. Kowarik I (1995) Time lags in biological invasions with regard to the success and failure of alien species. Plant invasions. General aspects and special problems (eds P Pyšek, K Prach, M Rejmánek & PM Wade), pp. 15-38. Amsterdam: SPB Academic Publ. The Netherlands. Krebs JR (1978) Optimal foraging: Decision rules for predators. Behavioural Ecology (ed Krebs JR & NB Davies), pp. 23-63. Blackwell, Oxford, UK.

150 General references

Krivánek M, Pyšek P, Jarošik V (2006) Planting history and propagule pressure as predictors of invasion by woody species in a temperate region. Conservation Biology 20:1487–1498. Lambdon PW (2008) Is invasiveness a legacy of evolution? Phylogenetic patterns in the alien flora of Mediterranean islands. Journal of Ecology 96:46–57. Lambdon PW, Pyšek P, Basnou C, Hejda M, Arianoutsou M, Essl F, Jarošik V, Pergl J, Winter M, Anastasio P, Andriopoulos P, Bazos I, Brundu G, Celesti-Grapow L, Chassot P, Delipetrou P, Josefsson M, Kark S, Klotz S, Kokkoris Y, Kühn I, Marchante H, Perglová I, Pino J, Vilà M, Zikos A, Roy D, Hulme P (2008) Alien flora of Europe: species diversity, temporal trends, geographical patterns and research needs. Preslia 80:101-149. Lambrinos JG (2002) The variable invasive success of Cortaderia species In a complex landscape. Ecology 83:518–529. Lanner RM (1998) Seed dispersal in Pinus. Ecology and biogeography of Pinus (ed DM Richardson), pp 281-295. Cambridge University. Press, Cambridge, UK. Lavery PB, Mead DJ (1998) Pinus radiata: a narrow endemic from North America takes on the world. Ecology and biogeography of Pinus (ed DM Richardson), pp 432-449. Cambridge University Press, Cambridge, UK. Levine JM, Vilà M, D’Antonio C, Dukes J S, Grigulis k, Lavorel S (2003) Mechanisms underlying the impacts of exotic plant invasions. Proceedings of the Royal Society of London B 270:775-781. Levine JM, Adler PB, Yelenik SG (2004) A meta-analysis of biotic resistance to exotic plant invasions. Ecology Letters 7:975–989. Lind EM, Parker JD (2010) Novel weapons testing: Are invasive plants more chemically defended than native plants? PLoS ONE 5: e10429. Littell RC, Milliken GA, Stroup W, Wolfinger RD (1996) SAS system for mixed models. SAS Institute Inc., Cary, North Carolina, USA. Littell RC, Milliken GA, Stroup WW, Wolfinger RD, Schanbenberber OD (2006) SAS for mixed models, 2nd ed. SAS Press Series, Cary, North Carolina, USA. Lockwood JL, Cassey P, Blackburn T (2005) The role of propagule pressure in explaining species invasions. Trends in Ecology and Evolution 20:223-228. Lombardero M, Vázquez-Mejuto P, Ayres M (2008) Role of plant enemies in the forestry of indigenous vs. nonindigenous pines. Ecological Applications 18:1171–1181. Lonsdale WM (1999) Global patterns of plant invasions and the concept of invasibility. Ecology 80:1522-1536.

151 General references

López-González G (2002) Guía de los árboles y arbustos de la Península Ibérica y Baleares. Ediciones Mundi-Prensa, Madrid, España. Maron J, Vilà M (2001) When do herbivores affect plant invasion? Evidence for the natural enemies and biotic resistance hypothesis. Oikos 95:361-373. Mason WL (2007) Changes in the management of British forests between 1954 and 2000 and possible future trends. Ibis 149:41-62. Masutti L, Battisti A (1990) Thaumetopoea pityocampa (Den. and Schiff.) in Italy. Bionomics and perspectives of integrated control. Journal of Applied Entomology 110:229-234. McCay TS, McCay DH (2009) Processes regulating the invasion of European buckthorn (Rhamnus cathartica) in three habitats of the northeastern United States. Biological Invasions 11:1835–1844. Moles AT, Warton DI, Westoby M (2003) Do small-seeded species have higher survival through seed predation than large-seeded species. Ecology 84:3148-3161. Moles AT, Westoby M (2004) Seedling survival and seed size: as synthesis of the literature. Journal of Ecology 92:372-383. Moles AT, Westoby M (2006) Seed size and plant strategy across the whole life cycle. Oikos 113:91-105. Montero G, Roig S, Martín B, De Miguel J, Alía R (2005) Red de parcelas de introducción de especies del IFIE-INIA (1966-1983). Instituto Nacional de Investigación y Tecnología Agraria y Alimentaria, Ministerio de Educación y Ciencia, Madrid, España. Moreira X, Costas R, Sampedro L, Zas R (2008) A simple method for trapping Hylobius abietis (L.) alive in northern Spain. Investigación Agraria: Sistemas y Recursos Forestales 17:188-192. Moreira X, Sampedro L, Zas R (2009) Defensive responses of Pinus pinaster seedlings to exogenous application of methyl jasmonate: Concentration effect and systemic response. Environmental and Experimental Botany 67:94-100. Morrison WE, Hay ME (2011) Herbivore preference for native vs. exotic plants: Generalist herbivores from multiple continents prefer exotic plants that are evolutionarily naïve. Plos One:e17227 Mortenson SG, Mack RN (2006) The fate of alien conifers in long-term plantings in the USA. Diversity and Distributions 12:456-466. Muth NZ, Pigliucci M (2006) Traits of invasives reconsidered: Phenotypic comparisons of introduced invasive and introduced non-invasive plant species within two closely related clades. American Journal of Botany 93:188-196.

