Ecología de la especie invasora Ailanthus altissima (Mill

Ecología de la especie invasora Ailanthus
altissima (Mill.) Swingle. Bases para su control
y
erradicación
en
Espacios
Naturales
Protegidos.
Soraya D. Constán Nava.
Facultad de Ciencias
Departamento de Ecología
TESIS DOCTORAL
Ecología de la especie invasora Ailanthus altissima (Mill.) Swingle.
Bases para su control y erradicación en Espacios Naturales Protegidos.
Soraya D. Constán Nava
Director: Andreu Bonet Jornet
2012
Departamento de Ecología
Ecología de la especie invasora Ailanthus altissima (Mill.) Swingle.
Bases para su control y erradicación en Espacios Naturales Protegidos.
Memoria presentada por:
Soraya D. Constán Nava
para optar al titulo de Doctora en Ciencias Biológicas
Director:
Andreu Bonet Jornet
Alicante, 2012
Dr. Andreu Bonet Jornet, Profesor titular del Departamento de Ecología de la Facultad
de Ciencias de la Universidad de Alicante
HACE CONSTAR:
Que el trabajo presentado en esta memoria, con el titulo: “Ecología de la especie
invasora Ailanthus altissima (Mill.) Swingle. Bases para su control y erradicación en
Espacios Naturales Protegidos” ha sido realizado bajo su dirección por Soraya D.
Constán Nava en el Departamento de Ecología y reúne todos los requisitos para su
aprobación como Tesis Doctoral.
Alicante, Enero de 2012
D. Andreu Bonet Jornet
A Santi, mi mundo
A mi familia, los que están y los que se fueron
ÍNDICE
Resumen
Introducción. Objetivos y estructura de la tesis. Metodología y área de
1
estudio. Resultados generales. Discusión general. Conclusiones.
Bibliografía
Capitulo 1
Distribution and performance of populations of the invasive species
49
Ailanthus altissima on Mediterranean Protected Areas
Capitulo 2
Genetic variability modulates the effect of habitat−type and
69
environmental conditions on early invasion success of Ailanthus
altissima in Mediterranean ecosystems
Capitulo 3
Direct and indirect effects of invasion by the alien tree Ailanthus
altissima
on
riparian
plant
communities
and
99
ecosystem
multifunctionality
Capitulo 4
Long-term control of the invasive tree Ailanthus altissima: Insights 133
from Mediterranean protected forests
Agradecimientos
153
Afiliación de los coautores
155
RESUMEN
Introducción. Objetivos y estructura de la tesis. Metodología
y área de estudio. Resultados generales. Discusión general.
Conclusiones. Bibliografía
2
INTRODUCCIÓN
DEFINICIÓN. ANTECEDENTES. FASES GENERALES
Las especies exóticas invasoras (sensu Richardson et al. 2000) son aquellas especies
capaces de sobrevivir, establecerse y reproducirse en un lugar biogeográficamente
distinto al original, superando los limitantes a su llegada, establecimiento y
reproducción, tanto bióticos como abióticos (Williamson 1996; Mack et al. 2000;
Balaguer 2004). Las invasiones biológicas por especies vegetales obedecen a un
fenómeno natural, pero la presencia del ser humano ha acelerado este proceso
exponencialmente (Vitousek et al. 1997; Mack et al. 2000; Mooney y Hobbs 2000; Vilà
2000). Actualmente, se consideran como la tercera causa principal del cambio global,
solamente por detrás de la destrucción de hábitats y la fragmentación del paisaje
(Vitousek 1990; Vitousek 1994; Williamson 1996), causando un enorme impacto sobre
la biodiversidad a escala global (IUCN 2000; Mack et al. 2000; Didham et al. 2005). La
introducción de especies vegetales ha sido tanto voluntaria como involuntaria por parte
del ser humano, quien ha actuado como elemento diseminador de primer orden
(antropocoria) (Vitousek et al. 1997; Vilà 2000; Sanz et al. 2001; Hulme et al. 2008).
Entre las vías de introducción se encuentran el turismo, las migraciones y el
desplazamiento y los intercambios comerciales, que aunque fueron iniciados
principalmente en la Edad Moderna a partir del descubrimiento de las rutas
transoceánicas, actualmente son de carácter global (McNeely et al. 2001; Vilà 2000;
Kowarik 2005).
El proceso de invasión está considerado como un continuum invasiónnaturalización en el que las especies han de atravesar diferentes tipos de barreras:
geográficas, ambientales, reproductivas, de dispersión, ambientales en hábitats
perturbados y finalmente, ambientales en hábitats naturales (Richardson et al. 2000;
Pyšek y Richardson 2010). Solo unas pocas de las especies introducidas (alrededor del
10%) sobreviven a las nuevas condiciones a las que se enfrentan y pocas son las que se
naturalizan, estableciéndose y reproduciéndose en el nuevo lugar, de manera que el
número de individuos aumenta y la población comienza a expandirse (Williamson
1996; Mack et al. 2000; Balaguer 2004). En el caso de las especies invasoras, esta
explosión demográfica normalmente se caracteriza por una alta tasa de crecimiento
poblacional, es decir, el tamaño poblacional y la superficie ocupada crecen rápidamente
3
(McNeely et al. 2001). La mayoría de las especies naturalizadas no causan alteración en
su nueva área de distribución, pero un 1 % se vuelven invasoras (Williamson y Fitter
1996; Mack et al. 2000; Mooney y Hobbs 2000; Balaguer 2004), estableciendo
interacciones ecológicas y evolutivas con la biocenosis de la comunidad invadida (Vilà
2000).
La invasión no solo depende de la especie introducida, sino que es el resultado
de la interacción compleja con diferentes factores, incluyendo el clima y el hábitat, así
como las especies nativas y otras especies invasoras que pueden acelerar y amplificar la
invasión y sus efectos en el lugar invadido a través de procesos de retroalimentación
(proceso conocido como invasional meltdown, Simberloff y Von Holle 1999;
Richardson et al. 2000; Pyšek y Richardson 2010).
LA ECOLOGÍA DE LA INVASIÓN. MARCO CONCEPTUAL
La ecología de la invasión estudia todos los componentes y procesos que interfieren en
la invasión de especies exóticas (Mack et al. 2000; Richardson y Pyšek 2006; Pyšek y
Richardson 2010). Desde el libro “The ecology of invasions by animals and plants”
(Elton 1958) hasta la actualidad, se puede encontrar numerosa literatura científica sobre
estos componentes (véase por ejemplo: Lodge 1993; Richardson et al. 2000; Pyšek et
al. 2004; Blackburn et al. 2011). Foxcroft et al. (2011) han propuesto un marco
conceptual que integra todos estos mecanismos y el contexto en el que interactúan (Fig.
1). Este marco general está compuesto por tres componentes principales: las
características de las especies invasoras, el contexto donde se desarrolla la invasión y la
susceptibilidad del hábitat receptor.
4
Figura 1 Marco conceptual de la ecología de la invasión. Adaptado de Foxcroft et al. 2011
Respecto al primer componente, las características de una especie vegetal
invasora, se encuentran numerosos mecanismos que pueden potenciar el carácter
invasor de una especie, tales como la producción de propágulos para el éxito de
reproducción (numero y viabilidad de semillas, tasa de germinación, crecimiento de
plántulas), competitividad (por ejemplo, presencia de sustancias alelopáticas),
reproducción vegetativa para expandirse, genotipos con elevada plasticidad fenotípica
ya que se propagan por diferentes medios, y una elevada tasa de crecimiento (Rejmánek
1996; Vilà 2000; Castro-Díez et al. 2004; Foxcroft et al. 2011). Pero, no todas las
especies vegetales invasoras poseen todas estas características, ni tampoco una especie
con esas características actúa como invasora (Vilà 2000; Balaguer 2004). Existen
diversos estudios comparativos entre especies vegetales nativas vs invasoras (e.g.
Godoy et al. 2009; van Kleunen et al. 2010; Godoy et al. 2011) donde se ha encontrado
que existen diferencias entre características relacionadas con el desarrollo (tasa de
crecimiento, tamaño) siendo mayor en las invasoras, pero no hay diferencias en
plasticidad fenotípica (por ejemplo, como respuesta a gradientes de nutrientes y luz). En
este sentido, Thompson y Davis (2011) han concluido que las plantas invasoras poseen
5
rasgos característicos de las especies más exitosas, independientemente si son nativas o
exóticas.
El segundo componente es el contexto del sistema, es decir, dentro del cual se
desarrolla la invasión. Este componente conecta las características de la especie
invasora con la susceptibilidad del hábitat. Entre los mecanismos específicos que se
incluyen en este componente se encuentran:
− la conectividad entre hábitats. Por ejemplo, las zonas verdes, vertederos, áreas
residenciales y parques periurbanos, dado que están sujetas al estrés ambiental y a
causa de su proximidad a sitios de introducción de especies exóticas y su función
como borde de otros hábitats, son el principal hábitat para las especies invasoras,
las cuales se extienden hacia hábitats menos urbanos (Kowarik 1983; Gulezian et
al. 2010; Hitchmugh 2011).
− la presencia de vías de expansión que actúan como vías de dispersión, por
ejemplo, debido al movimiento de semillas (Spellerberg 1998; Vilà y Pujadas,
2001; Hansen y Clevenger 2005; Kowarik y von deer Lippe 2006)
− la presión de propágulos, es decir, nuevos eventos de introducción de individuos
y su densidad (Von Holle y Simberloff 2005; Lockwood et al. 2009).
− el tiempo de residencia (es decir, cuánto tiempo lleva una especie invasora en un
determinado lugar (Lockwood et al. 2009; Vilà y Ibáñez 2011)
− los valores y percepciones humanos, importante de cara a la gestión de las
especies invasoras (Vitousek et al. 1997; Simberloff, 2009; Foxcroft et al. 2011).
El tercer componente es la susceptibilidad del hábitat, y los mecanismos
específicos que entran dentro de este componente están relacionados con las
condiciones bioclimáticas, la presencia de depredadores (generalmente baja, pues
escapan de sus enemigos naturales), la disponibilidad de recursos y la heterogeneidad
del paisaje, que incluye la presencia de espacios vacíos (áreas perturbadas de origen
antrópico) (Hobbs y Huenneke 1992; Maron y Vilà 2001; Sanz et al. 2004).
EFECTOS
DE LAS ESPECIES VEGETALES EXÓTICAS INVASORAS
La presencia de especies vegetales invasoras supone nuevas situaciones ambientales
tanto para las especies invasoras como para el ecosistema receptor (Castro-Díez et al.
2004), de manera que cambian las reglas de la existencia para las especies (Vitousek et
al. 1997). Los efectos que causan estas especies ocurren a diferentes niveles.
6
Genéticamente pueden afectar mediante la hibridación de especies exóticas con
nativas (Huxel 1999; Ellstrand y Schierenbeck 2000; Vilà 2000; Lee 2002).
En el ámbito paisajístico afectan a través de la homogeneización de los hábitats
por la dominancia de especies exóticas invasoras (Vilà 2000; Hastings et al. 2005;
Vuilleumier et al. 2011).
Ecológicamente se encuentran la alteración de la estructura y funcionamiento del
ecosistema invadido, mediante la alteración de ciclos biogeoquímicos, modificación de
procesos de erosión y sedimentación, alteración de la fertilidad de los suelos, reducción
o agotamiento de los niveles de agua, alteración de los patrones de drenaje, y/o
modificación de los regímenes de incendios (Mack et al. 2000; Corbin y D´Antonio 2004;
Didham et al. 2005; Castro-Díez et al. 2009). Asimismo, pueden competir con especies
nativas, por ejemplo, mediante mecanismos competitivos que no están presentes en las
comunidades donde invaden como es a través de sustancias alelopáticas (Callaway y
Aschehoug 2000; Ens et al. 2009), o por competencia por polinizadores (Jakobsson et
al. 2009; Morales y Traveset 2009). Esta competencia puede llegar a reemplazar a las
especies nativas (Vilà 2000; Simberloff 2001). Por otro lado, pueden favorecer la
invasión de otras especies exóticas y causar un impacto sobre la regeneración y
dinámica natural de muchos ecosistemas terrestres (Groves 1986; Mooney y Drake
1986; Simberloff 2001; Valladares et al. 2004; Murphy et al. 2006).
En el ámbito socioeconómico implican elevados costes debido a las inversiones
realizadas en medidas de prevención (análisis de riesgos, educación ambiental),
detección temprana (localización de especies invasoras, evaluación de impactos),
control y erradicación, así como en medidas de restauración de ecosistemas invadidos
(Dukes y Mooney 1999; Mack et al. 2000; Vilà 2000; Castro-Díez et al. 2004; Moore
2005). Por ejemplo, en Europa se ha estimado que el coste económico en relación a las
especie invasoras es de 9-12 billones de euros anuales, siendo realmente mayor, pues en
muchos países están comenzando a analizar los efectos de este problema (European
Commission 2008).
En el ámbito cultural implican una alteración de la percepción del ser humano,
ya que en muchos casos las especies invasoras son consideradas como nativas (Balaguer
2004; Davis et al. 2011), por ejemplo, debido al uso común de especies exóticas en
jardinería.
7
MARCO LEGISLATIVO SOBRE ESPECIES EXÓTICAS INVASORAS
En la actualidad existe un gran interés en el desarrollo de normativas adecuadas,
manifestada en diferentes mandatos internacionales de carácter ambiental, que regulen
específicamente las especies exóticas invasoras. Hasta el momento se puede encontrar
leyes que instan al control, prevención y protección frente a las especies invasoras.
En el ámbito europeo, el Convenio de Berna, uno de los primeros antecedentes
legales relativo a la conservación de la vida silvestre y el medio natural en Europa
(1979), obliga a las partes contratantes a realizar un control estricto de la introducción
de especies no nativas (Art. 11, 2b). Posteriormente, el Reglamento (CE) n.º 338/97 del
Consejo, de 9 de diciembre de 1996, relativo a la protección de especies de la fauna y
flora silvestres mediante el control de su comercio, incluye especies que son una
amenaza ecológica para las especies silvestres autóctonas dentro de la Unión Europea.
Posteriormente, dichas disposiciones son recogidas y ampliadas en la Directiva
92/43/CEE, del Consejo, de 21 de mayo de 1992, relativa a la conservación de los
hábitats naturales y de la fauna y la flora silvestres, incluyendo la regulación de la
introducción de especies exóticas. Asimismo, Comisión Europea (2008) adopta la
Comunicación «Hacia una Estrategia de la Unión Europea sobre especies invasoras»
(COM(2008) 789 final).La estrategia internacional queda recogida en el Convenio de
Naciones Unidas sobre la Diversidad Biológica (1993), la cual manda a las partes
firmantes a prevenir la introducción, control o erradicación de las especies no nativas.
En el ámbito nacional, la Ley 4/1989 de 27 de marzo, de Conservación de los
Espacios Naturales y de la Flora y Fauna Silvestre dicta normas en relación a la
introducción o liberación no autorizada de especies alóctonas perjudiciales,
considerándose como delito contra el medio ambiente en la Ley orgánica 10/1995, de
23 de noviembre, del Código Penal. Asimismo, la Ley 43/2002, de 20 de noviembre de
Sanidad Vegetal incluye restricciones y prohibiciones relacionadas con especies
vegetales alóctonas. La Ley 26/2007, de 23 de octubre, de responsabilidad
medioambiental, a través del Real Decreto 2090/2008, de 22 de diciembre, de desarrollo
parcial de dicha Ley, incluye a las especies vegetales invasoras como agente causante de
daño biológico. La Ley 42/2007, de 13 de diciembre, de Patrimonio Natural y de la
Biodiversidad, contempla la prohibición de la introducción de especies alóctonas por
parte de las administraciones públicas competentes en el caso de que puedan afectar a
las especies nativas (art. 52.5) y crea el Catálogo Español de Especies Exóticas
8
Invasoras (art. 61.1) de ámbito estatal. Finalmente, el Real Decreto 1628/2011, de 14 de
noviembre, regula el listado y Catálogo Español de Especies Exóticas Invasoras.