152 General references

Myster RW (2003) Effects of species, density, patch-type, and season on post- dispersal seed predation in a Puerto Rican pasture. Biotropica 35:542-546. Navarro Cerrillo RM, Navarro M, Salas C, González Dugo, Fernández PL, Rodríguez-Silva (1998) Evaluación de grados de afectación producidos por un incendio. Aplicación de imágenes Landsat-TM a su caracterización y seguimiento Servicio de evaluación de recursos naturales, Consejería de medio ambiente (Junta de Andalucía, Andalucía, Spain) y Servicio centralizado de información del territorio, Universidad de Córdoba, Córdoba, Spain. Nepstad DC, Uhl C, Serrao EA (1991) Recuperation of a degraded Amazonian landscape: forest recovery and agricultural restoration. AMBIO 20:248-255 Nichols J (1987) Damage and performance of the green spruce aphid, Elatobium abietinum on twenty spruce species. Entomologia Experimentalis et Applicata 45:211-217. Nordlander G, Bylund H, Orlander G, Wallertz K (2003) Pine weevil population density and damage to coniferous seedlings in a regeneration area with and without shelterwood. Scandinavian Journal of Forest Research 18:191–198. Nuñez M, Simberloff D, Relva M (2008) Seed predation as a barrier to alien conifer invasions. Biological Invasions 10:1389-1398. Nuñez M, Pauchard A (2009) Biological invasions in developing and developed countries: does one model fit all? Biological Invasions 12:707-714. Nuñez MA, Horton TR, Simberloff D (2009) Lack of belowground mutualisms hinders Pinaceae invasions. Ecology 90:2352-2359. Ordóñez J L, Retana J (2004) Early reduction of post-fire recruitment ofPinus nigra by post-dispersal seed predation in different time-since-fire habitats. Ecography 27:449–458. Orians CM, Ward D (2010) Evolution of plant defenses in nonindigenous environments. The Annual Review of Entomology 55:439–59. Oxbrough AG, Gittings T, O’Halloran J, Giller PS, Kelly TC (2006) The influence of open space on ground-dwelling spider assemblages within plantation forests. Forest Ecology and Management 237:404-417. Parry WH (1979) Summer survival of the green spruce aphid, Elatobium abietinum, in North East Scotland. Oecologia 41:235-244. Parry WH (1980) Studies on the factors affecting the population levels of Adelges cooleyi (Gillette) on Douglas fir. 3. Low temperature mortality. Zeitschrift für angewandte Entomologie 90:133-141.

153 General references

Parry WH, Kelly JM, McKillop AR (1990) The role of nitrogen in the feeding strategy of Strophosomus melanogrammus (Forster) (Col., Curculionidae) in a mixed woodland habitat. Journal of Applied Entomology 109:367-376. Pauchard A, Cavieres L, Bustamante R, Becerra P, Rapoport E (2004) Increasing the understanding of plant invasions in southern South America: first Symposium on alien plant invasions in Chile. Biological Invasions 6:255–257. Pedersen L, Bille-Handsen J (1995) Effects of nitrogen load to the forest floor in Sitka spruce stands (Picea sitchensis) as affected by difference in deposition and spruce aphid infestations. Water, Air and Soil Pollution 85:1173-1178. Peñuelas J, Boada M (2003) A global change-induced biome shift in the Montseny mountains (NE Spain). Global Change Biology 9:131-140 Peters SH, Macdonald SE, Boutin S, Moses RA (2004) Postdispersal seed predation of white spruce in cutblocks in the boreal mixedwoods: a short-term experimental study. Canadian Journal of Forest Research 34:907–915. Pimentel D, Zuniga R, Morrison D (2005) Update on the environmental and economic costs associated with alien-invasive species in the United States. Ecological Economics 52:273–288. Price RA, Liston A, Strauss SH (1998) Phylogeny and systematics of Pinus. Ecology and biogeography of Pinus (ed DM Richardson), pp. 49–68 .Cambridge Univ. Press, Cambridge, UK. Pringle A, Bever J, Gardes M, Parrent JL, Matthias C, Rillig MC, Klironomos JN (2009) Mycorrhizal symbioses and plant invasions. The Annual Review of Ecology, Evolution and Systematics 40:699–715. Pulkkinen M (1989) The distribution and ecology of the Strobilomyia flies (Diptera, Anthomyiidae) infesting larch seed and cones in Finland. Annales Entomologici Fennici 55:41 -47. Pyke DA, Thompson JN (1986) Statistical analysis of survival and removal rate experiments. Ecology 67:240-245. Pyšek P, Sádlo J, Mandák B (2002) Catalogue of alien plants of the Czech Republic. Preslia 74:97–186. Pyšek P, Richardson DM, Rejmánek M, Webster G, Williamson M, Kirschner J (2004) Alien plants in checklists and floras: towards better communication between taxonomists and ecologists. Taxon 53:131–143. Pyšek P, Jarošik V (2005) Residence time determines the distribution of alien plants. Invasive Plants: Ecological and Agricultural Aspects (ed S Inderjit), pp. 77–96. Birkhäuser Verlag, Basel.

154 General references

Pyšek P, Richardson DM (2007) Traits associated with invasiveness in alien plants: where do we stand? Biological invasions (ed W Nentwing), pp. 97– 126. Ecological Studies 193, Springer-Verlag, Berlin and Heidelberg. Pyšek P, Richardson DM, Pergl J, Jarošík V, Sixtová Z, Weber E (2008) Geographical and taxonomic biases in invasion ecology. Trends in Ecology Evolution 23:237-244. Quine CP, Humphrey JW, Watts K (2004) Biodiversity in the UK’s forests— recent policy developments and future research challenges. Towards the sustainable use of Europe’s forest ecosystem and landscape research: scientific challenges and opportunities (eds R Paivanen & A Franc), pp. 237–248. European Forest Institute, Joensuu, Finland. Rausher MD (2001) Co-evolution and plant resistance to natural enemies. Nature, 411:857-864. Rejmánek M (1989) Invasibility of plant communities. Biological invasions. A global perspective (eds JA Drake, H Mooney, F di Castri, R Groves, F Kruger, M Rejmánek & M Williamson), pp. 369-388. John Wiley & Sons, Chichester. Rejmánek M (1996) Species richness and resistance to invasions. Diversity and Process in Tropical Forest Ecosystems (eds GHOrians, R Dirzo & JH Cushman), pp. 153-172. Springer-Verlag, Berlin. Rejmánek M, Richardson DM (1996) What attributes make some plant species more invasive? Ecology 77:1655-1661. Rejmánek, M (2000) Invasive plants: approaches and predictions. Australian Ecology 25:497–506. Rey PJ, Alcántara JM (2000) Recruitment dynamics of a fleshy-fruited plant Olea( europaea): connecting patterns of seed dispersal to seedling establishment. Journal of Ecology 88:622-633. Richardson DM, Brown PJ (1986) Invasion of mesic mountain fynbos by Pinus radiata. South African Journal of Botany 52:529-536. Richardson DM, van Wilgen BW (1986) Effects of thirty five years of afforestation with Pinus radiata on the composition of Mesic Mountain Fynbos near Stellenbosch. South African Journal of Botany 52:309-315. Richardson DM, Williams PA, Hobbs RJ (1994) Pine invasions in the Southern Hemisphere: determinants of spread and invadability. Journal of Biogeography 21:511-527. Richardson DM, Bond WJ (1991) Determinants of plant distribution: Evidence from pine invasions. The American Naturalist 137:639-668. Richardson DM, Higgins SI (1998) Pines as invaders in the Southern Hemisphere. Ecology and Biogeography of Pinus (ed DM Richardson), pp. 450–470. Cambridge University Press, Cambridge, UK. 155 General references