En el ámbito autonómico, podemos encontrar algunos ejemplos, como el
Decreto 213/2009 sobre control de especies exóticas invasoras de la Comunidad
Valenciana, el Plan Andaluz para el Control de las Especies Exóticas Invasoras, que
aborda el problema de las invasiones biológicas, y la Estrategia Canaria de la
Biodiversidad.
PROBLEMÁTICA
INVESTIGACIÓN
EN LOS ESPACIOS NATURALES PROTEGIDOS.
EL
PAPEL DE LA
EN LA GESTIÓN DE LAS ESPECIES INVASORAS
Los espacios naturales protegidos generalmente representan las últimas áreas alteradas
(Vitousek et al. 1997), pero no dejan de ser susceptibles a la presencia de especies
invasoras, ya que muchas especies vegetales exóticas han invadido con éxito hábitats
con un alto grado de conservación (Drake 1988; Luken 1988; Broncano et al. 2005).
Este fenómeno, directa o indirectamente relacionado con la actividad humana, se está
generalizando como un problema de manejo prioritario (Luken y Thieret 1997). Cada
vez son más las áreas protegidas que se encuentran amenazadas por las invasiones
biológicas (Luken y Thieret 1997; MacDougall y Turkington 2004; Traveset et al.
2008). Ante este problema, los planes de gestión de las áreas protegidas incluyen o
deben incorporar el control de estas especies (Europarc-España 2002; Balaguer 2004;
Worboys et al. 2005; Andreu et al. 2009; Davis et al. 2011).
La investigación juega un papel de vital importancia en el ámbito de la gestión
mediante el aporte de información sobre las características de las especies invasoras, la
evaluación de sus impactos, la detección, el análisis de riesgos de expansión, las
medidas de control más eficaces, etc. (Pyšek y Richardson 2010). Todo esto mejora el
éxito en la gestión de las especies invasoras y puede implicar una reducción en los
costes económicos.
AILANTHUS ALTISSIMA, ESPECIE OBJETO DE ESTA TESIS DOCTORAL
Ailanthus altissima (Mill) Swingle 1916 (ailanto, árbol del cielo) pertenece a la familia
Simaroubaceae, y es una de las cinco especies del género Ailanthus Desf. Este árbol,
nativo de China y del Norte de Vietnam, fue introducido en Francia en los años 1740
por el misionero Pierre d´Incarville en forma de semillas, y posteriormente (1751), al
ser confundida con la especie Rhus verniciflua, fue transportada a Londres y a otras
9
partes de Europa y América (Hu 1979). Actualmente, esta especie está catalogada como
especie invasora en todos los continentes salvo en la Antártida (Fig. 2; Kowarik y
Säumel 2007). Está considerada como una de las plantas invasoras más extendidas en
Europa y ha sido incluida en la lista Top 20 de especies de prioridad en esta zona
(Sheppard et al. 2006; Pyšek et al. 2009). Asimismo, se encuentra en las Islas
Mediterráneas (Moragues y Rita 2005; Bacchetta et al. 2009; Podda et al. 2010), donde
está considerada como una de las 15 especies de mayor impacto dentro del proyecto
Exotic Plant Invasions: Deleterious Effects on Mediterranean Island Ecosystems
(EPIDEMIE) (Balaguer 2004; Moragues 2005). En España está incluida en la lista
negra preliminar de especies exóticas invasoras (Capdevila et al. 2006), en el Decreto
213/2009 sobre control de especies exóticas invasoras de la Comunidad Valenciana, así
como en el Anexo I del Real Decreto 1628/2011, de 14 de noviembre, por el que se
regula el listado y catálogo español de especies exóticas invasoras, estando prohibida su
introducción en el medio natural, posesión, transporte, tráfico y comercio.
Figura 2 Área de expansión de Ailanthus altissima con una diferenciación del área nativa
en China y Norte de Vietnam (rallado) y su distribución secundaria a lo largo del Mundo
(negro) a partir de su introducción en Europa en los años 1740 (Kowarik y Saümel 2007)
A. altissima es una especie pionera y de crecimiento rápido (1 a 1,5 m/año)
(Zasada y Little 2002). El rango de tamaño que se puede encontrar es de 27−30 m en
zonas templadas y de 18−20 m en zonas meridionales (Hunter 2000; Arnaboldi et al.
2003). Tiene una gran capacidad de rebrotado, formando densos rodales (Kowarik
1995). Es un árbol dioico; algunos autores señalan que las flores pueden ser bisexuales
o los árboles monoicos, aunque esto aún no está demostrado (Kowarik y Säumel 2007).
10
Produce semillas en forma de sámaras y se dispersan de manera anemócora a partir de
septiembre-octubre permaneciendo muchas de ellas en el árbol hasta mayo del año
siguiente (Fig 3; Bory y Clair-Maczulajtys 1980). También se ha encontrado dispersión
hidrocora a través de cursos de agua, la cual juega un importante papel para su
expansión a larga distancia (Kowarik y Säumel 2007, 2008; Kaproth y McGraw 2008;
Säumel y Kowarik 2010). La producción de sámaras es elevada, de hasta 325.000
sámaras por individuo (Little 1974; Bory y Clair-Maczulajtys 1980). Produce una raíz
pivotante y numerosas raíces laterales de una longitud media de 114,4 cm, donde
almacena la mayor parte de las reservas de carbohidratos y proteínas (Dubroca y Bory
1981). La corteza de la raíz, y de otras partes de la especie, así como hojas, y sámaras
contienen componentes herbicidas y alelopáticos que son tóxicos para muchas especies
(Heisey 1990, 1996; Lawrence et al. 1991). En otras palabras, esta especie posee todas
las características necesarias para ser una especie invasora: especie pionera de rápido
crecimiento juvenil, rebrotadora, con sistema radicular potente y producción sustancias
alelopáticas y herbicidas.
Figura 3 Ejemplar ornamental adulto femenino de A. altissima y detalle de panícula de
semillas
Está considerada como una especie oportunista, intolerante a la sombra (Grime
1965), y como una pionera colonizadora de espacios abiertos. En el caso de ecosistemas
forestales, se describe como “gap-dependent”, dependiente de la presencia de huecos en
el dosel del bosque para su desarrollo (Knapp et al. 2000).
11
Fuera de su rango nativo está sujeta a una baja presión herbívora, lo que se ha
atribuido a los componentes tóxicos de sus tejidos (Ohmoto y Koike 1984). En Europa
se ha encontrado dos especies de artrópodos que se alimentan de esta especie,
Hyphantria cunea en Austria y Samia cynthia en Italia y se han identificado 46
artrópodos fitófagos en su rango nativo (Kowarik y Säumel 2007). Asimismo, se han
descrito 65 especies de hongos asociados a la especie (Ding et al. 2006).
La distribución de la especie fuera de su área originaria se debe a su empleo
como especie forestal y ornamental, así como en restauración de taludes de carreteras
(Ruiz de la Torre et al. 1990).
Existe un elevado número de espacios naturales protegidos en el mundo que han
sido invadidos por A. altissima, como por ejemplo los Parques Nacionales de Richmond
y Petersburg en Estados Unidos (Akerson et al. 2001), el Parque Nacional del Danubio
(Austria) (Drescher y Ließ 2006) o el Parque Nacional Aggtelek (Hungria) (Kowarik
2003) en Europa. Concretamente en España, se encuentra en el Parque Nacional de
Sierra Nevada (García et al. 2003) y el Parc Natural de la Sierra de Collserola
(Barcelona) (Meggaro y Vilà 2002), en el Parque Natural de la Sierra de Cardeña y
Montoro, donde su erradicación está planteado grandes dificultades (Algarra et al.
2005), así como en los Parques Naturales del Carrascal de la Font Roja y de Sierra de
Mariola (Alicante y Valencia) (Constán-Nava et al. 2007, 2010).
Para la gestión de especies vegetales invasoras en espacios naturales protegidos
es importante seguir varios pasos (ver arriba apartado Problemática en los espacios
naturales protegidos. El papel de la investigación en la gestión de las especies
invasoras).
Como primer paso, es importante conocer la distribución de la especie invasora
y el grado de invasión que presenta en los diferentes ecosistemas que existen en el área
natural, donde lo prioritario es la conservación de las especies nativas y los ecosistemas
que lo conforman (Vilà et al. 2007; Chytrý et al. 2009). Por un lado, es necesario
conocer la superficie ocupada por la especie invasora. Por otro lado, se ha de determinar
los ecosistemas que han sido invadidos para llevar a cabo actuaciones de control,
priorizando los ecosistemas de mayor interés para la conservación teniendo en cuenta el
paisaje (Reid et al. 2009; Pyšek y Richardson 2010; Vilà y Ibañez 2011). En relación a
A. altissima, estudios previos muestran que ha invadido tanto ecosistemas antrópicos
(áreas periurbanas, cultivos en activo o abandonados, taludes de carretera, Fig. 4), como
ecosistemas naturales, por ejemplo, pinares o bosques de ribera (Kowarik 1983; Danin
12
2000; Constán-Nava et al. 2007; Kowarik y Säumel 2007; pero ver Affre et al. 2010), y
es menor la frecuencia a mayores altitudes (Traveset et al. 2008), pero son escasos los
estudios que tienen el cuenta el paisaje en el proceso de invasión, así como el desarrollo
de la especie en los diferentes hábitats que lo conforman.
Figura 4 Expansión de A. altissima junto a borde de carretera (izquierda) e invasión en
campo de cultivo de olivos (derecha).
En segundo lugar, es necesario el conocimiento de rasgos ecológicos de la
especie invasora que puedan influir en su expansión (Lake y Leishman 2004; Finnoff y
Tschirhart 2005; Andreu y Vilà 2010). Entre estos aspectos, se encuentran la
germinación y el establecimiento. Se conoce que A. altissima germina en diversos tipos
de suelo, y es mayor la germinación con mayor disponibilidad lumínica y a bajas
altitudes (< 1000 m), aunque parece no depender del tipo de hábitat (Mihulka 1998;
Kota et al. 2007; Vilà et al. 2008; Moore y Lacey 2009) o de la fuente de semillas
(Kota et al. 2007; Delgado et al. 2009). Aunque es importante el conocimiento del
efecto de los factores genéticos como ambientales sobre el desarrollo de la especie
invasora, no existen estudios que contemplen los efectos de ambos factores así como
sus interacciones.
Por otro lado, es importante el conocimiento sobre los efectos tanto directos
como indirectos que puede provocar las especies vegetales invasoras en los ecosistemas
invadidos. Estudios previos indican que A. altissima altera el ciclo del nitrógeno
(Castro-Díez et al. 2009), y puede enlentecer la descomposición de hojarasca, aumentar
el N total, C orgánico y pH del suelo, así como la proporción C/N (Vilà et al. 2006;
Godoy et al. 2010; pero ver Castro-Díez et al. 2011). Asimismo, tiene un efecto en el
funcionamiento ecosistémico y en la dinámica y estructura de la vegetación,
13
disminuyendo la biodiversidad (Lawrence et al. 1991; Vilà et al. 2006). A pesar de esto,
aún no se han desarrollado estudios que analicen los efectos directos e indirectos
(mediados por su efecto en la biodiversidad) sobre el múltiples funciones ecosistémicas
así como sobre la diversidad filogenética, en comunidades vegetales invadidas, todo ello
simultáneamente.
Finalmente, dado la problemática que supone la presencia de A. altissima en los
espacios naturales, es imprescindible el desarrollo de actuaciones que contemplen el
control de las especies invasoras y su posible eliminación. En esta línea, se han
desarrollado diversos métodos (manual, mecánico, químico, quema) para combatir a A.
altissima (Hoshovsky 1988; Hunter 2000). El método aplicado más habitual ha sido el
de eliminación mecánica, pero su tendencia a rebrotar hace que sea poco efectiva
(Hoshovsky 1988; Bory et al. 1991; Hunter 2000). En áreas de clima templado se ha
encontrado que el empleo de tratamiento mecánico junto con químico (en concreto, el
glifosato) es el más eficaz (Meloche y Murphy 2006). Sin embargo, bajo clima
mediterráneo no se conoce una técnica a largo plazo que sea efectiva, a pesar de la
necesidad de su control, la cual está incluida en los planes de gestión de numerosos
espacios naturales (Andreu y Vilà 2007; Andreu et al. 2009).
OBJETIVOS Y ESTRUCTURA DE LA TESIS
El objetivo general de esta tesis es analizar diferentes procesos que intervienen en la
invasión de Ailanthus altissima en hábitats mediterráneos para mejorar el éxito en su
gestión en áreas protegidas.
Los objetivos específicos que se plantean en esta tesis son los siguientes:
Analizar la distribución actual y el grado de invasión de la especie invasora A.
altissima en la LIC Serra de Mariola i Carrascal de la Font Roja, así como determinar
las variables ambientales que afectan a su desarrollo (Capitulo 1)
Identificar la importancia relativa tanto de los factores genéticos como
ambientales (variabilidad genética, condiciones climáticas y tipo de hábitat) y sus
interacciones sobre la emergencia y establecimiento temprano de la especie invasora A.
altissima bajo condiciones mediterráneas (Capitulo 2)
14
Analizar los efectos tanto directos como indirectos (mediados o no por su efecto
sobre biodiversidad) de A. altissima sobre la multifuncionalidad ecosistémica en
hábitats riparios bajo clima Mediterráneo (Capitulo 3)
Determinar la mejor estrategia para controlar la invasión de A. altissima a largo
plazo en bosques mediterráneos (Capitulo 4)
Figura 5 Diagrama sintético en el que se incluyen los capítulos que conforma la tesis
doctoral. Modelo adaptado de Foxcroft et al. 2011
Los cuatro capítulos que conforman esta tesis están escritos en inglés y se
presentan en formato de artículo para su publicación en revistas científicas de carácter
internacional. Esto implica algunas redundancias en relación a la descripción de la
especie y de algunas de las áreas de estudio que son comunes entre algunos de los
capítulos.
15
METODOLOGÍA Y ÁREA DE ESTUDIO
ÁREA DE ESTUDIO
El área general de estudio es el Lugar de Interés Comunitario “Serra de Mariola
i Carrascal de la Font Roja” (LIC, de aquí en adelante; Directive 92/43/EEC). Está
situado al Sureste de España, entre el Norte de la provincia de Alicante y Sur de la
provincia de Valencia, cuenta con una superficie de 19945.9 ha e incluye los Parques
Naturales del Carrascal de la Font Roja y de la Sierra de Mariola (Fig. 6). Para el
capítulo 1 se ha incluido el área general así como las carreteras que rodean el área
natural, ya que se consideran los hábitats más invadidos por A. altissima, con una
superficie total de muestreo de 22757 ha. Los capítulos 2 y 4 fueron desarrollados
dentro del P.N. del Carrascal de la Font Roja. Para el capítulo 3, se seleccionaron los
bosques de ribera presentes en la LIC.
Figura 6 Mapa de localización del área general de estudio de la tesis doctoral
Por un lado, el P. N. de la Sierra de Mariola está situado entre los municipios de
Alcoy, Cocentaina, Muro de Alcoy, Agres, Alfafara, Onteniente, Bocairente, Bañeres de
Mariola e Ibi y cuenta con una superficie de 16.926 ha, incluyendo el área de
amortiguamiento (PORN, Fig. 6.; Decreto 76/2001). Por otro lado, el P. N. del
16
Carrascal de la Font Roja (situado entre los términos de Alcoy e Ibi, Alicante) se
extiende en dirección este-oeste, e incluye la Sierra del Menejador, así como los valles y
llanuras que hay a su alrededor. Comprende una superficie de 2298 ha, que junto a su
área de influencia (Área PORN) ocupan una superficie de 6326 ha (Decreto 121/2004;
Fig. 6).