Richardson DM (1998) Forestry trees as invasive aliens. Conservation Biology 12:18-26. Richardson DM, Allsopp N, D’Antonio CM, Milton SJ, Rejmánek M (2000) Plant invasions: the role of mutualisms. Biological Review 75:65-93. Richardson DM, Rejmánek M (2004) Conifers as invasive aliens: a global survey and predictive framework. Diversity and Distributions 10:321-331. Richardson DM (2006) Pinus: a model group for unlocking the secrets of alien plant invasions? Preslia 78:375-388. Richarson DM, Pyšek P (2006) Plant invasions: merging the concepts of species invasiveness and community invisibility. Progress in Physical Geography 30:409-431. Richardson DM, van Wilgen BW, Nuñez M (2008) Alien conifer invasions in South America: short fuse burning? Biological Invasions 10:573-577. Rodgerson L (1998) Mechanical defence in seeds adapted for ant dispersal. Ecology 79:1669-1677. Rogers WE, Siemann E (2005) Herbivory tolerance and compensatory differences in native and invasive ecotypes of Chinese tallow tree (Sapium sebiferum). Plant Ecology 181:57–68. Romero H, Jaffe K (1989) A comparison of methods for sampling ants (Hymenoptera, Formicidae) in savannas. Biotropica 21:348–352. Rossiter NA, Settrefield SA, Douglas MM, Hutley LB (2003) Testing the grass-fire cycle: alien grass invasion in the tropical savannas of northern Australia. Diversity and Distribution 9:169–176. Rouget M, Richardson DM, Nek JA, van Wilgen WB (2002) Commercially important trees as invasive aliens- towards spatially explicit risk assessment at a nacional scale. Biological Invasions 4:397-412. Sádlo J, Chytrý M, Pyšek P (2007) Regional species pools of vascular plants in habitats of the Czech Republic. Preslia 79:303–321. Sagnard F, Pichot C, Dreyfus P, Jordano P, Fady B (2007) Modelling seed dispersal to predict seedling recruitment: Recolonization dynamics in a plantation forest. Ecological Modelling 203:464–474. Sampedro L, Moreira X, Martíns P, Zas R (2009) Growth and nutritional response of Pinus pinaster after a large pine weevil (Hylobius abietis) attack. Trees 23:1189–1197. Sampedro L, Moreira X, Llusia J, Peñuelas J, Zas R (2010) Genetics, phosphorus availability and herbivore-derived induction as sources of phenotypic variation of leaf volatile terpenes in a pine species. Journal of Experimental Botany 61:4437-4447.

156 General references

Sampedro L, Moreira X, Zas R (2011a) Resistance and response of Pinus pinaster seedlings to Hylobius abietis after induction with methyl jasmonate. Plant Ecology, doi: 10.1007/s11258-010-9830-x (published online on 29- aug-2010). Sampedro L Moreira X, Zas R (2011b) Costs of constitutive and herbivore-induced chemical defences in pine trees emerge only under low nutrient availability. Journal of Ecology doi: 10.1111/j.1365-2745.2011.01814.x. Sanz-Elorza M, Dana Sánchez E, Sobrino Vesperinas E (2004) Atlas de las Plantas Alóctonas Invasoras en España. Dirección General para la Biodiversidad, pp. 384, Madrid,España . Sarasola MM, Rusch VE, Schlichter TM, Ghersa CM (2006) Invasión de coníferas forestales en áreas de estepa y bosques de ciprés de la cordillera en la Región Andino Patagónica. Austral Ecology 16:143-156. Schaffner U, Ridenour WM, Wolf VC, Bassett T, Müller C, Müller-Schärer H, Sutherland S, Lortie CJ, Callaway RM (2011) Plant invasions, generalist herbivores and novel defense weapons. Ecology 92:829 –835. Schupp EW (1988) Seed and seedling predation in the forest understorey and in treefall gaps. Oikos 51:71-78 Schupp EW (1995) Seed-seedling conflicts, habitat choice, and patterns of plant recruitment. American Journal of Botany 82:399-409. Shafroth PB, Auble GT, Scott ML (1995) Germination and establishment of the native plains cottonwood (Populus deltoides Marshall subsp. Monilifera) and the exotic Russian-olive (Elaeagnus angustigolia L.). Conservation Biology 9:1169-1175. Simberloff D, Relva MA, Nuñez M (2002) Gringos en el bosque: introduced tree invasion in a native Nothofagus/Austrocedrus forest. Biological Invasions 4:35–53. Simberloff D, Nuñez MA, Ledgard NJ, Pauchard A, Richadson DM, Sarasola M, wan Wilgen BW, Zalba SM, Zenni RD, Bustamante R, Peña E, Ziller SR (2010) Spread and impact of introduced conifers in South America: Lessons from other southern hemisphere regions. Austral Ecology 35:489–504. Sol D, Vilà M, I. Kühn I (2008) The comparative analysis of historical alien introductions. Biological Invasions 10:1119-1129. Spies SE, Franklin J (1989) Gap characteristics and vegetation response in coniferous forsts of the Pacific Northwest. Ecology 70:543-545 Staines BW, Welch D, Catt DC, Scott D, Hinge MD (1985) Habitat use and feeding by deer in Sitka spruce plantations. Institute of Terrestrial Ecology Annual Report, ITE, Abbots Ripton.

157 General references

Stowe KA, Marquis RJ, Hochwender CG, Simms EL (2000) The evolutionary ecology of tolerance to consumer damage. Annual Review of Ecology and Systematics 31:565–595. Strauss SY, Campbell OW, Salamin N (2006) Exotic taxa less related to native species are more invasive. Proceedings of the Natural Academy of Sciences of the United States of America 103:5841–5845. Straw N, Fielding N, Green G, Price J (2005) Defoliation and growth loss in young Sitka spruce following repeated attack by the green spruce aphid, Elatobium abietinum (Walker). Forest Ecology and Management 213:349-368. Straw N, Fielding N, Green G, Price J (2006) Seasonal changes in the distribution of green spruce aphid Elatobium abietinum (Walker) (Homoptera: Aphididae) in the canopy of Sitka spruce. Agricultural and Forest Entomology 8:139-154. Tapias R, Climent J, Pardos JA, Gil L (2004) Life history of Mediterranean pines. Plant Ecology 171:53–68. Tellería J (1983) La invernada de aves en los bosques montanos del País Vasco Atlántico. MUNIBE 35:101-108. Thompson K, Band SR, Hodgson JG (1993) Seed size and shape predict persistence in soil. Functional Ecology 7:236-241 Tiberi R, Niccoli A, Curini M, Epifano F, Marcotullio MC, Rosati O (1999) The role of the monoterpene composition in Pinus spp. needles, in host selection by the pine processionary caterpillar, Thaumetopoea pityocampa. Phytoparasitica 27:263-272. Torre I, Arrizabalaga A (2009) Species richness and abundance of small mammals along an elevational gradient of a Mediterranean mountain. Vie et Milieu- Life and Environment 59:203-212. Traveset A, Gulias J, Riera N, Mus M (2003) Transition probabilities from pollination to establishemnt in a rare dioecious shrub species (Rhamnus ludovici-salvatoris) in two habitats. Journal of Ecology 91:427-437. Underwood N, Anderson K, Inouye D (2005) Induced vs. constitutive resistance and the spatial distribution of insect herbivores among plants. Ecology 86:594-602. Van Kleunen M, Dawson W, Schlaepfer D, Jeschke JM, Fischer M (2010) Are invaders different? A conceptual framework of comparative approaches for assessing determinants of invasiveness. Ecology Letters 13: 947–958. Vander Wall SB (1992) The role of animals in dispersing a ‘‘wind-dispersed’’ pine. Ecology, 73:614–621