Geológicamente, el área general de estudio se engloba en el prebético y están
formados mayoritariamente por rocas carbonatadas del Eoceno y Mioceno, y también
con materiales del Triásico, facies de Keuper, de composición básicamente arcillosa
(IGME 2010).
El clima es Mediterráneo con influencia continental, con precipitación media
anual y temperatura de 647 mm y 14,7 ºC, respectivamente (estación meteorológica de
Bocairente, localizada en el área de estudio, a 641 m s.n.m., datos del periodo 19852006 para la temperatura y 1996-2006 para la precipitación; Rívas Martínez et al. 2007).
El área de estudio incluye diversos hábitats. Por un lado se encuentran los
bosques caducifolios (Acer granatense Boiss., Fraxinus ornus L., Quercus faginea
Lam., Sorbus aria L., Polygonatum odoratum Mill.) y los encinares (Quercus ilex
subsp. ballota (Desf.) Samp.). Asimismo aparecen encinares abiertos (Quercus
coccifera L., Juniperus phoenicea L., Rhamnus lycioides L.), pinares (Pinus halepensis
Mill.), con presencia de muchas especies aromáticas en el caso de la Sierra de Mariola
(Serra 2007), y matorrales (Genista scorpius (L.) DC., Juniperus sp., Quercus coccifera
L.). También aparecen bosques de ribera (Salix L. sp, Umus minor Mill., Populus alba
L., Populus nigra L.), con especies de interés (Adiantum capillus-veneris L.,
Trachelium caeruleum L.), y que se encuentran incluidos como uno de los hábitats
prioritarios (Directive 92/43/EEC; Serra 2007). En la Sierra de Mariola se encuentra un
elevado número de endemismos iberolevantinos y setabenses (Serra 2007; Serra y Soler
2011).
METODOLOGÍA
A continuación se presenta un breve resumen de cada capitulo y la metodología
utilizada en cada uno de ellos.
17
ANÁLISIS DE LA DISTRIBUCIÓN Y DESARROLLO DE POBLACIONES DE AILANTHUS ALTISSIMA
EN ÁREAS PROTEGIDAS MEDITERRÁNEAS (CAPITULO 1)
Este capítulo comprende el análisis de la distribución y grado de invasión de A.
altissima en un área protegida a nivel nacional e internacional (LIC Serres de Mariola i
el Carrascar de la Font Roja) y la influencia de variables ambientales en el desarrollo de
las poblaciones de la especie. La hipótesis principal es que A. altissima invade
numerosos hábitats, principalmente corredores de transporte, donde su desarrollo es
mayor, y su crecimiento está influido por variables ambientales, como la altitud. Se
localizaron todas las poblaciones presentes en el área natural, incluyendo hasta las
carreteras que bordean el mismo (en total 22757 ha de área de estudio). Asimismo, se
estimó el área de cada población y se indicó el tipo de uso del suelo en el que se
desarrollaban. Se analizó el área invadida en los diferentes usos del suelo, así como
respecto a las figuras de protección que abarca el área muestreada (parques naturales,
LIC y área adyacente) mediante sistemas de información geográfica. Por otro lado, se
seleccionaron aleatoriamente 99 poblaciones de A. altissima, se estimaron variables de
crecimiento (área poblacional y densidad) y variables ambientales (altitud, pendiente,
orientación, hábitat, presencia/ausencia de carreteras).
LA
VARIABILIDAD GENÉTICA MODULA EL EFECTO DEL TIPO DE HÁBITAT Y DE LAS
CONDICIONES AMBIENTALES EN EL ESTABLECIMIENTO TEMPRANO DE
A.
ALTISSIMA
(CAPITULO 2)
Este capítulo se centra en el estudio de los factores ambientales y genéticos que limitan
la expansión de A. altissima en las fases iniciales del proceso de invasión. La hipótesis
principal de este capítulo es que los factores ambientales, no genéticos, afectan a la
germinación de A. altissima. En este capítulo se analizó la germinación y el
establecimiento temprano de semillas procedentes de diferentes árboles de A. altissima
en experimentos de laboratorio y de campo. En cámara de germinación se analizó el
efecto de la variabilidad genética, de la temperatura y sus interacciones en la
germinación y viabilidad de las semillas, así como del tiempo de almacenamiento de las
mismas. En campo, se consideraron seis hábitats mediterráneos (pinar solana, pinar
umbría, encinar abierto, talud de carretera, borde de camino y cultivo abandonado) y se
analizaron la germinación y establecimiento temprano de A. altissima, examinando las
diferencias entre fuentes de semillas, hábitats, variables ambientales (relacionados o no
con el tipo de hábitat) y condiciones climáticas durante el periodo de estudio.
18
EFECTOS DIRECTOS E INDIRECTOS DE LA INVASIÓN DE A. ALTISSIMA SOBRE COMUNIDADES
RIPARIAS Y MULTIFUNCIONALIDAD ECOSISTÉMICA (CAPITULO 3)
Este capítulo incluye el análisis de los efectos directos e indirectos (mediados por su
efecto sobre biodiversidad) de A. altissima sobre el funcionamiento ecosistémico en
comunidades vegetales de bosques de ribera mediterráneos. Las hipótesis principales
son 1) A. altissima reduce además de la riqueza de especies, la diversidad filogenética,
2) los efectos de A. altissima sobre la diversidad vegetal reducen indirectamente la
multifuncionalidad ecosistémica y 3) la reducción de la multifuncionalidad por A.
altissima no es solo indirecta, mediada por su efecto sobre la diversidad vegetal, sino
que también directamente por sus efectos conocidos sobre ciclos de nutrientes y
propiedades del suelo. Para ello, se midieron variables de la vegetación (riqueza de
especies y cobertura, diversidad filogenética), funciones del suelo (biomasa vegetal y
actividades enzimáticas) y propiedades del suelo (pH, conductividad eléctrica, materia
orgánica, P disponible) en diez parcelas de 100 m2 invadidas por A. altissima y en diez
parcelas control (no invadidas).
CONTROL
DE
A.
ALTISSIMA A LARGO PLAZO EN ESPACIOS NATURALES MEDITERRÁNEOS
(CAPITULO 4)
Este capítulo se centra en determinar la mejor estrategia de control de la especie A.
altissima en espacios naturales protegidos mediterráneos. Las hipótesis principales son
1) el tratamiento de desbroce anual incrementará el crecimiento de rebrotes y
supervivencia durante los siguientes años, implicando cambios ecofisiológicos que le
permitirán crecer y mantenerse, 2) los tratamientos de dos desbroces anuales y de
desbroce junto a aplicación de herbicida serán los métodos más eficientes, porque
afectarán negativamente a la especie invasora, reduciendo el desarrollo de la especie
como resultado de una reducción de las reservas radiculares así como por sus efectos en
la morfología causados por el herbicida, lo que reducirá la capacidad competitiva de la
especie frente a las especies nativas. Se seleccionaron diferentes poblaciones de A.
altissima y en cada una de ellas se aplicó un tratamiento: control, un desbroce anual, dos
desbroces anuales y desbroce junto a aplicación de herbicida. Cada tratamiento fue
aplicado sobre cada población a lo largo de 4 años consecutivos. Diferentes variables
fueron medidas a lo largo de un periodo de cinco años de muestreo: biomasa, altura
densidad de rebrotes, así como potencial hídrico, conductancia estomática e índice de
área foliar (LAI).
19
RESULTADOS GENERALES
ANÁLISIS DE LA DISTRIBUCIÓN Y DESARROLLO DE POBLACIONES DE AILANTHUS ALTISSIMA
EN ÁREAS PROTEGIDAS MEDITERRÁNEAS (CAPITULO 1)
La presencia de A. altissima en el área de estudio estuvo principalmente asociada a
carreteras, construcciones antrópicas y caminos (Fig. 7). A partir de estas zonas se
encontró invadiendo numerosos tipos de uso del suelo, desde cultivos (en activo y
abandonados) hasta ecosistemas semi-naturales y naturales, como matorrales, pinares,
encinares y bosques de ribera. La pendiente, la orientación y el hábitat influyeron sobre
el desarrollo de la especie.
Figura 7 Localización de A. altissima (círculos rojos) en el área de estudio (línea gris:
carreteras; línea verde: vía de tren abandonada). Arriba izquierda: borde de camino; arriba
derecha: vía de tren abandonada; abajo izquierda: borde de carretera; abajo derecha:
construcción antrópica
20
Tabla 1 Presencia de las poblaciones de A. altissima en los diferentes tipos de uso del suelo del
área de estudio
Área ocupada Área ocupada Área total
Uso del suelo
N
%
min (m2)
máx.(m2)
ocupada (m2)
Construcción antrópica
46
16,4 0,8
795,8
2765,4
Desde construcción antrópica
Invadiendo cultivos agrícolas
11
3,9 0,2
534,6
731,5
Invadiendo espacios verdes
2
0,7 6,5
79,7
86,2
Invadiendo bordes de camino
36
514,7
1927,5
12,9 0,7
Invadiendo encinares
2
0,7 72,1
506,8
578,9
Invadiendo pinares
6
2,1 1,1
226,2
474,5
Invadiendo matorrales
6
2,1 32,9
198,9
810,0
Bordes de carretera
Desde bordes de carretera
Invadiendo cultivos agrícolas
Invadiendo cultivos abandonados
Invadiendo encinares
Invadiendo pinares
Invadiendo bosques de ribera
Invadiendo matorrales
48
17,1 0,2
235,4
1431,0
5
3
1
28
10
2
1,8
1,1
0,4
10,0
3,6
0,7
197,9
87,0
45,0
1505,6
4887,4
7436,9
454,2
107,6
45,0
8124,2
10325,1
13102,4
Bordes de camino
Desde bordes de camino
Invadiendo pinares
Invadiendo bosques de ribera
Invadiendo matorrales
32
11,4 0,2
883,5
2876,2
4
1
3
1,4
0,4
1,1
50,0
20,1
54,1
1906,8
20,1
135,3
2777,2
20,1
260,6
Vía de tren abandonada
16
5,7
0,1
1841,0
4321,6
Bosques
Encinares
Pinares
Bosques de ribera
1
1
1
0,4
0,4
0,4
4,2
50,7
325,8
4,2
50,7
325,8
4,2
50,7
325,8
Canteras abandonadas
Cultivos agrícolas
Cultivos abandonados
Total
2
12
1
280
0,7
4,3
0,4
100
626,0
0,2
557,0
737,5
28,7
557,0
1363,5
107,8
557,0
53628,4
LA
3,1
3,1
45,0
0,8
4,0
5665,5
VARIABILIDAD GENÉTICA MODULA EL EFECTO DEL TIPO DE HÁBITAT Y DE LAS
CONDICIONES AMBIENTALES EN EL ESTABLECIMIENTO TEMPRANO DE
A.
ALTISSIMA
(CAPITULO 2)
La germinación y la supervivencia temprana de A. altissima estaban principalmente
afectadas por factores externos, tales como las condiciones climáticas (pulsos de lluvia),
el tipo del hábitat, el sitio y el suelo desnudo (Fig. 8). Sin embargo, la influencia de
estos factores varió con la procedencia de las semillas (Fig. 9), sugiriendo que la
21
preferencia de hábitat y el óptimo ambiental varían dependiendo de la fuente de
semillas.
180
A
T min
T max
T media
140
30
120
20
100
80
10
60
40
0
Temperatura media (ºC)
Precipitación (mm)
160
40
20
0
16
B
Borde de carretera
Borde de camino
Pinar de umbría
Pinar de solana
Encinar abierto
Cultivos abandonados
Germinación (%)
14
12
10
= 0.07; P = 0.791
T: F
1,270
P: F
= 3.8; P = 0.0541
1, 270
8
Ti: F
= 14.9; P = 0.0001
12, 270
6
H: F
= 1.2; P = 0.3707
S: F
= 2.1; P = 0.0159
5, 270
12, 270
4
Ti × H: F
= 1.6; P = 0.011
Ti × S: F
= 1.4; P = 0.0162
70, 270
168, 270
2
0
J
F
M
A
M
J
J
A
S
O
N
D
J
F
M
A
M
J
2010
2009
Mes y año
Figura 8 Precipitación mensual (barras grises) y temperatura (puntos) (A) y curvas de
germinación (media ± ES, n = 6) de A. altissima en seis tipos de hábitat incluidos (B) durante el
periodo del estudio (T: temperatura, P: precipitación mes anterior, Ti: tiempo, H: hábitat, S:
sitio)
22
30
Borde de camino
Borde de carretera
1
2´
3
4
6
7´
8
9
10´
11
12
20
10
Germinación (%)
0
30
Pinar de umbría
H: F
20
10
Pinar de solana
= 1.1; P = 0.3609
5,198
F: F
= 4.2; P = 0.0001
10,198
S: F12, 198= 8.2; P = 0.0001
H × F: F
= 1.4; P = 0.0781
50,198
F × S: F
= 0.6; P = 0.9984
120,198
0
30
Cultivos abandonados
Encinar abierto
20
10
0
J F M A M J J AS O N D J F M A M J
2009
J F M A M J J AS O N D J F M A M J
2010
2009
2010
Mes y año
Figura 9 Curva de germinación (media ± ES) de semillas de A. altissima de 11 árboles en cada
tipo de hábitat a lo largo del periodo de estudio (H: hábitat, F: fuente de semillas, S: sitio)
EFECTOS DIRECTOS E INDIRECTOS DE LA INVASIÓN DE A. ALTISSIMA SOBRE COMUNIDADES
RIPARIAS Y MULTIFUNCIONALIDAD ECOSISTÉMICA (CAPITULO 3)
La riqueza de especies, la diversidad filogenética y la multifuncionalidad ecosistémica
fueron reducidas en presencia de A. altissima (Tabla 2). El efecto de la especie invasora
sobre la multifuncionalidad fue indirecto y estuvo mediado principalmente por su efecto
sobre la diversidad filogenética y en menor medida sobre la riqueza de especies. La
cobertura y composición de especies se vieron afectadas en las parcelas invadidas, pero
no las propiedades del suelo.
23
Tabla 2 Resumen de los efectos de la invasión de A. altissima en los atributos del ecosistema
medidos. Se incluyen el estadístico F y el P valor de los efectos analizados y la varianza (R2)
explicada por el modelo. Aunque algunas de las variables utilizadas son derivadas de múltiples
datos (p. ej. Propiedades del suelo y multifuncionalidad, ver capitulo 3), se ha añadido una
interpretación de los resultados. Las etiquetas “efectos directos o indirectos” indican si se
usaron los datos brutos o los residuales de regresiones lineales, respectivamente. En el último
caso, la variable predictiva utilizada en la regresión lineal (con la multifuncionalidad como
variable de respuesta) se muestra entre paréntesis. Atributos del suelo = primer eje del Análisis
de Componentes Principales desarrollado con pH, conductancia eléctrica, materia orgánica, P
disponible y cobertura de hojarasca. Diversidad filogenética = resultado del índices de similitud
de especies filogenéticas. Multifuncionalidad = índice M construido mediante el promedio de
Z-scores de tres variables funcionales utilizadas (actividades enzimáticas glucosidasa y
fosfatasa, biomasa vegetal)
F1,18
R2
P-valor
Interpretación
Atributos suelo
0,85
0,05
0,369
No efectos en pH, EC, OM o P
disponible del suelo
Riqueza especies
12,7
0,41
0,002
Diversidad
filogenética (PSE)
29,4
0,62
<0,0001
11,3
0,39
0,004
3,23
0,15
0,089
1,36
0,07
0,26
EFECTOS DIRECTOS
Atributos Ecosistema
EFECTOS
INDIRECTOS
Multifuncionalidad
Multifuncionalidad
(riqueza especies)
CONTROL
Multifuncionalidad
(PSE)
DE
A.