158 General references

Vander Wall SB, Borchert M, Gworek J (2006) Secondary dispersal of big cone Douglas-fir Pseudotsuga( macrocarpa) seeds. Acta Oecologica 30:100-106. Varbergen AJ, Raymond B, Pearce IS, Watt AD, Hails RS, Hartley SE (2003) Host shifting by Operophtera brumata into novel environments leads to population differentiation in life-history traits. Ecological Entomology 28:604-612. Vilà M, D’Antonio CM (1998) Fitness of invasive Carpobrotus (Aizoaceae) hybrids in coastal California. Ecoscience 5:191-199. Vilà M, Gimeno I (2003) Seed predation of two alien Opuntia species invading Mediterranean communities. Plant Ecology 167:1-8. Vilà M, Lloret F (2000) Seed dynamics of the mast seeding tussock grass Ampelosdesmos mauritanica in Mediterranean shrublands. Journal of Ecology 88:479-491. Vilà M, Maron J, Marco L (2004) Evidence for the enemy release hypothesis in Hypericum perforatum L. Oecologia 142:474-479. Vilà M, Bartomeus I, Gimeno I, Traveset A, Moragues E (2006) Demography of the invasive geophyte Oxalis pes-caprae across a Mediterranean island. Annals of Botany 97: 1055-1062. Vilà M, Basnou C, Pyšek P, Josefsson M, Genovesi P, Gollasch S, Nentwig W, Olenin S, Roques A, Roy D, Hulme P, DAISIE partners (2010) How well do we understand the impacts of alien species on ecosystem services? A pan-European cross-taxa assessment. Frontiers in Ecology and the Environment 8:135-144. Vitousek PM, D’Antonio CM, Loope LL, Rejmánek M, Westbrooks R (1997) Introduced species: a significant component of human-caused global change. New Zealand Journal of Ecology 21:1–16. Walsh PM, Halloran JO, Giller PS, Kelly TC (1999) GreenfinchesCarduelis chloris and other bird species feeding on cones of Noble Fir Abies procera. Bird Study 46:119-121. Watt AD, Evans R, Varley T (1992) The egg laying behaviour of a native insect, the winter moth Operophtera brumata (L.) (Lep., Geometridae), on an introduced tree species, Sitka spruce Picea sitchensis. Journal of Applied Entomology 114:1-4. Welch D, Scott D (1998) Bark-stripping damage by red deer in a Sitka spruce forest in western Scotland. IV. Survival and performance of wounded trees. Forestry 71:225-235. Wilan RL (1985) A guide to forest seed handling with special reference to the tropics. FAO Forestry Paper, no. 20/2, FAO, Rome (Italy).

159 General references

Williamson M (1993) Invaders, weeds and the risk from genetically manipulated organisms. Cellular and Molecular Life Sciences 49:219-224. Williamson M (1996) Biological Invasions. Chapman Hall, London, UK. Willoughby I, Stokes V, Poole J , White J, Hodge S (2007) The potential of 44 native and non-native tree species for woodland creation on a range of contrasting sites in lowland Britain. Forestry 80:531-553. Wilson WL, Day KR (1995) The comparative effectiveness of chemical traps, and fir, spruce and larch billets, for the estimations of pine weevil (Hylobius abietis L.) (Col., Curculionidae) density indices. Journal of Applied Entomology 119:157-160. Wittenberg R, Cock MJW (2001) Invasive alien species: a toolkit of best prevention and management practices. CAB Intrernational, Wallingford. Wittenberg R (2005) An inventory of alien species and their threat to biodiversity and economy in Switzerland. CABI Bioscience Switzerland Centre report to the Swiss Agency for Environment, Forests and Landscape. Zas R, Sampedro L, Prada E, Lombardero MJ, Fernández-López J (2006) Fertilization increases Hylobius abietis L. damage in Pinus pinaster Ait. seedlings. Forest Ecology and Management 222:137–144. Zas R, Sampedro L, Moreira X, Martíns P (2008) Effect of fertilization and genetic variation on susceptibility of Pinus radiata seedlings to Hylobius abietis damage. Canadian Journal of Forest Research 38:63–72.

160 Annexes

161 162 Annexes

Annex 1.

List of references for data sources used in Table 2.1 of Chapter 2.

1- Alonso-Zarazaga MA, Goldarazena A (2005) Presencia en el País Vasco de Rhyephenes humeralis (Coleoptera, Curculionidae), plade de Pinus radiata procedente de Chile. Boletín de la Sociedad Entomológica Aragonesa 36:143-146. 2- Altenkirch V (1986) The Nun Moth (Lymantria monacha L.) in northwestern Germany in 1977 to 1980. Anz Schiidlingskde, Pflanzenschutz, Umweltschutz 59:67-74. 3- Amezaga I (1996) Monterrey pine (Pinus radiata D. Don) suitability for the pine shoot beetle (Tomicus piniperda L.) (Coleoptera: Scolytidae). Forest Ecology and Management 86:73-79. 4- Amezaga I, Rodríguez MA (1998) Resource partitioning of four sympatric bark beetles depending on swarming dates and tree species. Forest Ecology and Management 109:127-135. 5- Angst A, Scheurer S, Forster B (2007) First record of Cinara curvipes (Patch) (Homoptera, Aphidina, Lachnidae) on Abies concolor in Switzerland. Bulletin Société Entomologique Suisse 80:247–252. 6- Annila E (1982) Diapause and population fluctuations inMegastigmus specularis Walley and Megastigmus spermotrophus Wachtl. (Hymenoptera, Torymidae). Annales Entomologici Fennici 48:33-36. 7- Antunes SC, Pereira R, Sousa JP, Santos MC, Gonçalves F (2008) Spatial and temporal distribution of litter arthropods in different vegetation covers of Porto Santo Island (Madeira Archipelago, Portugal). European Journal of Soil Biology 44:45– 56. 8- Arbea JI, Jordana R (1988) Efecto de la repoblación con alerce (Larix kaempferi) en la zona norte de Navarra, sobre la estructura de la poblaciones de colémbolos edáficos. Biología ambiental: actas del Congreso de Biología Ambiental (II Congreso Mundial Vasco), pp.253-265, Pamplona, España. 9- Arnaldo PS, Torres LM (2005) Estudo sobre complexo de parasitóides oófagos de Thaumetopoea pityocampa (Den. and Schiff.) em Portugal. Revista de Biologia (Lisboa) 23:113-120. 10- Arnaldo PS, Torres LM (2006) Effect of different hosts on Thaumetopoea pityocampa populations in Northeast Portugal. Phytoparasitica 34:535- 530.