A. altissima reduce la riqueza
y diversidad filogenética
A. altissima reduce la
multifuncionalidad del
ecosistema. Sin embargo, este
efecto está mediado
principalmente por su efecto
en la diversidad filogenética y
en menor medida, sobre la
riqueza de especies
ALTISSIMA A LARGO PLAZO EN ESPACIOS NATURALES MEDITERRÁNEOS
(CAPITULO 4)
El tratamiento más efectivo a largo plazo para controlar y reducir a la especie invasora
A. altissima fue el de corte y aplicación del herbicida glifosato, por su reducción en
biomasa, e índice de área foliar de los rebrotes (Tabla 3). Los tratamientos que solo
incluyen métodos de corte no redujeron a la especie.
24
Tabla 3 Biomasa, diámetro basal (DB), altura y densidad de A. altissima (mean ± ES, n = 3)
para cada tratamiento y año de estudio (ANOVA y test de Tukey, P < 0.05). Leyenda: Control;
1CT: tratamiento de un corte anual; 2CT: tratamiento de dos cortes anuales; CHT: tratamiento
corte y aplicación de herbicida. Diferentes letras indican diferencias significativas entre
tratamientos
Tratamiento
2005
Biomasa (gr m-2 )
Control
1318 ±603 a
1CT
1879 ±1205 a
2CT
2262 ±1825 a
CHT
828 ±246 a
DB (mm)
Control
11.7 ±2.2 a
1CT
14.8 ±4.0 a
2CT
14.5 ±3.6 a
CHT
12.8 ±1.1 a
Altura (m)
Control
1.2 ±0.3 a
1CT
1.3 ±0.3 a
2CT
1.4 ±0.4 a
CHT
1.0 ±0.2 a
Densidad (num m-2)
Control
8 ±2 a
1CT
4 ±0 a
2CT
8 ±3 a
CHT
5 ±1 a
2006
2007
2008
2009
2422 ±1042 a
283 ±85 ab
341 ±144 ab
36 ±25 b
3330 ±1511 a
802 ±463 ab
414 ±163 ab
115 ±42 b
3129 ±1366 a
235 ±81 ab
534 ±55 a
45 ±30 b
3560 ±1656 a
160 ±69 a
206 ±101 a
15 ±15 b
14.3 ±3.2 a
7.4 ±1.3 ab
7.7 ±0.5 ab
4.4 ±0.9 b
15.5 ±3.4 a
7.6 ±1.8 ab
7.1 ±0.2 ab
4.5 ±0.3 b
15.7 ±3.1 a
5.7 ±1.6 bc
6.9 ±0.7 ab
2.4 ±0.7 c
17.3 ±4.3 a
6.2 ±2.1 a
6.0 ±0.3 a
1.4 ±0.5 b
1.3 ±0.4 a
0.6 ±0.1 ab
0.6 ±0.2 ab
0.2 ±0.0 b
1.5 ±0.3 a
0.7 ±0.2 ab
0.6 ±0.0 b
0.3 ±0.0 b
1.4 ±0.3 a
0.5 ±0.1 b
0.5 ±0.0 b
0.1 ±0.0 b
1.6 ±0.5 a
0.5 ±0.1 b
0.4 ±0.0 b
0.07 ±0.0 b
8 ±2 a
8 ±1 a
10 ±3 a
3 ±1 a
6 ±1 a
11 ±2 a
13 ±5 a
6 ±3 a
7 ±1 a
14 ±4 a
16 ±6 a
5 ±2 a
7 ±1 a
12 ±6 a
11 ±4 a
2 ±1 a
25
Figura 10 Diagrama de los resultados más significativos de la tesis doctoral (modelo adaptado de Foxcroft et al. 2011)
DISCUSIÓN GENERAL
Esta tesis doctoral abarca diferentes aspectos que intervienen en el proceso de invasión
de la especie vegetal exótica Ailanthus altissima bajo clima Mediterráneo. Se han
considerado sus características ecológicas, la susceptibilidad de los hábitats y el
contexto del sistema, con la finalidad de incrementar nuestro conocimiento sobre la
especie y sus efectos sobre ecosistemas nativos y sobre la mejora en la gestión de esta
especie en espacios naturales protegidos. La invasión de A. altissima en ecosistemas
mediterráneos está influida tanto por características genéticas como ambientales las
cuales afectan al desarrollo de la especie. A partir de su uso antrópico ha invadido
numerosos ecosistemas, entre ellos los bosques de ribera, donde afecta negativamente
sobre la riqueza de especies, la diversidad filogenética y múltiples funciones
ecosistémicas. El método más efectivo para su control a largo plazo es el de desbroce y
aplicación de glifosato.
DISTRIBUCIÓN DE A. ALTISSIMA, LA IMPORTANCIA DE LA DISPERSIÓN SECUNDARIA
Al igual que en estudios previos que fundamentan el carácter pionero de esta especie
(Kowarik 1983; Hulme 2004; Kowarik y Säumel 2007), se ha observado que la
aparición de A. altissima está principalmente asociada a perturbaciones de origen
antrópico, principalmente las asociadas a las vías de transporte y las construcciones
urbanas, periurbanas y rurales (Capitulo 1). El uso de la especie invasora en áreas
mediterráneas como especie ornamental y la recomendación de su uso en restauración
del paisaje hasta la actualidad (Ruiz de la Torre et al. 1990; Valladares et al. 2011), ha
provocado que se extienda en estos lugares usando las vías de transporte y los cursos de
agua como corredores, siendo el tráfico rodado y los cursos de agua mecanismos de
dispersión secundaria a largas distancias bien conocidos para ésta y otras especies
invasoras (p.ej. Timmins y Williams 1991; Tyser y Worley 1992; Kota 2005; Kowarik y
von der Lippe 2006; 2011). La dispersión de las semillas, fase importante para el
movimiento de las especies (Harper 1977) puede ser primaria (alrededor de la planta
madre) y secundaria (mediada por agentes tras la dispersión primaria). La existencia de
ambos tipos de dispersión contribuyen a la supervivencia de las especies vegetales (p.ej.
Forget 1990; Moore 1997; Ruiz et al. 2010), especialmente en el caso de las especies
invasoras, las cuales han visto aumentada su área de colonización debido en muchos
27
casos, a la intervención humana en la dispersión (Hodkinson y Thompson 1997;
Kowarik y von der Lippe 2006; 2011; von der Lippe y Kowarik 2007). Concretamente,
A. altissima, gracias a la dispersión secundaria, y su uso como especie ornamental, ha
invadido hábitats naturales o semi-naturales rodeados por áreas antropizadas, como son
comunidades ruderales, matorrales, pinares, cultivos abandonados, e incluso hábitats de
interés para la conservación, como encinares y bosques de ribera (Kowarik 1983; Lepart
y Debussche 1991). La presencia de espacios abiertos (gaps) en estos ecosistemas
podría haber facilitado la entrada de la especie invasora (Davies 1944; Kowarik 1995;
Knapp y Canham 2000; Kota 2005; Capítulo 2). La configuración del paisaje
circundante (composición y estructura espacial) es también muy importante en relación
a la presencia y establecimiento de especies invasoras (Pauchard y Alaback 2004; Vilà e
Ibáñez 2011). Esta configuración de paisaje puede explicar la invasión de A. altissima
en diferentes hábitats, por ejemplo, muchas de las poblaciones expandidas desde los
bordes de carreteras o caminos a pinares cercanos, o aquellas poblaciones
extendiéndose desde áreas construidas, principalmente rurales o periurbanas, a campos
de cultivos circundantes.
RASGOS
ECOLÓGICOS DE LA ESPECIE VS CARACTERÍSTICAS DEL HÁBITAT Y
CONDICIONES AMBIENTALES EN EL PROCESO DE INVASIÓN
A pesar de que la presencia de las vías de comunicación (carreteras, pistas forestales)
explican la distribución y colonización de A. altissima, estos factores no explican el
desarrollo de tales poblaciones. El éxito de la invasión de especies exóticas puede estar
influido por la disponibilidad de hábitats, de ahí la variación en el grado de invasión en
los diferentes hábitats (Hansen y Clevenger 2005), y por el hábitat per se, ya que la
dispersión de semillas está influida por el grado de perturbación de los mismos
(Landerberg et al. 2007). En este sentido, en esta tesis doctoral se han observado
diversas variables ambientales predictivas del desarrollo de las poblaciones de A.
altissima. En primer lugar, el área poblacional ha sido mayor en umbría (Capitulo 1).
Estudios previos son contradictorios. A pesar de que A. altissima ha sido considerada
como especie intolerante a la sombra (Miller 1990; Facelli y Pickett 1991; Knapp y
Canham 2000), Espenschied-Reilly y Runkle (2008) no encontraron un efecto
significativo de la orientación en la presencia de A. altissima, y Kota et al. (2007)
encontraron que A. altissima crecía más en umbría a partir del segundo año de
crecimiento. El área poblacional ha sido también mayor en pendientes elevadas, lo que
28
puede ser debido a las condiciones del micrositio en áreas con gran pendiente, que
favorecen la expansión de la especie invasora (Le Maitre et al. 1996; Kohama et al.
2006; Kowarik y Säumel 2007). La altitud no ha influido significativamente en el
tamaño de la población, al contrario que lo encontrado en otros estudios (Traveset et al.
2008), excepto por su limite altitudinal, menor a 1050 m (Kowarik y Säumel 2007). En
relación a la densidad de pies, se han encontrado mayores densidades en
bosques/matorrales frente a los otros hábitats considerados. Esto podría deberse a un
efecto secundario de las medidas de gestión forestal desarrolladas previamente al
análisis. En estos hábitats, las medidas de control consistentes únicamente en el
desbrozado de la especie han sido aplicadas con mayor frecuencia respecto a los otros
hábitats (por ejemplo, en bosques de ribera aún no se han aplicado medidas de control)
en base a los planes de gestión del área de estudio (Decreto 76/2001; Decreto
121/2004), provocando un incremento en los rebrotes y por tanto, de la densidad de pies
(Hoshovsky 1988; Bory et al. 1991; Constán-Nava et al. 2010). Asimismo, Traveset et
al. (2008) también encontraron diferencias entre hábitats, por lo que el tipo de hábitat
puede influir de forma directa en el desarrollo de la especie.
Otra condición ambiental importante que influye sobre A. altissima es la
temperatura, la cual ha tenido un efecto significativo sobre la germinación, con mayores
tasas de germinación a temperaturas más bajas (Capitulo 2). Estos resultados contrastan
con estudios previos (Little 1974, Graves 1990) y con nuestros resultados en campo
(Capitulo 2): los pulsos emergentes en el campo están relacionados con pulsos de lluvia,
pero no con la temperatura. La explicación más posible para estos resultados
contradictorios es que la temperatura en el campo fluctúa (tanto diaria como
estacionalmente) y no es fija; la germinación, por lo tanto, no puede ser relacionada con
un único valor de temperatura. Por otro lado, y en contraste con estudios previos usando
la misma especie (Vilà et al. 2008) ha habido diferencias significativas en las curvas de
germinación entre los diferentes tipos de hábitats considerados, los cuales están
relacionados con las diferencias encontradas en relación a variables ambientales. El tipo
de hábitat con menores tasas de germinación y supervivencia temprana es el talud de
carretera, lo que puede estar relacionado con la baja disponibilidad de agua y la baja
fertilidad en los suelos normalmente encontrados en estas áreas (Bochet and GarciaFayos 2004; García-Palacios et al. 2010). Los taludes de carretera son ecosistemas
comúnmente invadidos por A. altissima bajo condiciones Mediterráneas (Kowarik
1983; Danin 2000; Constán-Nava et al. 2007; Traveset et al. 2008). El efecto del tipo de
29
hábitat en la tasa de germinación ha variado según las estaciones o las características
microambientales de cada sitio dentro del hábitat. El experimento de campo ha revelado
que el 61% de la varianza en la tasa de germinación de las semillas estaba explicado por
el porcentaje de suelo desnudo en la parcela. Esto demuestra la importancia de las áreas
con alto porcentaje de suelo desnudo para la germinación de A. altissima,
particularmente en hábitats con relativamente poco estrés (por ejemplo encinares
abiertos), como se ha indicado previamente para esta y otras especies invasoras (p.ej.
Burke y Grime 1996; Bartuszevige et al. 2007; Kota et al. 2007). Esta dependencia con
el suelo desnudo puede ser explicado por la baja competencia con otras plantas y una
baja presencia de hojarasca en estas áreas, lo que puede tener efectos tanto efectos
directos negativos (una menor disponibilidad de recursos e hidratación física de la
emergencia de semilla) y efectos indirectos por el incremento de herbivoría por insectos
(Facelli y Pickett 1991; Facelli 1994). El efecto del sitio en la germinación de A.
altissima puede estar relacionado con factores microambientales no incluidos en el
estudio de campo, tales como la humedad del suelo o la disponibilidad de luz (esta
última fue medida indirectamente mediante la cobertura vegetal, pero no fue
estadísticamente significativa), lo que ha sido encontrado relevante en el éxito de A.
altissima (Kota et al. 2007; González-Muñoz et al. 2011). El hecho de que todas las
plántulas germinadas en campo de A. altissima muriesen, ya fuese en invierno o verano,
sugiere que las condiciones climáticas podrían ser el principal factor limitante para el
éxito invasor de la especie. Sin embargo, esta conclusión debe ser considerada con
cuidado, ya que los arboles de los que procedían las semillas utilizadas en los
experimentos de laboratorio y campo se encontraban en un único tipo de ambiente.
Además de los factores ambientales, los rasgos ecológicos son importantes para
el desarrollo de las especies. En el caso de las especies invasoras, los rasgos
relacionados con el desarrollo, como las tasas de crecimiento, son mayores que en
especies nativas (e.g. Godoy et al. 2009; van Kleunen et al. 2010; Godoy et al. 2011), y
no se han encontrado diferencias en plasticidad fenotípica (por ejemplo, como respuesta
a gradientes de nutrientes y luz). En A. altissima, los rasgos ecológicos influyen en fases
tempranas de la invasión, presentando una elevada plasticidad a la sequia mediterránea
por medio de la estrategia de ahorro del agua, estrategia que permite sobrevivir bajo
condiciones de déficit hídrico al igual que muchas especies nativas (Levitt 1980;
Vilagrosa et al. 2003; Vilagrosa et al. 2005). Esta estrategia le permite a la especie
invasora crecer y desarrollarse durante periodos de sequía (Levitt 1980). El alto
30
potencial hídrico al alba encontrado en las plantas de A. altissima, principalmente
durante la sequía estival, puede indicar una mejor rehidratación frente a especies nativas
(Capitulo 4). Por ejemplo, Fraxinus ornus, un árbol nativo caducifolio, muestra
potenciales más negativos (-2,56 ± 0,56 MPa) frente a A. altissima (-0,6 ± 0,04 MPa)
bajo las mismas condiciones ambientales (Constán-Nava et al. 2009). Esta elevada
rehidratación de A. altissima puede permitirle un mejor crecimiento y mejorar sus
capacidades competitivas frente a las especies nativas bajo situaciones de sequía.
Esta plasticidad a la sequia mediterránea demostrada por A. altissima se ve
modificada al aplicar métodos de control sobre ella (Capitulo 4). En los rebrotes de los
tres tratamientos aplicados (un desbroce, dos desbroces y desbroce junto a aplicación de
herbicida) se detectaron cambios ecofisiologicos sustanciales. Los rebrotes tratados con
un desbroce y dos desbroces mostraron altas tasas de conductancia estomática durante
primavera, especialmente al mediodía, lo que podría ayudar a la especie invasora a
recuperarse y crecer tras los tratamientos, lo que explicaría la falta de reducción de
biomasa. Por otro lado, la alta conductancia estomática registrada en las hojas de los
rebrotes de desbroce y aplicación de herbicida, principalmente durante la sequia estival,
puede ser debida a los cambios producidos por el herbicida sobre la morfología foliar
(S. Constán-Nava, observ. pers. Fig. 11), lo que podría perjudicar al estado hídrico de la
especie y reducir su desarrollo, corroborando así, el resto de los resultados en este
tratamiento.
Figura 11 Ejemplar de A. altissima con efectos foliares como resultado de la aplicación de
desbroce y herbicida en el tocón del que ha surgido
Asimismo, en esta tesis se muestra que existe una variabilidad genética que
influye en la capacidad germinativa y en el establecimiento temprano, la cual modula su
capacidad de colonizar diferentes hábitats (Capitulo 2). Aunque se han encontrado
31
variaciones en la germinación de semillas entre fuentes maternales en otras especies
(p.ej. Baskin y Baskin 1998), en el caso de A. altissima no se ha encontrado ningún
efecto de este factor sobre la germinación (solo en el peso de las semillas; Kota et al.
2007; Delgado et al. 2009). Sin embargo, en esta tesis se demuestra que, tanto en
condiciones de campo como de laboratorio, los factores genéticos afectan a la capacidad
germinativa y establecimiento temprano de A. altissima. La tasa de supervivencia y de
crecimiento de A. altissima difieren entre procedencias (Feret 1985), lo que sugiere la
existencia de algún componente genético que afecta al desarrollo de la especie. Estos
efectos genéticos vendrían determinados por el árbol maternal y su interacción con el
ambiente, así como aquellos procedentes de las contribuciones parentales (Roach y
Wulff 1987; Baskin y Baskin 1998; Bischoff et al. 2006). En este sentido, las
condiciones ambientales que experimenta el árbol maternal durante la fructificación,
como las diferencias en disponibilidad hídrica, nutrientes o luz, podrían explicar
cambios en las tasas de germinación de las semillas procedentes del mismo árbol, pero
recogidas en diferentes años. Esta variabilidad entre años distintos dentro del mismo
árbol también puede estar relacionada con las contribuciones genéticas paternales, que
pueden diferir de un año a otro. Independientemente de la variabilidad genética, se
encontró un porcentaje de germinación en campo muy bajo (siempre menor de 35 %),
algo que también se ha observado en otras zonas (Kota et al. 2007; Vilà et al. 2008). La
baja viabilidad y germinación obtenidas para esta especie contrastan con el alto grado
de invasión de A. altissima en muchos hábitats en todo el mundo (Kowarik y Säumel
2007). Posibles explicaciones para estos resultados contradictorios son 1) la alta
fecundidad de la especie, con un único árbol capaz de producir un elevado número de
semillas (Little 1974; Bory y Clair-Maczulajtys 1980), 2) el uso de esta especie en
restauración de carreteras y el papel del tráfico rodado como dispersor secundario, 3) la
capacidad rebrotadora de la especie (Kowarik y Säumel 2007), y 4) el sistema radicular
potente que posee, con raíces laterales de casi 30 m de longitud (Kiermeier 1987).
EFECTOS DE LA ESPECIE INVASORA SOBRE LOS ECOSISTEMAS NATIVOS
Numerosos estudios han evaluado los efectos de las especies vegetales invasoras a
diferentes niveles.
En relación a las especies, las especies vegetales invasoras han causado, en
general, efectos negativos sobre la biodiversidad vegetal y animal (Richardson et al.
1989; Maerz et al. 2005; Hejda et al. 2009; Vilà et al. 2011; Watling et al. 2011), con
32
algunas excepciones (ver Sax y Gaines 2003; Meffin et al. 2010). En el caso de A.
altissima, se ha visto un efecto negativo sobre la biodiversidad (Vilà et al. 2006; Motard
et al. 2011), causando una alteración y reducción de la composición y cobertura de las
especies, de la riqueza especifica y de la diversidad filogenética (Capitulo 3). En
relación a la composición de especies, se ha encontrado otras especies invasoras como
Robinia pseudoacacia L. en las áreas invadidas por A. altissima. Se sabe que la
presencia de especies invasoras acelera la invasión de otras especies exóticas y
amplifica sus efectos en las comunidades nativas (invasional meltdown; ver Simberloff
y Von Holle 1999; Richardson et al. 2000; Pyšek y Richardson 2010). Además, entre las
especies presentes en las áreas no invadidas, se ha encontrado especies raras de alto
interés para la conservación, como Cephalanthera damasonium (Mill.) Druce.
Asimismo, la reducción de la diversidad filogenética bajo la presencia de A. altissima
podría causar efectos negativos sobre los servicios y múltiples funciones ecosistémicos,
ya que las comunidades formadas por especies funcionalmente más diversas (es decir,
más filodiversas) son más probables que mantengan mayores niveles de funciones del
ecosistema (Zavaleta et al. 2010).
Por otro lado, las especies vegetales invasoras han causado efectos sobre las
propiedades del suelo, ciclos de nutrientes o comunidades nativas microbianas así como
sobre los procesos ecosistémicos asociados (p.ej. Vitousek and Walker 1989; Ehrenfeld
2003; van der Putten et al. 2007; Weidenhamer and Callaway 2010). Estudios previos
sobre A. altissima indican efectos directos de la especie sobre varias funciones del
ecosistema y atributos del suelo (descomposición de la hojarasca y pH, Godoy et al.
2010; ciclo del N, Castro-Díez et al. 2011); evidencias también encontradas en estudios
observacionales (pH, C orgánico, proporción C/N, Vilà et al. 2006; Gómez-Aparicio y
Canham 2008).
El estudio de los efectos de las especies vegetales invasoras se ha centrado
básicamente sobre la biodiversidad o sobre el funcionamiento de los ecosistemas por
separado, pero no conjuntamente. En esta tesis se incluye uno de los pocos estudios que
analizan el efecto de una especie invasora sobre diversas medidas de biodiversidad y la
riqueza de especies, y sobre múltiples funciones del ecosistema, todo ello
simultáneamente (Capitulo 3). Este es también el primer estudio que separa efectos
directos e indirectos (mediados por su efecto sobre la biodiversidad) de plantas
invasoras sobre el funcionamiento ecosistémico. Así, aunque A. altissima tiene un
rápido crecimiento (Zasada y Little 2002), y puede incrementar la fertilidad del suelo y
33
la fijación de C (Vilà et al. 2006; Gómez-Aparicio y Canham 2008), los resultados
encontrados en esta tesis muestran un efecto negativo neto de A. altissima en la
productividad de herbáceas y arbustivas y sobre los ciclos de C y P. Estos resultados
instan a futuras investigaciones que incluyan estudios manipulativos y la medida de
múltiples funciones del ecosistema simultáneamente para concluir si la especie invasora
altera directa o indirectamente la funcionalidad del ecosistema.
La reducción de la multifuncionalidad en presencia de A. altissima encontrada
en esta tesis está indirectamente mediada por la reducción en la diversidad filogenética
y en menor medida, en la riqueza de especies. Entre estas funciones, y en línea con
investigaciones previas (Strauss et al. 2006; Diez et al. 2008; Zavaleta et al. 2010) es
posible que las comunidades más filodiversas (por lo tanto con mayores niveles de
funciones en el ecosistema) puedan prevenir futuras invasiones.
IMPLICACIONES PARA LA GESTIÓN
Además de conocer la distribución, características que influyen el establecimiento y los
efectos sobre los ecosistemas nativos, es necesario conocer cómo combatir a las
especies invasoras. Existen numerosos métodos empleados sobre las especies vegetales
invasoras, desde métodos mecánicos, biológicos, hasta químicos. Concretamente en A.
altissima, el método usado más común ha sido el de desbroce de la parte aérea de la
especie (Hoshovsky 1988; Hunter 2000). En áreas templadas se han aplicado diferentes
herbicidas siendo el glifosato el más eficaz (Meloche y Murphy 2006), al igual que en
bajo clima mediterráneo a largo plazo (Capitulo 4). En esta tesis se han probado 4
metodologías diferentes: control (ninguna actuación), un desbroce anual, dos desbroces
anuales, y desbroce y aplicación de herbicida (glifosato). Ni el tratamiento de un
desbroce ni el de dos desbroces han reducido significativamente la densidad de rebrotes,
al igual que en otros estudios desarrollados en áreas templadas (Bory et al. 1991; Burch
y Zedaker 2003; Meloche y Murphy 2006). El tratamiento combinado de desbroce y
aplicación de glifosato sobre el tocón recién cortado es el más eficiente, principalmente
debido a sus efectos negativos sobre el crecimiento de la parte aérea de los rebrotes de
la especie (reducción de la biomasa, índice de área foliar, además de efectos sobre
rasgos ecofisiológicos). Este tratamiento finalmente ha eliminado los rebrotes de más de
la mitad de las parcelas tras cinco años de aplicación repetida, lo que sugiere que, si se
prolonga por más tiempo, una aplicación más persistente del desbrozado junto a la
aplicación de herbicida puede resultar en un control total. Además, este tratamiento
34
parece reducir su capacidad competitiva frente a las especies nativas como sugiere el
hecho de que se ha observado recolonización natural tras cinco años de aplicación del
tratamiento. Por ejemplo, existe una recuperación natural de especies nativas como
Thymus vulgaris L. subsp. vulgaris, Brachipodium retusum (Pers.) P. Beauv., Cistus
albidus L., o plántulas de Quercus ilex L., Viburnum tinus L.and Pinus halepensis Mill
en parcelas con este tratamiento, pero no en el resto de las técnicas empleadas (S.
Constán-Nava, observ. pers.). Esto puede indicar una reducción en la competición por
luz y en la producción de componentes alelopáticos por una menor biomasa e índice de
área foliar.
Es importante reducir el impacto y crecimiento de A. altissima en áreas
antropogénicas, y evitar la construcción de corredores de transporte junto a hábitats que
son especialmente sensibles a la invasión. Asimismo, los modelos de distribución
potencial de A. altissima (Albrigth et al. 2010) en espacios naturales protegidos pueden
ser una herramienta muy útil para identificar las áreas potencialmente más propensas a
la invasión, es decir, áreas disponibles para la especie invasora, de forma que se puedan
determinar áreas de actuación prioritarias, por lo que deberían incluirse en los planes de
gestión. En las áreas afectadas, medidas de restauración que incluyan la eliminación de
A. altissima podrían favorecer la colonización de especies nativas y ayudar a preservar
la riqueza (especialmente en el caso de especies amenazadas). Por tanto, en la
aplicación de tratamientos de control y erradicación, habría que priorizar, aquellos
hábitats de interés de conservación, por ejemplo encinares y bosques de ribera.
Como se ha demostrado, A. altissima provoca efectos negativos en los
ecosistemas. Mediante la eliminación de la especie invasora y la reintroducción de
especies nativas, junto con otras actuaciones complementarias, se podría restablecer los
servicios y funciones ecosistémicos importantes. En este aspecto, la reducción en la
diversidad filogenética observada en parcelas invadidas por A. altissima, junto con su
efecto indirecto sobre el funcionamiento ecosistémico, sugiere que el control directo
sobre esta especie, sin otras medidas necesarias, podría ayudar a restaurar múltiples
funciones y servicios ecosistémicos. Por otro lado, se sabe que la entrada de especies
invasoras puede estar influida por la estructura filogenética de la comunidad receptora
(Vacher et al. 2010), con especies exóticas más alejadas a la comunidad receptora con
más probabilidades de acabar como invasoras en estas comunidades (Strauss et al. 2006;
Diez et al. 2008). Aunque no ha sido estudiado antes, el conjunto de los resultados de
ambas líneas de investigación sugiere que conservar conjuntos de especies
35
filogenéticamente más diversas podría incrementar la resistencia de las comunidades a
futuras invasiones a la vez que acelera el restablecimiento de las funciones
ecosistémicas perdidas. Comunidades formadas por especies funcionalmente más
diversas (es decir, con mayor diversidad filogenética) son más probables que mantengan
mayores niveles de funciones del ecosistema tales como la resistencia a la invasión de
plantas (Zavaleta et al. 2010), y que incluyan entre esas especies algunos taxones
cercanamente relacionados a posibles invasoras, por lo tanto les previene del
establecimiento e invasión en la comunidad (Strauss et al. 2006; Diez et al. 2008).
Como resultado de esta tesis doctoral se ha generado el manual para la gestión de
Ailanthus altissima en espacios naturales protegidos de ámbito mediterráneo.
PERSPECTIVAS DE FUTURO
A partir de esta tesis doctoral, se han abierto diferentes líneas de interés para la
investigación.
Por un lado, se encuentra el desarrollo de modelos de distribución potencial de la
especie invasora a partir de los datos obtenidos en campo, así como la comparación de
la misma con respecto a modelos de distribución potencial de especies nativas,
constituyendo una potente herramienta para la toma de decisiones.
Por otro lado, es interesante conocer la diversidad genética y genotípica de A.
altissima en el Mediterráneo (Dallas et al. 2005) para determinar los efectos genéticos
en el desarrollo de la especie.
Son necesarios más estudios relacionados con los efectos del tipo hábitat en la
viabilidad de las semillas de A. altissima. En esta tesis, las semillas analizadas
procedían de un tipo de hábitat, por lo que serían complementarios posteriores estudios
que incluyan este factor.
El uso de técnicas como las empleadas para determinar la multifuncionalidad
ha mostrado ser muy útil para determinar conjuntamente efectos directos e indirectos de
la especie invasora en ecosistemas invadidos como son los bosques de ribera. La
aplicación de esta metodología en otros ecosistemas así como sobre otras especies
puede ser de gran interés, así como el estudio a nivel filogenético.
En relación al estudio sobre los tratamientos de control sobre A. altissima, sería
interesante el seguimiento de las colonizaciones de especies nativas a largo plazo para
determinar si existe una recuperación natural total de las áreas afectadas, o si
finalmente, son necesarias medidas de restauración complementarias.
36
CONCLUSIONES
1.
Las poblaciones de A. altissima están asociadas principalmente a perturbaciones
de origen antrópico, como carreteras, construcciones y caminos. Los resultados
sugieren que la especie invasora, a partir de estos usos del suelo, coloniza hábitats
adyacentes como pinares, bosques de ribera o ecosistemas que si no fuera por la
presencia de este tipo de perturbaciones no llegaría, como los encinares
2.
Diferentes variables ambientales influyeron sobre el desarrollo de las poblaciones
de la especie invasora, como la orientación, la pendiente y el hábitat
3.
Es importante conocer la localización y el grado de invasión en áreas protegidas
así como tener el cuenta la configuración del paisaje en el desarrollo de estrategias
de control en áreas naturales
4.
El componente genético no solo afecta al desarrollo de A. altissima, sino que
también modula su respuesta a factores ambientales, tales como la lluvia, o el
suelo desnudo, los cuales parecen ser los principales conductores de la
germinación y el establecimiento temprano de la especie invasora
5.
Tanto la preferencia del tipo de hábitat como el óptimo ambiental varía según la
fuente de semillas
6.
Existe una reducción de la riqueza, de la diversidad filogenética y de la
multifuncionalidad ecosistémica en presencia de A. altissima en bosques de ribera
mediterráneos
7.
Es importante determinar los efectos directos e indirectos de las especies
invasoras para mejorar las estrategias de gestión en espacios naturales. Esta tesis
incluye una metodología fácil para analizar tales efectos directos e indirectos
mediante datos observacionales
37
8.
El método de control más efectivo para eliminar a A. altissima en ecosistemas
mediterráneos a largo plazo es el de desbroce y aplicación de herbicida
(glifosato), una metodología que debe ser incluida en los planes de manejo de las
áreas protegidas
9.
Aquellos métodos basados únicamente en la eliminación mecánica (aún repitiendo
el tratamiento dos veces al año) no reducen a la especie. La gestión pasiva basada
en la ausencia de control de A. altissima debe tener en cuenta todos los efectos
ecológicos negativos que causa
10.