163 Annexes

11- Arroyo J, Bolger T (2007) Oribatid (Acari: Oribatida) and gamasida (Acari: Gamasida) mite communities in an Irish Sitka spruce Picea sitchensis (Bong.) Carr. stand with three first records for Ireland. Irish Naturalists’ Journal 28:452-458. 12- Bajo J, Santamarıía O, Diez J (2008) Cultural characteristics and pathogenicity of Pestalotiopsis funerea on Cupressus arizonica. Forest Pathology 38:263–274. 13- Barbagallo S, Binazzi A, Ortu S (2005) On the presence in Italy of the nearctic aphid Essigella californica (Essig) living on American pines. Redia 88:79-84. 14- Barbour DA (1985) Two records of Eupithecia abietaria Goeze. Entomologist’s Record 97:146. 15- Binazzi A (1996) Contributions to the knowledge of the conifer aphid fauna. XXVI. Schizolachnus obscurus Börner and S. pineti (Fabricius) in Italy: notes on the morphology and bioecology (Aphididae Lachninae). Redia 79:171- 176. 16- Blick T, Goßner M (2006) Spiders from branch-beating samples in tree crowns (Arachnida: Araneae), with remarks on Cinetata gradata (Linyphiidae) and Theridion boesenbergi (Theridiidae). Arachnologische Mitteilungen 31:23-39. 17- Boag B (1981) Observations on the population dynamics and vertical distribution of trichodorid nematodes in a Scottish forest nursery. Annals of Applied Biology 98:463-469. 18- Boag B (1982) Observations on the population dynamics, life cycle and ecology of the plant parasitic nematode Rotylenchus robustus. Annals of Applied Biology 100:157-165. 19- Broncano M, Vilà M, Boada M (2005) Evidence of Pseudtosuga menziesii naturalization in montane Mediterranean forests. Forest Ecology and Management 211:257-263. 20- Buhn O, Karsholt O, Larsen K, Palm E, Karsten S (1984) Records of Microlepidoptera from Denmark in 1983 (Lepidoptera). Entomologiske Meddelelser 52:1-21. 21- Butterfield J (1997) Carabid community succession during the forestry cycle in conifer plantations. Ecography 20:614-625. 22- Buxton RD (1990) The influence of host tree species on timing of pupation of Thaumetopoea pityocampa Schiff. (Lep., Thaumetopoeidae) and its exposure to parasitism by Phryxe caudata Rond. (Dipt., Larvaevoridae). Journal of Applied Entomology 109:302-310. 23- Carle P, Granet AM, Perrot JP (1979) Contribution a l´etude de la dispersion et de l’agressivité chez Dendroctonus micans Kug. (Col. Scolytidae) en France. Bulletin de la Société Entomologique Suisse 52:185-196.

164 Annexes

24- Caroppo S, Ambrogioni L, Cavalli M, Coniglio D (1998) Occurrence of the pine wood nematodes, Bursaphelenchus spp., and their possible vectors in Italy. Nematologia Mediterranea 26:87-92. 25- Carrascal L, Tellería J (1990) Impacto de las repoblaciones de Pinus radiata sobre la avifauna forestal del norte de España. Ardeola 37:247-266. 26- Carter CI (1976) A gall forming adelgid (Pineus similis (Gill.)) new to Britain with a key to the adelgid galls on Sitka spruce. Entomologist’s Monthly Magazine 111:29-32. 27- Charpin D, Calleja M, Lahoz C, Pichot C, Waisel Y (2005) Allergy to cypress pollen. Allergy 60:293–301. 28- Chytrý M, Maskell L, Pino J, Pyšek P, Vilà M, Font X, Smart S (2008) Habitat invasions by alien plants: a quantitative comparison among Mediterranean, subcontinental and oceanic regions of Europe. Journal of Applied Ecology 45:448–458. 29- Craik C, Van der Kraan A, Van der Kraan M, Austin S (2007) Armyworms-a second site in the British Isles. The Bulletin of the Amateur Entomologists Society 66:241-243. 30- Craik JC, Wormell P, Smith JE, Menzel F (2005) Long columns of “army worms” in west Scotland-the first record ofSciara militaris Nowicki (Diptera, Sciaridae) in the British Isles? Dipterists Digest 12:21-27. 31- Craik JC, Wormell P, Austin S (2006) Armyworms (Dip: Sciaridae) in West Scotland in 2005. Entomologist’s Record and Journal of Variation 118:115-122. 32- Crook CS (1997) A (very) provisional checklist of conifers in the British Isles. In: BSBI NEWS 75 (ed G Ellis), pp. 42-47. Lostock Hall, Preston, Lancaster, UK. 33- Cuadros S (1995) Evaluación de las experiencias sobre adaptación y crecimiento de diversas especies forestales en el montano mediterráneo español. Monte vertiente sur del Parque Natural de Sierra Nevada, Lanjarón (Granada). PhD Thesis, Departamento de Ingeniería Rural, Escuela Técnica Superior de Ingenieros Agrónomos y de Montes, Universidad de Córdoba, España. 34- Cuadros S, Francia JR (1999) Caracterización del sitio de ensayo de especies forestales de Lanjarón, vertiente sur de Sierra Nevada. Aspectos climatológicos y fitoclimáticos. Investigación Agraria, Sistemas y Recursos Forestales 1:143-158. 35- Day K (1984) Systematic differences in the population density of green spruce aphid, Elatobium abietinum in a provenance trial of Sitka spruce, Picea sitchensis. Annals of Applied Biology 105:405-412.

165 Annexes

36- Day K (1984) The growth and decline of a population of the spruce aphid Elatobium abietinum during a three year study, and the changing pattern of fecundity, recruitment and alary polymorphism in a Northern Ireland Forest. Oecologia 64:118-124. 37- Day KR (1986) Population growth and spatial patterns of spruce aphids (Elatobium abietinum) on individual trees. Journal of Applied Entomology 102:505-515. 38- Day KR, Armour HL, Henry CJ (1999) The performance of the green spruce aphid (Elatobium abietinum Walker) on provenances of Sitka spruce (Picea sitchensis (Bong.) Carr.). Colloques de l’INRA; Physiology and genetics of tree-phytophage interactions: International symposium:199-210. 39- Day KR, Docherty M, Leather SR, Kidd NA (2006) The role of generalist insect predators and pathogens in suppressing green spruce aphid populations through direct mortality and mediation of aphid dropping behaviour. Biological Control 38:233-246. 40- Domes R (2003) A new species of the genus Nalepella on Canada Hemlock (Tsuga canadensis (L.) Carr), Pinaceae. Acarologia 43:267-270. 41- Du Merle P, Brunet S, Cornic JF (1992) Polyphagous potentialities of Choristoneura murinana (HB) (Lep., Tortricidae): A “monophagous” folivore extending its host range. Journal of Applied Entomology 113:18-40. 42- Entwistle PF, Adams PH, Evans HF, Rivers CF (1983) Epizootiology of a nuclear polyhedrosis virus (Baculoviridae) in European spruce sawfly Gilpinia( hercyniae): spread of disease from small epicentres in comparison with spread of baculovirus diseases in other hosts. Journal of Applied Ecology 20:473-487. 43- Evans HF (1985) Great spruce bark beetle, Dendroctonus micans: an exotic pest new to Britain. Antenna 9:117 -121. 44- Evans H (1987) Sitka spruce insects: past, present and future. Proceedings of the Royal Society of Edinburgh 93B:157-167. 45- Fabre JP (1997) Geographical distribution of the tortricid moth Epinotia cedricida (Lepidoptera: Tortricidae) within the natural and artificial ranges of the cedar treesCedrus atlantica and C. libani. European Journal of Entomology 94:485-494. 46- Fabre JP, Auger-Rozenberg MA, Chalon A, Boivin S, Roques, A (2004) Competition between exotic and native insects for seed resources in trees of a Mediterranean forest ecosystem. Biological Invasions 6:11–22.

166 Annexes

47- Fabre JP, Chalon A (2005) Multiplication possibilities of an ecotype of the aphid Cedrobium laportei (Homoptera Lachnidae) on various provenances of the genus Cedrus. Entomological Research in MediterraneanForest Ecosystems (eds F Lieutier & D Ghaioule), pp. 123-139. INRA Editions, Versailles Cedex, France. 48- Faccoli M, Battisti A (2001) On the occurrence of Pityophthorus carniolicus Wichmann (Coleoptera, Scolytidae) in Italy. Bulletin de la Société Entomologique Suisse 74:7-11. 49- Fernandez FA, Evans PR, Dunstone, N (1994) Local variation in rodent communities of Sitka spruce plantations: the interplay of successional stage and site-specific habitat parameters. Ecography 17:305-313. 50- Fernandez FA, Evans PR, Dunstone N (1996) Population dynamics of the wood mouse Apodemus sylvaticus (Rodentia: Muridae) in a Sitka spruce successional mosaic. Journal of Zoology London 239:717-730. 51- Fernandez FA, Dunstone N, Evans PR (1999) Density-dependence in habitat utilisation by wood mice in a Sitka spruce successional mosaic: the roles of immigration, emigration, and variation among local demographic parameters. Canadian Journal of Zoology 77:397-405. 52- Friberg N, Jacobsen D (1994) Feeding plasticity of two detritivore-shredders. Freshwater Biology 32:133-142. 53- Goßner M, Gruppe A, Simon U (2005) Aphidophagous insect communities in tree crowns of the neophyte Douglas-fir Pseudotsuga[ menziesii (Mirb.) Franco] and Norway spruce (Picea abies L.). Journal of Applied Entomology 129(2):81-88. 54- Gossner M, Liston A, Späth J (2007) Sawflies in the crowns of native and exotic trees, sampled with flight-interception traps in southern Germany (Hymenoptera: Symphyta). Entomologia Generalis 30:273-282. 55- Gurnell J, Lurz PW, Shirley MD, Cartmel S, Garson PJ, Magris L, Steeles J (2004) Monitoring red squirrels Sciurus vulgaris and grey squirrels Sciurus carolinensis in Britain. Mammal Review 34:54-74. 56- Habermann M (2000) The larch casebearer and its host tree: I. Population dynamics of the larch casebearer (Coleophora laricella Hbn.) from latent to outbreak density in the field. Forest Ecology and Management 136:11-22. 57- Hadincová V, Münzbergová Z, Wild J, Šajtar L, Marešová J (2008) Dispersal of invasive Pinus strobus in sandstone areas of Czech Republic. Plant Invasions: Human perception, ecological impacts and management (eds B Tokarsha-Guzik, JH Brock, G Brundu, L Child , CC Daehler & P Pyšek ), pp. 117-132. Backhuys Publishers, Leyden, The Netherlands.

167 Annexes

58- Hanzélyová D (1998) A comparative study of Pinus strobus L. and Pinus sylvestris L.: growth at different soil acidities and nutrient levels. Plant invasions: ecological mechanisms and human responses (eds U Starfinger, K Edwards, I Kowarik & M Williamson), pp. 185-194. Backhuys Publishers, Leiden, The Netherlands. 59- Harriman R, Morrison BR (1982) Ecology of streams draining forested and non forested catchments in an area of central Scotland subject to acid precipitation. Hydrobiologia 88:251-263. 60- Heie OE (1989) Aphids in Denmark in the spring following the mild winter of 1988-89. Entomologiske Meddelelser 57:173-175. 61- Holusa J, Grodzki W (2008) Occurrence of Ips duplicatus (Coleoptera: Curculionidae, Scolytinae) on pines (Pinus sp.) in the Czech Republic and southern Poland – Short Communication. Journal of Forest Science 54:234–236. 62- Hunter MD, Watt AD, Docherty M (1991) Outbreaks of the winter moth on Sitka spruce in Scotland are not influenced by nutrient deficiencies of trees, tree budburst, or pupal predation. Oecologia 86:62-69. 63- Irizar I, Laskurain N, Herrero J (2004) Wild boar frugivory in the atlantic Basque Country. Galemys 16:125-133. 64- Jarry M, Candau JN (1997) Impact of emigrating seed chalcid, Megastigmus spermotrophus Watchl (Hymenoptera: Torymidae), on seed production in a Douglas-fir seed orchard in France and modelling of orchard invasion. The Canadian Entomologist:7-19. 65- Kohlert A, Roth M (2000) The effects of exotic tree species (Douglas fir: Pseudotsuga menziesii) on litter decomposing arthropods and soil surface dwelling regulator. Mitteilungen der Deutschen Gesellschaft für allgemeine Entomologie 12:71-74. 66- Kowarik I (2005) Urban ornamentals escaped from cultivation. Crop Ferality and Volunteerism (ed J Gressel), pp. 97-121. CRC Press, Boca Raton. 67- Krivánek M, Pyšek P, Jarošik V (2006) Planting history and propagule ressure as predictors of invasion by woody species in a temperate region. Conservation Biology 20:1487–1498. 68- Lavery PB, Mead DJ (1998) Pinus radiata: a narrow endemic from North America takes on the world. Ecology and Biogeography of Pinus. (ed DM Richardson), pp. 432-449. Cambridge University Press, Cambridge, UK. 69- Lombardero M, Vázquez-Mejuto P, Ayres M (2008) Role of plant enemies in the forestry of indigenous vs. nonindigenous pines. Ecological Applications 18:1171–1181.