El desarrollo de estudios de investigación basados en la ecología de especies
invasoras, como A. altissima, ayudan a mejorar la gestión desarrollada en áreas
protegidas, pudiendo disminuir desde efectos ecológicos a económicos
38
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CAPITULO 1
Distribution and performance of populations
of the invasive species Ailanthus altissima on
Mediterranean Protected Areas1
1
Manuscrito enviado
Autores: Soraya Constán-Nava, Andreu Bonet
49
50
ABSTRACT
If we are to improve the control and eradication of invasive species within protected
natural areas, we need to know their distribution and degree of invasion in relation to
environmental conditions. A. altissima is a tree from China and North Vietnam that has
invaded numerous ecosystems around the world. Our work aimed to analyze the
occurrence of A. altissima in diverse land use types in a Site of European Community
Importance with a Mediterranean climate, as well as determining the influence of
several environmental variables on its performance. To do this, we located and
measured the area of all populations in a whole natural protected area (22757 ha,
including surrounding lands with roads). The distribution of such populations was
related to land use type. To study the performance of these A. altissima populations
according to environmental conditions, we selected 99 populations and measured their
area and stem density. We also measured the aspect, slope, altitude, habitat and
presence/absence of roads as environmental and land-use variables of interest. Our
results show that the distribution of A. altissima in the studied area was highly
associated with roadsides, built-up areas and pathsides. Our study suggests that this
species uses these anthropogenic land use types to further its colonization to such
surrounding habitats as pine forests, riparian forests and scrublands. Our results agree
with previous literature and indicate that this species uses roadsides as ecological
corridors to reach habitats that it could not otherwise invade, such as oak forests. The
mean population area was larger on pronounced slopes and moister north-slope aspects.
Stem density, however, was mainly affected by habitat, with denser populations in
scrublands/forests than in the other habitats studied (agricultural fields, urban areas,
riparian forests). Our study highlights the importance of considering landscape when
analyzing the invasion of a natural area by an exotic species. By quantifying the degree
of invasion according to different land use types, it may be possible to develop adequate
management strategies, prioritizing habitats with more conservational importance or
those more prone to invasion. In the case of A. altissima under Mediterranean
conditions, we suggest focusing first on controlling their population of anthropogenic
areas to prevent further expansion.
51
52
INTRODUCTION
I
nvasive species are causing numerous ecological and economic impacts in most
ecosystems worldwide (Vitousek et al. 1997; Mack et al. 2000; Simberloff 2001).
Protected areas, including those with Mediterranean environments, are not free from
this global problem and have been successfully colonized by many invasive species
(Drake 1988; Luken 1988; Luken and Thieret 1997; Broncano et al. 2005). In these
areas, the control of invasive species has been prioritized in management plans in order
to conserve native species and ecosystems (Luken and Thieret 1997; Pauchard et al.
2003; Andreu et al. 2009).
Early detection of alien species is necessary when prevention is not enough to
avoid their invasion of protected areas, (Pauchard et al. 2003; Paurchard and Alabarck
2004; Rout et al. 2011). When analyzing the distribution and performance of alien plant
populations according to environmental factors and the sensitivity of the different
habitats to an invasion, it is important to determine the risk and current impact of said
invasion and to develop focused control actions (Daehler et al. 2004; Hulme 2006;
Hortal et al. 2010). Moreover, analysis of alien plant distribution should include not
only environmental information, but also landscape composition and structure, which
may thus reveal the effect of significant differences in land management (Pauchard and
Alaback 2004; Vilà and Ibáñez 2011). Undoubtedly, increasing our knowledge about
how invasive species colonize and grow in different environments will enhance the
efficiency of management strategies in protected areas by prioritizing the most
vulnerable areas to invasion, or by identifying the ecological corridors that accelerate
such invasion (Hansen and Clevenger 2005; Hulme 2006; Gulezian et al. 2010; Säumel
and Kowarik 2010). As an example, linear habitats such as roads and rivers are known
to act as ecological corridors enhancing the dispersal of invasive species into adjacent
ecosystems (Pyšek and Prach 1994; Forman and Alexander 1998; Parendes and Jones
2000; Säumel and Kowarik 2010). Furthermore, a high degree of disturbance, such as
that found in anthropogenic systems, is known to increase the sensitivity of a given area
to an invasion (Vilà and Ibáñez 2011).
Ailanthus altissima (Mill.) Swingle is a tree originates from China and North
Vietnam and was introduced into Europe in the 18th century. It has successfully invaded
several habitats with a Mediterranean climate including roadsides, old fields and pine,
oak and riparian forests (Kowarik 1983; Constán-Nava et al. 2007; Kowarik and Säumel
53
2007). A. altissima is a hardy invasive species due to two main ecological
characteristics: 1) its ability to reproduce both sexually and through resprouting (Little
1974; Bory and Clair-Maczulaijtys 1980; Kowarik 1995; Kowarik and Säumel 2007),
and 2) its herbicidal and allelopathic compounds, which make it highly competitive
compared to native plants (Heisey 1990, 1996; Heisey and Heisey 2003). Previous
studies indicate that A. altissima grows in different soil types and at different altitudes
(with a altitude limit at about 1000 m.a.s.l.), grows better in well-lit environments, and
varies its performance among habitats, with larger populations on roadsides and old
fields than in temporary streams (Kota el al. 2007; Traveset et al. 2008; Moore and
Lacey 2009).
This study was designed to evaluate the distribution and performance of
Ailanthus altissima, regarding land use types and the environmental (both ecological
and anthropogenic) conditions found throughout an entire Mediterranean protected area.
To do so, we located all the populations of A. altissima in the study area, measured their
performance (area and stem-density) and established the environmental characteristics
for each population. The main hypotheses were that A. altissima invades several
habitats, principally linear transport corridors, where its performance is greater and its
growth is influenced by environmental variables such as altitude (Kowarik and Säumel
2007; Traveset et al. 2008).
MATERIAL AND METHODS
STUDY SITE
The study was conducted at the Site of European Community Importance (hereafter
SCI; Directive 92/43/EEC) Serres de Mariola i el Carrascar de la Font Roja in
southeastern Spain. This conservation area stretches over 19,945.9 ha and consists of
two Natural Parks with their corresponding buffer zones for protection: Sierra de
Mariola (16,926 ha; 38º 44´ 1´´ N, 0º 35´ 30´´ O) and Carrascal de la Font Roja (6,301
ha; 38º 38´ 51´´ N, 0º 32´ 46´´ O) situated between north-western Alicante and southwestern Valencia, Spain. We included the roads surrounding this protected area in our
study because they are considered to be the habitats most prone to invasion by A.
altissima (Kowarik and Säumel 2007), giving a final studied area of 22,757 ha. The
climate is Mediterranean, with a mean annual precipitation of 408-493 mm and a mean
annual temperature ranging from 13-14.5 ºC (Baradello and Bañeres de Mariola
54
meteorological stations located in the study area, at 788 and 729 m.a.s.l.; data from
period 2005-2010). The soils are xerorthents on limestone (Soil Survey Staff 2006),
with the presence of impermeable clays. The study area includes diverse habitats such
as deciduous forests (Acer granatense Boiss., Fraxinus ornus L., Quercus faginea
Lam.), oak forests (Quercus ilex subsp. ballota (Desf.) Samp.), Aleppo pine forests
(Pinus halepensis Mill.), riparian forests (Salix sp, Ulmus minor Mill.) and scrublands
(Genista scorpius (L.) DC., Juniperus sp. pl, Quercus coccifera L.). Invasion of the
studied area by A. altissima started at least 75 years ago (Natural Park staff, personal
comm.), because of its use as an ornamental tree on private property and for roadside
and railway restoration.
DISTRIBUTION ANALYSIS
All the populations of A. altissima present in the study area were located and marked
using a global positioning system (GARMIN GPS 12; Fig. 1). In the field, we assigned
a land use type (roadside, pathside, built-up area, agricultural fields, oak forests, etc.)
and used the GPS to measure the area of each population. These areas were later
introduced as polygons into a Geographic Information System (ArcGis 9.2; ESRI 2008)
to create a new layer (hereafter A. altissima layer). To determine the degree of invasion
according to conservation designations, the layer of A. altissima was crossed with
Conservation designation layers (Source: Conselleria de Infraestructuras, Territorio y
Medio Ambiente). Layer crossing was conducted using the ArcGis 9.2 join function,
which permits analysis of the total invaded area in the different categories within the
other layers included (Peña 2006). The invaded area in both Natural parks was
compared using the χ2 test.
PERFORMANCE ANALYSIS
We randomly selected a set of 99 populations from those described above and measured
their area and stem density. To measure stem density, we applied two methodologies
because of logistic constraints: we counted and measured the root collar diameter
(henceforth RCD) of all the stems in 54 randomly selected populations and for the rest
of the populations we randomly located three 2 × 2 m plots and counted and measured
the RCD of all shoots within them. Both methodologies were used for 12 of these
populations and the results gathered were highly correlated (r = 0.73; P < 0.001).
Moreover, we used these 12 populations to establish a correction factor for the number
55
of stems obtained for the large populations and thus homogenized the data obtained
with both methodologies. We do not therefore expect any influence on our results or
conclusions as a result of the differential methodology used depending on population
area. Stem density (stems/m2) and the distribution of size classes were estimated using
this sampling procedure. The following environmental variables were recorded for each
population: habitat (grouped in agricultural fields, forests/scrublands, riparian forests,
urban areas; based on Traveset et al. 2008), altitude, slope (low, medium, high), aspect
(South, North), and presence/absence of roads.
Separate multiple regressions for each performance variable used (population
area and stem-density) were conducted to determine which environmental variables
(habitat, altitude, slope, aspect and presence/absence of roads) were most important
with regard to the performance of A. altissima. Population area and density were log10
transformed to meet normality and homoscedasticity assumptions. Frequency analysis
was carried out on RCD. Multiple regression analyses were conducted using SPSS v.15
(SPSS Inc., Chicago, IL, USA).
RESULTS
DISTRIBUTION ANALYSIS
A total of 280 populations of A. altissima were located in the study area, with a total
invaded area of 5.36 ha, making up 0.02 % of the study area (Table 1). Differences in
the invaded area were found between the Sierra de Mariola and Carrascal de la Font
Roja Natural Parks (Table 1).
Table 1 Invaded area (ha) and % invaded by A. altissima according to the conservation
designation of the study area
Conservation designation
Area (ha)
Invaded area (ha) % invaded
Total study area
SCI
22,757
19,946
5.36
4.89
0.020
0.020
Sierra de Mariola N.P.
16,926
0.74*
0.004
Carrascal de la Font Roja N.P.
6,301
5.36*
0.080
2
*Significant differences in invaded area between both Natural Parks, χ > 7.88, P < 0.005
The occurrence of A. altissima was principally associated with roadsides, built-up areas
and pathsides (Fig. 1, Table 2). From these areas, the invasive species encroached into
the surrounding natural and anthropogenic habitats. For example, A. altissima
56
frequently invaded pathsides from built-up areas (ca. 3,000 m2). From roadsides, the
invasive species encroached principally upon pine forests (more than 8,000 m2) and
riparian forests (more than 10,000 m2), and, despite low-frequency invasion of
scrubland, this is where the greatest invaded area was seen (over 13,000 m2). From
pathsides, it invaded approximately 3,000 m2 of pine forests. A. altissima was also
frequent on railway lines, with 4,321.6 m2 of invaded area. Isolated populations were
found in old fields and in oak, pine and riparian forests.
Table 2 Presence of A. altissima populations in different land use types in the study
area. N: number of populations; %: percentage of the total invaded area; Min. occupied
area: minimum occupied area; max occupied area: maximum occupied area
16.4
Min occupied Max occupied
area (m2)
area (m2)
0.80
795.8
Total occupied
area (m2)
2,765
11
2
36
2
6
6
3.90
0.70
12.9
0.70
2.10
2.10
0.20
6.50
0.70
72.1
1.10
32.9
534.6
79.70
514.7
506.8
226.2
198.9
731.5
86.20
1,927
578.9
474.5
810.0
48
17.1
0.20
235.4
1,431
5
3
1
28
10
2
1.80
1.10
0.40
10.0
3.60
0.70
3.10
3.10
45.0
0.80
4.00
5,665
197.9
87.00
45.00
1,506
4,887
74,37
454.2
107.6
45.00
8,124
10,325
13,102
Land use type
N
%
Built-up area
From built-up area
Encroaching into agricultural fields
Encroaching into green spaces
Encroaching into pathsides
Encroaching into oak forests
Encroaching into pine forests
Encroaching into scrublands
46
Roadsides
From roadsides
Encroaching into agricultural fields
Encroaching into old fields
Encroaching into oak forests
Encroaching into pine forests
Encroaching into riparian forests
Encroaching into scrublands
Pathsides
From pathsides
Encroaching into pine forests
Encroaching into riparian forests
Encroaching into scrublands
32
11.4
0.20
883.5
2,876
4
1
3
1.40
0.40
1.10
50.0
20.1
54.1
1,907
20.10
135.3
2,777
20.10
260.6
Railway line
16
5.70
0.10
1,841
4,322
Forests
Oak forests
Pine forests
Riparian forests
1
1
1
0.40
0.40
0.40
4.20
50.7
325.8
4.20
50.70
325.8
4.20
50.70
325.8
Abandoned quarries
Agricultural fields
Old fields
Total
2
12
1
280
0.70
4.30
0.40
100
626
0.20
557
737.5
28.70
557.0
1,363
107.8
557.0
53,628
57
Figure 1 Distribution of A. altissima (red circles) in the study area (grey lines: roads; green line: abandoned railway line). Above left: pathsides,
above right: railway line, below left: roadsides, below right: built-up area
58
PERFORMANCE ANALYSIS
Multiple regression analyses showed that aspect and slope affected the area of A.
altissima populations (Table 3) whilst the rest of the variables entered were not
significant (P > 0.05). Slope aspect was the most important predictor, with population
areas significantly greater on North than on South-sloping aspects. Slope was a second
order predictor, with larger populations on pronounced (high) slopes than on the rest. In
contrast to population area, habitat was the only significant predictor variable for stemdensity (Table 3), with denser populations in scrubland/forests than in the other habitats
included.
Table 3 Results of multiple regressions analysing the effect of environmental variables
on A. altissima performance
Analysis of variance
2
R
2
Area (m )
df
0.43 97
Density (number of shoots/m2)
0.21 88
Predictor variable
F-value
P
11.11
<0.0001
4.25
0.04
t
P
Aspect
4.11
<0.0001
Slope
2.18
0.03
Habitat
-2.06
0.04
RCD presented an asymmetric distribution, ranging from 0.003 cm to 58.93 cm, with
the major frequency being lower than 10 cm (Fig. 2).
12000
Min 0.003 cm
Max 58.93 cm
Mean 1.001 cm
10000
Total shoots
8000
6000
4000
2000
0
0
10
20
30
40
50
60
RCD (cm)
Figure 2 Frequency distribution of the basal diameter (RCD) of A. altissima shoots in
the study area (n = 11950)
DISCUSSION
Similarly to previous studies (Kowarik 1983; Hulme 2004; Kowarik and Säumel 2007),
the occurrence of Ailanthus altissima was principally associated with anthropogenic
disturbances, mainly linear transport corridors and built-up areas. The much higher
abundance of populations of A. altissima in these areas can be explained by the common
use of this invasive species for ornamental purposes and the fact that it was
recommended for use in landscape restoration until recently (Ruiz de la Torre et al.
1990; Valladares et al. 2011). From this initial colonization, many of the populations
could have expanded due to road traffic acts as a secondary dispersal mechanism for A.
altissima, thus transporting the seeds over long distances (Kowarik and von der Lippe
2006; Kowarik and von der Lippe 2011). This secondary dispersal explains the presence
of isolated populations along road or pathsides, where the initial populations could be
dispersed by wind or vehicles (Kota 2005; Kowarik and von der Lippe 2006; Kowarik
and von der Lippe 2011).
As in other parts of the Mediterranean Basin (Kowarik 1983), A. altissima
colonized oak forests and scrubland close to roads (Table 2). Paths or roads are
disturbances which cross administrative boundaries in different ways, for example,
between private or public properties, or between city areas and permit the entrance of
invasive species which could not access the areas in any other way (Timmins and
Williams 1991; Tyser and Worley 1992). Furthermore, transport corridors such as roads
are often placed close to administrative boundaries and are altered and artificial areas
(Forman and Moore 1992; Landres et al. 1998), which are known to affect alien species
spreading into natural areas, increasing their colonization in adjacent habitats (Tyser
and Worley 1992; Spellerberg 1998; Parendes and Jones 2000; Pauchard and Alaback
2004), as well as being regarded as a severe threat to the native species in many nature
reserves (Usher 1988; Spellerberg 1998).
A. altissima was present in built-up areas, both abandoned and non-abandoned,
where it was introduced as an ornamental tree. The invasive species appeared forming
dense stands in the abandoned areas, where population size could have increased over
time since abandonment (Domènech et al. 2005; Vilà and Ibáñez 2011). On the other
hand, A. altissima has also invaded non-abandoned built-up areas, which could have
been caused by the high propagule pressure from ornamental trees (to 325,000 samaras
per tree, Little 1974; Bory and Clair-Maczulajtys 1980) and also because ineffective
60
control measures applied to new shoots (Such as mechanical removal, S. Constán-Nava,
pers. observ.) may have increased its density (Hoshovsky 1988; Bory et al. 1991;
Constán-Nava et al. 2010).
The colonization of surrounding habitats from anthropogenic areas could be
caused by the possible presence of gaps, such as oak forests which could have
facilitated the entrance of the species (Davies 1944; Kowarik 1995; Knapp and Canham
2000; Kota 2005). Land use type has an influence on invasion, with anthropogenic areas
being the most susceptible. Nevertheless, the surrounding landscape configuration is
also very important with regard to the presence and establishment of alien species
(Pauchard and Alaback 2004; Vilà and Ibáñez 2011). This landscape configuration
could explain the A. altissima invasion of the different habitats, for example, the many
populations that have expanded from road or pathsides into nearby pine forests, or those
populations extending from built-up areas into surrounding agricultural fields.
The degree of invasion of A. altissima from road and pathsides to surrounding
habitats varied (Table 2). This may due to the suitability of the habitats, which could
have an influence on the success of the invasion (Hansen and Clevenger 2005), and also
due to the habitat, where seed dispersion differs between habitats according to the
degree of disturbance (Landerberg et al. 2007). A. altissima invaded riparian forests, as
has been found in other studies (Kowarik 1983; Lepart and Debussche 1991). Riparian
forests were mainly found on roadsides, and water presence combined with seed
dispersion by water, wind and traffic could aid its colonization (Kowarik and von der
Lippe 2006; Kota 2005; Kaproth and McGraw 2008; Kowarik and von der Lippe 2011).
Presence/absence of roads was not a variable predictor of A. altissima
performance (considered in terms of population area and stem-density), despite the fact
that the invasive species is very frequently found at roadsides. Our results show that,
despite roads explaining the distribution and colonization of A. altissima, these areas do
not explain the growth of such populations. In this sense, environmental variables that
were good predictors of the performance of the invasive species were found (Traveset et
al. 2008). First, the population area was greater at shady sites. Previous studies are
contrasted. Despite A. altissima has been reported as a shade-intolerant species (Miller
1990; Facelli and Pickett 1991; Knapp and Canham 2000), Espenschied-Reilly and
Runkle (2008) did not find any significant effect of this aspect on the presence of A.
altissima, and Kota et al. (2007) found that A. altissima grew more in shady sites as
from the second growth. The population area was also greater on steep slopes. This
61
could be due to light resources and microsite conditions on areas with steep slopes
favouring the expansion of the invasive species (Le Maitre et al. 1996; Kohama et al.
2006; Kowarik and Säumel 2007). Our results indicated that altitude had no influence
on population size, in contrast to previous studies (Traveset et al. 2008), except for its
absence at > 1,200 m (Kowarik and Säumel 2007). This could be due to the control
measures of the invasive species counteracting the effect of altitude.
Secondly, habitat was the predictor variable for stem-density, with the density
being greater in forests/scrublands than in the other habitats. In these habitats, control
measures based on mechanical methods have been more frequently applied to A.
altissima than in the other habitats (for example, in riparian forests, control measures
have not yet been applied) based on management plans (Decreto 76/2001; Decreto
121/2004). Consequently, density increased (Hoshovsky 1988; Bory et al. 1991;
Constán-Nava et al. 2010).
A wide range of various sizes of A. altissima was found in the study, from
seedlings and small resprouts to adult trees, with a much higher frequency of young
shoots. Despite A. altissima being recommended for the landscape restoration of road
slopes, the annual obligatory roadside cleaning (Real Decreto-ley 11/2005) using
mechanical or chemical methods that do not involve the correct herbicide could increase
invasive species stem density and reduce size class (Hoshovsky 1988; Bory et al. 1991;
Meloche and Murphy 2006; Constán-Nava et al. 2010).
A. altissima invaded a larger area in Carrascal de la Font Roja than in Sierra de
Mariola. These differences could be explained by the high density of roads and paths
and tourist vehicles in the former, combined with greater use of the invasive species in
built-up areas. Our results highlight the importance of understanding the degree of
invasion and spatial distribution for improving control on different levels of
conservation. Furthermore, these results highlight the importance of reducing the impact
and growth of A. altissima in anthropogenic areas, and of avoiding the construction of
transport corridors close to habitats that are especially sensitive to invasion. Therefore,
obligatory control actions should be applied, prioritizing habitats of conservation
interest (oak and riparian forests; Decreto 76/2001; Decreto 121/2004). Potential models
of the current distribution of A. altissima (Albrigth et al. 2010) in natural protected areas
should be included in management plans because they could help to identify areas
potentially prone to invasion.
62
CONCLUSIONS
A. altissima is related with human disturbances by means invades surrounding habitats,
including conservation interest areas such as oak forests, and threatening native species
in the natural area. Moreover, the performance of the species was found to be affected
by environmental variables (aspect, slope and habitat).
Our results highlight the importance of identifying the location and degree of invasion
in protected areas and of taking into account the composition and configuration of the
landscape when developing conservation strategies for protected areas.
ACKNOWLEDGEMENTS
S. Soliveres provided helpful comments on an early version of this manuscript. We also
thank Language Centre (University of Alicante) for improving the English of this
manuscript. We are grateful for the collaboration of the staff at the Natural Parks and all
private landowners and for the permits granted. C. Constán, A. Constán, M.J. Nava, E.
Pastor, N. Vizcaíno, S. Soliveres, J. Monerris, M. Herrera, J. Acosta and the other
collaborators helped with the fieldwork. E. Pastor, Q. Rubio and S.M. Catalán helped
with the Geographic Information System. This research and SCN PhD fellowship were
supported by the GV06/02 projects founded by the Valencian Regional Government,
ESTRES (063/SGTB/2007/7.1) and RECUVES (077/RN08/04.1) founded by the
Spanish Ministry for the Environment. Font Roja Natura UA Scientific Station (ECFRN
UA), which depends on the Pro-Vice-Chancellorship for Research, Development and
Innovation (VIDI) at the University of Alicante, also supported this research.
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APÉNDICE FOTOGRÁFICO
Foto 1 Invasión de A. altissima en una
masía abandonada dentro de la LIC
Foto 2 Población de A. altissima
invadiendo matorral a partir de uso
ornamental
Foto 3 Individuos de A. altissima
creciendo dentro de carrascal
Foto 4 Individuos de A. altissima
invadiendo pinar
67
68
CAPITULO 2
Genetic variability modulates the effect of
habitat type and environmental conditions on
early invasion success of Ailanthus altissima in
Mediterranean ecosystems2
2
Este capitulo está en 2ª revisión en Biological Invasions
Autores: Soraya Constán-Nava, Andreu Bonet
69
70
ABSTRACT
At the early stages of an invasion by an exotic species, there are diverse environmental
and genetic factors that limit its expansion. Ailanthus altissima is a tree from China and
northern Vietnam which has become an invasive species in numerous ecosystems
around the word. Our objective was to identify the relative effect of both genetic and
environmental factors and how they interact with the emergence and early establishment
of this invasive tree under Mediterranean conditions. To achieve this, seed germination
and early establishment from different maternal sources were analyzed under contrasted
environmental conditions in a series of experiments, using both laboratory and field
approaches. Seed germination and early survival were affected by environmental factors
such as habitat type, the percentage of bare soil and climatic conditions (rainfall pulses),
although the influence of these factors changed depending on the maternal source. Our
study reveals that the genetic component affected not only the performance of A.
altissima, it also modulated its response to environmental factors, which seemed to be
the main drivers of germination and early establishment for this species. Our results
highlight the importance of considering both genetic and environmental factors when
studying plant invasion risk and success, and may be helpful in predicting and reducing
the spread of this species in Mediterranean ecosystems.
71
72
INTRODUCTION
T
he introduction of invasive plant species and their colonization depend not only on
the ecological traits of the invader species but also on the conditions of both
habitat and climate, resulting in varying distributions of invasive species in different
types of habitat (Gleadow and Rowan 1982; East et al. 1999; Koop 2004; Lloret et al.
2005). Previous studies show that disturbed habitats such as human-created
environments are vulnerable to plant invasion because of the lack of competition or gap
abundance (Elton 1958; Baker 1986; Hobbs and Huenneke 1992; Bruno et al. 2003).
Less disturbed habitats such as protected areas have also been successfully colonized by
many invasive species (Drake 1988; Luken 1988; Hiebert et al. 1993; Vitousek et al.
1997). Moreover, these habitats are not free of plant invasions (Broncano et al. 2005;
Traveset et al. 2008; Constán-Nava et al. 2010), and the factors allowing such invasions
are poorly understood.
When a plant has dispersed to a given area, there are numerous factors that can
enhance or limit its probability of success (Keddy 1992; Baskin and Baskin 1998;
Zamora et al. 2008). There exist environmental factors such as climatic conditions,
habitat type (defined by plant community, such as open areas with low vegetation
growth, open woodlands and forests, and which also represent, to a certain extent, a
degree of increasing disturbance), and the abiotic and biotic interactions within these
habitat types (e.g. litter depth, abundance of herbivores, etc.; Williams et al. 1990;
Bewley and Black 1994; Facelli 1994; Koop 2004). In addition to these external,
environmental factors, genetic factors also influence plant germination and early
establishment. These genetic factors derive from the mother tree, the mother tree’s
interaction with the environment (seed position in plant, climatic conditions, etc.) and
paternal genetic contributions (Roach and Wulff 1987; Baskin and Baskin 1998;
Bischoff et al. 2006). Thus, all these environmental and genetic factors, and their
interactions, can influence the success of a plant species when colonizing a given
habitat in very complex ways. Studying such factors separately may therefore result in
misleading conclusions, and more comprehensive approaches are needed to improve our
knowledge of the relative importance of such factors and how their interaction
influences plant establishment. These comprehensive approaches are especially needed
for invasive species, as they may be essential for predicting alien plant invasion and
73
developing management techniques to prevent such invasions in natural ecosystems,
including protected areas (Mack et al. 2000; Rejmánek 2000).
Ailanthus altissima (Mill.) Swingle is a tree from China and northern Vietnam,
where it grows under temperate conditions. This species has invaded several ecosystems
out of its original range, chiefly because of its ornamental use and in landscape
restoration (Kowarik and Säumel 2007). A. altissima is a dioecious species which
produces a large number of samaras assisted by wind and water dispersion; it also
grows quickly and has a high ability to resprout and form dense clonal stands (Kowarik
1995; Kowarik and Säumel 2007, 2008; Kaproth and McGraw 2008; Säumel and
Kowarik 2010). This species has successfully colonized several habitat types under
Mediterranean conditions, such as roadsides, old fields, and pine, oak and riparian
forests (Kowarik 1983; Danin 2000; Constán-Nava et al. 2007; Kowarik and Säumel
2007). Germination and the early establishment of A. altissima are known to occur in
many soil types and be enhanced by a high light availability and low altitudes
(>1000m), although they do not seem to depend on the habitat types (Mihulka 1998;
Kota et al. 2007; Vilà et al. 2008; Moore and Lacey 2009) or source tree (Kota et al.
2007; Delgado et al. 2009). However, the combined effect on the early stages of this
invasive species and the interaction between genetic and environmental factors, whether
or not they are related to habitat type, are poorly understood.
In this study, we aimed to evaluate the influence of both environmental and
genetic factors on the emergence and early establishment of Ailanthus altissima under
Mediterranean conditions by using a series of experiments including both field and
growth chamber approaches. Growth chamber experiments evaluated the effects of
temperature, seed storage time and genetic effects on the germination and viability of A.
altissima seeds. In the field experiment, we considered six contrasted Mediterranean
habitat types and analyzed the germination and early establishment of A. altissima,
examining the differences among these habitats, the environmental attributes (whether
or not they were related to the habitat type) and climatic conditions during the study
period. Our main hypothesis was that environmental, not genetic factors affect the
germination rate of A. altissima (Kota et al. 2007; Delgado et al. 2009). To our
knowledge, this is the first comprehensive approach evaluating the joint effects of
environmental and genetic factors and their interactions on A. altissima success.
74
MATERIAL AND METHODS
STUDY SITE
The study was conducted in the Carrascal de la Font Roja Natural Park, in southeast
Spain. The climate is Mediterranean, with a mean annual precipitation of 407 mm
during the study period (2006-2010) and mean monthly temperatures ranging from 4.4º
C in January to 24.9º C in July (Baradello meteorological station, representative of the
climate in the study area and located between 1.4 and 5.5 km from the plots used in the
field study, 38º 41´ 90´´ N, 00º 31´ 31´´ W, 788 m.a.s.l.). Soils are xerorthents on
limestone (Soil Survey Staff 2006), with the presence of impermeable clays. The
presence of A. altissima in this Natural Park was caused by its use in roadside
restoration and as an ornamental plant. The species was first planted in private
properties located throughout the park at least 75 years ago (Natural Park staff, personal
comm). A. altissima went on to infest various habitats in the Natural Park, such as pine
forests, scrubland, oak forests and riparian forests. Currently, 5.4 ha of the park have
been invaded by the species and management plans are now being applied to control its
spread (Constán-Nava et al. 2007; 2010).
VIABILITY TEST
In October 2006, seeds of A. altissima were collected from 12 maternal sources
randomly selected from disturbed areas (not ornamental trees) located along the
boundary of the Natural Park and at a minimum distance of 500 m from one another.
These seeds were stored separately in paper bags in the dark and under room conditions
with a temperature of 17-20º C for two months. Because A. altissima seeds are viable
after up to two years of storage, we did not expect their germination ability to be
affected by this storage time (Appendix 1; Little 1974; Hildebrand 2006). We compared
differences in seed viability among the maternal seed sources by applying the
tetrazolium test (ISTA 1966) to four replicates of 50 seeds per each maternal seed
source (totaling 200 seeds per maternal source). The seeds were dissected and imbibed
for 24 h with distilled water and then soaked with 2,3,5,-triphenol tetrazolium chloride
(TTC) at 1% for 24 h in dark conditions. Seeds containing living embryos were stained
red or pink and were counted as viable. Treated seeds with no stained contents were
considered non-viable.