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70- Mailleux AC, Roques A, Molenberg JM, Gregoire JC (2008) A North American invasive seed pest, Megastigmus spermotrophus (Wachtl) (Hymenoptera: Torymidae): Its populations and parasitoids in a European introduction zone. Biological Control 44:137–141. 71- Mason WL (2007) Changes in the management of British forests between 1954 and 2000 and possible future trends. Ibis 149:41-62. 72- Masutti L, Battisti A (1990) Thaumetopoea pityocampa (Den. and Schiff.) in Italy. Bionomics and perspectives of integrated control. Journal of Applied Entomology 110:229-234. 73- McKenzie AJ, Petty SJ, Toms MP, Furness RW (2007) Importance of Sitka Spruce Picea sitchensis seed and garden bird-feeders for Siskins Carduelis spinus and coal tits Periparus ater. Bird Study 54:236-247. 74- Moore R, Brixey JM, Milner AD (2004) Effect of time of year on the development of immature stages of the Large Pine Weevil (Hylobius abietis L.) in stumps of Sitka spruce (Picea sitchensis Carr.) and influence of felling date on their growth, density and distribution. Journal of Applied Entomology 128:167-176. 75- Morral LG (1992) Trissolcus flavipes, an egg parasitoid of Pentatomidae (Heteroptera), new for Western Europe (Hymenoptera: Scelionidae). Entomologische Berichten, Amsterdam 52:84-86. 76- Morrison BR, Wells DE (1981) The fate of fenitrothion in a stream environment and its effect on the fauna, following aerial spraying of a Scottish forest. The Science of the Total Environment 19:233-252. 77- Mouna M, Fabre JP (2005) Pest insects of cedars: Cedrus atlantica Manetti, C. libani A. Richard and C. brevifolia Henry in the Mediterranean area. Entomological Research in Mediterranean Forest Ecosystems (eds F Lieutier & D Ghaioule), pp. 89-104. INRA Editions, Versailles Cedex, France. 78- Nichols J (1987) Damage and performance of the green spruce aphid, Elatobium abietinum on twenty spruce species. Entomologia Experimentalis et Applicata 45:211-217. 79- Nielsen C, Eilenberg J, Harding S, Vestergaard S (2004) Biological control of weevils (Strophosoma melanogrammum and S. capitatum) in greenery plantations in Denmark. Department of Ecology, The Royal Veterinary and Agricultural University. Pesticides Research 91. 80- Olbrechtová J (2007) First occurrence of Cinara curvipes Patch 1912 on white fir Abies( concolor (Gordon et Glendinning) Hildebrand 1861) in the Czech Republic. Monograph aphids and other hemipterous insects (ed E Cichocka), 13. The John Paul II Catholic University of Lublin, Department of Nature Preservation, Lublin, Poland. 169 Annexes

81- Oxbrough AG, Gittings T, O’Halloran J, Giller PS, Kelly TC (2005) Structural indicators of spider communities across the forest plantation cycle. Forest Ecology and Management 212:171-183. 82- Oxbrough AG, Gittings T, O’Halloran J, Giller PS, Kelly TC (2006) The influence of open space on ground-dwelling spider assemblages within plantation forests. Forest Ecology and Management 237:404-417. 83- Oxbrough AG, Gittings T, O’Halloran J, Giller PS, Kelly TC (2006) The initial effects of afforestation on the ground-dwelling spider fauna of Irish peatlands and grasslands. Forest Ecology and Management, 237:478-491. 84- Parry WH (1979) Factors affecting low temperature survival of Cinara pilicornis eggs on Sitka spruce. International Journal of Biometeorology 23:185-193. 85- Parry WH (1979) Summer survival of the green spruce aphid, Elatobium abietinum, in North East Scotland. Oecologia 41:235-244. 86- Parry WH (1980) Studies on the factors affecting the population levels of Adelges cooleyi (Gillette) on Douglas fir. 3. Low temperature mortality. Zeitschrift für angewandte Entomologie 90:133-141. 87- Parry WH, Kelly JM, McKillop AR (1990) The role of nitrogen in the feeding strategy of Strophosomus melanogrammus (Forster) (Col., Curculionidae) in a mixed woodland habitat. Journal of Applied Entomology 109:367-376. 88- Pedersen L, Bille-Handsen J (1995) Effects of nitrogen load to the forest floor in Sitka spruce stands (Picea sitchensis) as affected by difference in deposition and spruce aphid infestations. Water, Air and Soil Pollution 85:1173-1178. 89- Petanovic R (1999) Three new species of Eriophyoid mites (Acari: Eriophyoidea) from Serbia with the notes on new taxa for the fauna of Yugoslavia. Acta Entomologica Serbica 4:127 -137. 90- Picozzi N, Catt DC, Moss R (1992) Evaluation of capercaillie habitat. Journal of Applied Ecology 29:751-762. 91- Pinder PS, Hayes AJ (1986) An outbreak of vapourer moth (Orgyia antiqua L.: Lepidoptera Lymantridae) on Sitka spruce (Picea sitchensis (Bong.) Carr.) in Central Scotland. Forestry 59:97-105. 92- Poljakovic-Pajnik L, Petrovic-Obradovic O (2002) Bow-legged fir aphid Cinara curvipes (Patch) new pest of Abies concolor in Serbia. Acta Entomologica Serbica 7:147–150. 93- Pulkkinen M (1989) The distribution and ecology of the Strobilomyia flies (Diptera, Anthomyiidae) infesting larch seed and cones in Finland. Annales Entomologici Fennici 55:41-47.

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94- Putman RJ, Edwards PJ, Mann JCE, How RC, Hill SD (1989) Vegetational and faunal changes in an area of heavily grazed woodland following relief of gazing. Biological Conservation 47:13-32. 95- Pyšek P, Sádlo J, Mandák B (2002) Catalogue of alien plants of the Czech Republic. Preslia 74:97–186. 96- Ramos J, Bugalho M, Cortez P, Iason G (2006) Selection of trees for rubbing by red and roe deer in forest plantations. Forest Ecology and Management 222:39–45. 97- Raymond B, Hartley SE, Cory JS, Hails RS (2005) The role of food plant and pathogen-induced behaviour in the persistence of a nucleopolyhedrovirus. Journal of Invertebrate Pathology 88:49-57. 98- Raymond B, Hails RS (2007) Variation in plant resource quality and the transmission and fitness of the winter moth,Operophtera brumata nucleopolyhedrovirus. Biological Control 41:237-245. 99- Robinson CH, Ineson P, Piearce TG, Parrington J (1996) Effects of earthworms on cation and phosphate mobilisation in limed peat soils under Picea sitchensis.Forest Ecology and Management 86:253-258. 100- Robinson CH, Ineson P, Piearce TG,and Rowland AP (1992) Nitrogen mobilization by earthworms in limed peat soils under Picea sitchensis. Journal of Applied Ecology 29:226-237. 101- Sánchez ME, Gibbs JN (1995) The ecology of fungal cankers on Cupressus macrocarpa in southern England. European Journal of Forest Pathology 252:266-273. 111- Staines BW, Welch D, Catt DC, Scott D, Hinge MD (1985) Habitat use and feeding by deer in Sitka spruce plantations. Institute of Terrestrial Ecology Annual Report:12-16. 102-Schneider VI (1985) Gnathotrichus materiarius Fitch (Col., Scolytidae) in pheromone traps of the European Larch Bark Beetle, lps cembrae (Heer) (Col., Scolytidae), a new location for Europe in North West Germany. Anz Schiidlingskde, Pflanzenschutz, Umweltschutz 58:51-55. 103-Shaposhnikov GK, Gabrid NV (1987) Aphids of the Family Hormaphididae (Homoptera, Aphidinea) from the coniferous trees. Entomologicheskoe Obozrenie 66:765-772. 104- Shaw G (1990) Timing and fidelity of breeding for SiskinsCarduelis spinus in Scottish conifer plantations. Bird Study 37:30-35. 105- Shaw G (1996) Dunnock feeding in spruce canopy. Scottish Birds 8:185. 106- Shaw P, Ozanne C, Speight M, Palmer I (2007) Edge effects and arboreal Collembola in coniferous plantations. Pedobiologia 51:287-293.

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107-Skuhravy V, Thuróczy C (2007) Parasitic Hymenoptera associated with Thecodiplosis brachyntera (Diptera: Cecidomyiidae) on the genus Pinus (Pinaceae) in the Czech Republic. Journal of Forest Science 53: 381–389. 108- Smith GF, Gittings T, Wilson M, French L, Oxbrough A, O’Donoghue S, O’Halloran J, Kelly D, Mitchell FJ, Kelly T, Iremonger S, Mckee AM, Giller, P (2008) Identifying practical indicators of biodiversity for stand-level management of plantation forests. Biodiversity and Conservation 17:991-1015. 109- Smith R, Baker R, Malumphy C, Hockland S, Hammon R, Ostojá- Starzewski J, Collins D (2007) Recent non-native invertebrate plant pest establishments in Great Britain: origins, pathways, and trends. Agricultural and Forest Entomology 9:307–326. 110- Spradbery JP, Kirk A (1981) Experimental studies on the responses of European siricid woodwasps to host trees. Annals of Applied Biology 98:179-185. 111- Staines BW, Welch D, Catt DC, Scott D, Hinge MD (1985) Habitat use and feeding by deer in Sitka spruce plantations. Institute of Terrestrial Ecology Annual Report:12-16. 112- Staines BW, Petty SJ, Ratcliffe PR (1987) Sitka spruce (Picea sitchensis (Bong.) Carr.) forests as a habitat for birds and mammals. Proceedings of the Royal Society of Edinburgh 93B:169-181. 113- Straw N, Fielding N, Green G, Price J (2005) Defoliation and growth loss in young Sitka spruce following repeated attack by the green spruce aphid, Elatobium abietinum (Walker). Forest Ecology and Management 213:349-368. 114- Straw N, Fielding N, Green G, Price J (2006) Seasonal changes in the distribution of green spruce aphid Elatobium abietinum (Walker) (Homoptera: Aphididae) in the canopy of Sitka spruce. Agricultural and Forest Entomology 8:139-154. 115- Straw NA, Green G (2001) Interactions between green spruce aphid (Elatobium abietinum (Walker)) and Norway and Sitka spruce under high and low nutrient conditions. Agricultural and Forest Entomology 3:263-274. 116- Talavera M, Magunacelaya JC, Tobar A (1999) Plant parasitic nematodes from a forest tree nursery in southern Spain with some notes about the influence of soil storage on the quantitative recovery of Meloidogyne arenaria. Nematology 1(3):261-266. 117- Tellería J (1983) La invernada de aves en los bosques montanos del País Vasco Atlántico. MUNIBE 35:101-108.

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118- Varbergen AJ, Raymond B, Pearce IS, Watt AD, Hails RS, Hartley SE (2003) Host shifting by Operophtera brumata into novel environments leads to population differentiation in life-history traits. Ecological Entomology 28:604-612. 119- Vasconcelos T, Horn A, Lieutier F, Branco M, Kerdelhué C (2006) Distribution and population genetic structure of the Mediterranean pine shoot beetle Tomicus destruens in the Iberian Peninsula and Southern France. Agricultural and Forest Entomology 8:103–111. 120- Wainhouse D, Ashburner R, Ward E, Boswell R (1998) The effect of lignin and bark wounding on susceptibility of spruce trees to Dendroctonus micans. Journal of Chemical Ecology 24:1551-1561. 121- Walsh PM, Halloran JO, Giller PS, Kelly TC (1999) GreenfinchesCarduelis chloris and other bird species feeding on cones of Noble Fir Abies procera. Bird Study 46:119-121. 122- Warrington S, Whittaker JB (1990) Interactions between Sitka Spruce, the green spruce aphid, sulphur dioxide pollution and drought. Environmental Pollution 65:363-370. 123- Watt AD, McFarlane AM (1991) Winter moth on Sitka spruce: synchrony of egg hatch and budburst, and its effect on larval survival. Ecological Entomology 16:387-390. 124- Watt AD, Evans R, Varley T (1992) The egg laying behaviour of a native insect, the winter moth Operophtera brumata (L.) (Lep., Geometridae), on an introduced tree species, Sitka spruce Picea sitchensis. Journal of Applied Entomology 114:1-4. 125- Weir AG, Ferns PN, Cowie RJ (1996) Manipulating seed production in commercial conifer plantations and the effect on vertebrate seed predators. Aspects of Applied Biology 44:17-24. 126- Welch D, Scott D (1998) Bark-stripping damage by red deer in a Sitka spruce forest in western Scotland. IV. Survival and performance of wounded trees. Forestry 71:225-235. 127- Williamson K (1971) Birdlife associated with a pure stand of Japanese larch. Bird Study 18:225-226. 128- Willoughby I, Stokes V, Poole J , White J, Hodge S (2007) The potential of 44 native and non-native tree species for woodland creation on a range of contrasting sites in lowland Britain. Forestry 80:531-553. 129- Wilson WL, Day KR (1995) The comparative effectiveness of chemical traps, and fir, spruce and larch billets, for the estimations of pine weevil Hylobius( abietis L.) (Col., Curculionidae) density indices. Journal of Applied Entomology 119:157-160. 173 Annexes

130- Wilson MW, Pithon J, Gittings T, Kelly TC, Giller PS, O'Halloran J (2006) Effects of growth stage and tree species composition on breeding bird assemblages of plantation forests. Bird Study 53:225-236. 131- Wittenberg R (2005) An inventory of alien species and their threat to biodiversity and economy in Switzerland. CABI Bioscience Switzerland Centre report to the Swiss Agency for Environment, Forests and Landscape.

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