75
TEMPERATURE AND GENETIC EFFECTS
We developed a full factorial experiment to assess the effect of different temperatures
and maternal seed sources on A. altissima germination. As seeds are the result of
maternal and paternal genetic contributions and of interaction with the environment
(Roach and Wulff 1987; Baskin and Baskin 1998; Bischoff et al. 2006), hereinafter the
effect of the maternal seed source is referred to as a genetic effect. In October 2007, A.
altissima seeds were collected from the 12 maternal sources (the same trees used for the
viability test, except for four of them, which were new and are marked with a comma in
Results to distinguish them from the others). All seeds were germinated under the same
photoperiod (16/8 h light/dark), but under three contrasted temperature regimes (15º,
20º and 30º C). Temperatures ranging between 20º and 30º C are considered optimal for
the species studied (Little 1974, Graves 1990). The temperature of 15º C was selected to
contrast with the 20º and 30º C temperatures, and as an approximation of the mean
annual temperature in the study area (14º C, Ninyerola et al. 2005). Four replicates of 20
non-stratified seeds were established for each maternal source and temperature (80
seeds x 12 maternal sources x 3 temperatures; 2,880 seeds in total). These seeds were
treated with a sodium hypochlorite solution (2%) for ten minutes to avoid fungal
infection and were rehydrated with distilled water for 24 hours. Each replicate was then
placed on a 9 cm Petri dish containing two layers of filter paper. Filter papers were kept
soaked throughout the experimental period (31 days). Petri dishes were randomly
moved to avoid position effects in the germination chamber (the same chamber was
used for the three temperature regimes but at different times, so no chamber effect was
expected). Seed germination was monitored daily and, when the radicle was visible, it
was considered germinated and removed from the Petri dish.
SEED GERMINATION AND EARLY SEEDLING ESTABLISHMENT UNDER
FIELD CONDITIONS
Six habitat types in the Natural Park were selected, defined by plant community and
following a decreasing disturbance gradient: local roadside (RS), pathside (PS), early
old fields (OLD), south-facing slopes pine forest (SPF), north-facing slopes pine forest
(NPF), and open oak forests (OF). RS had steep slopes. PS habitats were uncommon in
the study site and contained species from nearby habitats (such as NPF). NPF included
shade-tolerant species and a high density of bryophytes, whereas SPF presented more
heliophyllus species. SPF was the only studied habitat type with a south aspect. OF
76
included some deciduous species. OLD had previously been almond crop terraces, and
after being abandoned were colonized by shrubby species and several herb species.
In autumn 2008, seeds from A. altissima were collected from 11 of the 12
maternal sources selected in 2007 (one tree was excluded because it did not produce
fruit) and stored separately in paper bags in the dark under room conditions. In March
2009, three sites were selected by habitat type (n = 18 sites) covering a wide range of
altitudes (from 670 to 1150 m.a.s.l.; see Table 1). Although A. altissima has invaded
most of the selected habitats, no reproductive individuals of this species were present
closer than 200 m from any of the selected sites, avoiding uncontrolled seed dispersion
in the study sites (Landenberger et al. 2007). In each site, we randomly marked two 1 m
× 1 m plots, at a distance of 1-4 m from each other, resulting in six plots in each habitat
type, all on gentle slopes (except at RS, which had steep slopes) and north-facing slopes
(except SPF sites). Each plot was divided into 12 quadrats of 25 cm × 33 cm; in each
quadrat, 12 non-stratified seeds of a given seed source were buried at a depth of 1 cm
and at a distance of <5 cm apart (the last quadrat always remained empty because we
had no seeds from the 12th mother tree). Overall, 132 seeds were planted in each plot,
making for a total of 4,752 seeds. We acknowledge that the distance among the seeds
could be insufficient to avoid early competition among seedlings. However, due to the
high growth rate and the extensive root system of A. altissima, avoiding this effect
would have required large plots, which was not viable in the Natural Park. Seeds were
buried with the minimum disturbance possible to avoid removing structural attributes
characteristic of each plot. The number of emerging seedlings and their survival was
recorded every two to four weeks during the study period (spring 2009-summer 2010).
Upon completion of the study, all A. altissima seeds and seedlings were removed from
the study sites to avoid further invasion. To assess habitat characteristics that are
important for seed germination, such as light availability, litter layer depth and
competition with other species (Tilman 1988), we measured total cover (%) of litter,
bare soil, perennial herbs, shrubs and trees at each of the six plots per habitat type
immediately after seeding (Table 1). Five measurements of litter depth (cm) were also
taken in each plot. Although microclimatic measurements taken within each plot would
have been better, we believe that measuring these attributes is a good surrogate for the
availability of light, the condition of the soil and water competition experienced by the
seedlings (Raich 1989; Bartuszeyige et al. 2007; Pérez-Ramos et al. 2010).
77
Table 1 Environmental variables measured in the studied sites (Mean ± SE of the two
plots/site). Legend: RS: Roadside; PS: path side; NPF: North facing aspect pine forests; SPF:
South facing aspect pine forests; OF: open oak forests; OLD: old fields
Bare soil
Herb
Shrub
Tree cover Litter
Habitat Site Litter (%) (%)
cover (%) cover (%) (%)
depth (cm) Altitude
670
RS
1
7.5±2.5
15±5
12.5±2.5
67.5±2.5
0±0
0.14±0.02
750
2
12.5±2.5
0±0
87.5±2.5
0±0
0±0
0.37±0.03
730
3
50±25
2.5±2.5
72.5±2.5
0±0
45±15
1.21±0.85
810
PS
4
5±0
7.5±2.5
5±0
82.5±2.5
10±10
0.06±0.02
1160
5
57.5±32.5
22.5±22.5
0±0
50±20
95±5
0.28±0.24
860
6
25±5
25±5
0±0
50±10
0±0
0.49±0.21
770
NPF
7
20±10
20±10
0.25±0.25
60±20
50±30
0.72±0.02
790
8
1±1
4±4
0±0
95±5
7.5±7.5
0.01±0.01
810
9
7.5±2.5
0±0
0±0
92.5±2.5
47.5±32.5
0.44±0.26
880
SPF
10
20.5±19.5
5±5
0±0
70±30
0±0
0.19±0.05
910
11
17.5±2.5
15±0
0±0
67.5±2.5
20±5
0.24±0
910
12
27.5±17.5
25±15
0±0
47.5±2.5
0±0
0.19±0.07
1050
OF
13
1±0
47.5±47.5
0±0
52.5±47.5
50±50
0.04±0.02
920
14
8±7
70±10
0±0
22±3
0±0
0.08±0.06
1000
15
7.5±2.5
15±15
0±0
77.5±17.5
100±0
1.19±0.09
760
OLD
16
6±4
7±6
62.5±17.5
20±10
10±0
0.13±0.13
1150
17
0±0
32.5±2.5
0±0
67.5±2.5
0±0
0±0
796
18
0±0
70±5
12.5±2.5
17.5±2.5
7.5±7.5
0±0
STATISTICAL ANALYSES
Differences in viability between A. altissima maternal sources were assessed using onefactor analysis of variance (ANOVA) with “maternal source” as a random factor with
12 levels. Germination percentage at the laboratory experiment was analyzed using a
two-way mixed model with temperature and maternal seed source as fixed and random
factors, respectively. According to the effects of the time and climatic factors on
germination percentage in the field experiment, we tried to integrate all the maternal,
habitat-related and other environmental factors in an analysis with maternal seed source,
habitat, site, and sampling period. In order to avoid overhead problems using an overly
large matrix, we analyzed our data in different subsets. Firstly we averaged germination
rate by habitat type or site and tested for differences in germination times among
habitats and sites, and for the effect of current environmental conditions (rainfall and
temperature) on germination. To achieve this, the analysis included sampling period
(time) and habitat type (fixed factors), and site (nested within habitat type); monthly
temperature and precipitation in the previous month were introduced as covariates.
Second, germination percentage was analyzed using a three-way mixed model,
78
introducing the maternal seed source as a random factor, habitat type as a fixed factor
and site as a nested factor. This allowed us to test for differences in germination among
maternal sources or habitat types, to test for the effect of other environmental factors not
related to habitat type (the “Site” effect), and also to assess for interactions between
maternal and environmental factors (both habitat type and others). All these analyses
were performed using the semi-parametric PERMANOVA approach (Anderson 2001;
McArdle and Anderson 2001; Anderson and ter Braak 2003). Differences among
habitats in litter cover, bare soil, trees and perennial herbs were also analyzed using
PERMANOVA, with habitat type and site as fixed and random factors, the latter nested
within the former. Litter depth and shrub cover were excluded because they were
correlated to the other variables (Spearman’s Rho coefficient ≥ 0.5, P < 0.01). The other
measured variables were entered into the model as they did not correlate to one other,
meaning that PERMANOVA assumptions were not violated. All PERMANOVA
analyses were performed using Euclidean distance to create the resemblance matrix, and
a type III sum of squares (except for the analysis with covariates, where we used a type
I sum of squares), permutation of residuals under a reduced model (except for the
analysis of viability, which included one factor and the unrestricted permutation of raw
data was used), and 9999 permutations were used to calculate pseudo-F and P-values.
The Kaplan–Meier procedure (Fox 1993) was conducted to model differences in
A. altissima germination curves among the three temperatures assayed (factor) and
maternal seed sources (strata) in the growth chamber experiment, and among the
different habitat types (factor) and maternal seed sources (strata) in the field experiment.
Survival curves for the latter experiment were also analyzed with this procedure.
Comparisons of the cumulative germination/survival curves among different maternal
sources were tested using non-parametric Log-Rank tests (Pyke and Thompson 1986) in
both the growth chamber and the field experiment.
A regression tree was performed as an exploratory analysis to assess for the most
important environmental factors: bare soil, perennial herb, shrub and tree cover, and
litter depth on the germination percentage of A. altissima (Appendix 2). As this
exploratory analysis only revealed bare soil to be important, it was analyzed using a
linear regression.
All statistical analyses were performed using PERMANOVA+ for the PRIMER
statistical package for Windows (PRIMER-E Ltd., Plymouth Marine Laboratory, UK),
79
except for germination and survival curves, and linear regression, which were analyzed
using SPSS v.15 (SPSS Inc., Chicago, IL, USA).
RESULTS
VIABILITY AND GERMINATION TESTS
Seed viability ranged from 45.5 ± 6% to 90 ± 5% (mean ± SE) and was significantly
different among maternal seed sources (ANOVA: F11,47 = 9.46, P < 0.001). Both
maternal seed source and temperature affected germination rates in the growth chamber
experiment (Fig. 1). A significant maternal seed source × temperature interaction was
found (Fig. 1). In general, most maternal seed sources showed highest and lowest
germination rates at 15º C (17.4 ± 2.7%) and 30º C (3.65 ± 0.9%), respectively.
GERMINATION AND SEEDLING SURVIVAL UNDER FIELD CONDITIONS
Three germination pulses occurred during the entire study period, two in spring (2009
and 2010) and one in autumn (2009), with the latter showing the highest germination
values. Germination percentage was significantly affected by the time, the site and the
rainfall of the previous month, but not by the mean month temperature or habitat type
(Fig. 2). Significant interactions were observed between time × habitat type and time ×
site, i.e. germination varied throughout the time among habitat types and sites. In spring
2009, seeds only germinated at RS, PS and NPF, and showed low mean germination
rates (<1% in all cases, Fig. 2B.). A predation event occurred with a 2.3% seed
predation rate at one OLD site in spring 2009. In both autumn 2009 and spring 2010,
seeds germinated in all assayed habitats except for RS, where no germination was
registered in autumn and which showed the lowest germination rates the following
spring. The highest germination rates were found at SPF and OF in autumn 2009, and at
OF in spring 2010.
80
80
15 ºC
1
2´
3
4
5´
6
7´
8
9
10´
11
12
60
40
20
0
80
Germination (% )
20 ºC
60
40
20
0
80
T: F
1,108
= 6.5; P = 0.0023
MS: F
11,108
60
30 ºC
= 11.6; P = 0.0001
T × MS: F
22,108
= 6.4; P = 0.0001
40
20
0
0
5
10
15
20
25
30
Day
Figure 1 Germination curves of A. altissima seeds from 12 different maternal seed sources
(showed with different color dots) after 31 days in a growth chamber at 15 ºC, 20 ºC and 30 ºC.
Data are mean percentage ± SE (n = 4). Statistical results for the PERMANOVA are given (T:
temperature, MS: maternal source)
81
180
T min
T max
T mean
Rainfall (mm)
140
30
120
20
100
80
10
60
40
0
Mean temperature (ºC)
160
40
A
20
0
16
B
Roadside
Path side
North facing slopes pine forests
South facing slopes pine forests
Oak forests
Old fields
Germination (%)
14
12
10
T: F
8
= 0.07; P = 0.791
1,270
P: F
1, 270
= 3.8; P = 0.0541
Ti: F
= 14.9; P = 0.0001
12, 270
6
H: F
= 1.2; P = 0.3707
5, 270
S: F
4
2
12, 270
Ti × H: F
= 2.1; P = 0.0159
70, 270
Ti × S: F
= 1.6; P = 0.011
168, 270
= 1.4; P = 0.0162
0
J
F
M
A
M
J
J
A
S
O
N
D
J
F
M
A
M
J
2010
2009
Month and year
Figure 2 Monthly rainfall (grey bars) and temperature (black circles) (Baradello meteorological
station, representative of the climate in the study area) (A) and germination percentage curves
(mean ± SE, n = 3) of A. altissima at the six different habitat types studied (B) during the study
period. Statistical results for the PERMANOVA are given (T: temperature, P: precipitation, Ti:
time, H: habitat, S: site)
Final germination percentages differed significantly among maternal source and
site but not among habitat types, and a significant maternal source × habitat type
interaction was found, with seeds from each maternal source differing in their
germination rates depending on the habitat type (Fig.3A). Differences between
germination curves among habitat types and maternal seed sources were found (log–
rank test, P < 0.001). There were no significant differences among maternal seed
sources in separated analyses for each sampling period, except in November 2009
(Pseudo-F= 5.44; P < 0.001). Although all seedlings died during the study period,
82
differences between survival curves were found between habitats and maternal seed
sources (log-rank test, P < 0.001, Fig. 3B).
(A)
30
Roadside
H: F
5,198
20
Path side
= 1.1; P = 0.3609
MS: F
10,198
= 4.2; P = 0.0001
S: F12, 198= 8.2; P = 0.0001
H × MS: F
= 1.4; P = 0.0781
10
50,198
MS × S: F
120,198
= 0.6; P = 0.9984
Germination (%)
0
30
North facing slopes
pine forests
South facing slopes
pine forests
20
10
0
30
20
10
Oak forests
Old fields
1
2´
3
4
6
7´
8
9
10´
11
12
0
J F M A M J J AS O N D J F M A M J
2009
J F M A M J J AS O N D J F M A M J
2010
Month and year
83
2009
2010
(B)
100
Roadside
Pathside
80
2´
3
6
7´
8
9
10´
11
12
1
6
7´
8
10´
11
12
60
40
20
Survival (%)
0
100
South facing slopes
pine forests
North facing slopes
pine forests
80
1
2´
3
4
6
8
9
10´
11
12
60
40
20
1
2´
3
4
7´
8
9
10´
11
12
0
100
Oak forests
80
Old fields
1
2´
3
4
6
7´
8
9
10´
11
12
60
40
20
1
2´
3
4
6
7´
8
9
10´
11
12
0
0
1
2
3
4
0
1
2
3
4
Month after emergence
Figure 3 Germination (A) and survival (B) percentages curves (mean ± SE, n = 3) of A.
altissima seeds from 11 maternal sources at each habitat type throughout the field experiment
study period. Statistical results for the PERMANOVA are given (H: habitat, MS: maternal
source, S: site)
RS showed the poorest survival rates, with all seedlings dead after a month
(Fig.4). The highest survival duration was found in PS and OLD habitats, where
seedlings survived until the fourth month. Maternal seed source significantly affected
seedling survival in all habitats, except for RS (Fig.3B).
84
100
Roadside
Path side
North facing slopes pine forests
South facing slopes pine forests
Oak forests
Old fields
Survival (%)
80
60
40
20
0
0
1
2
3
4
Months after emergence
Figure 4 Survival percentage curves (mean ± SE, n = 3) of A. altissima seedlings after
emergence at the six different habitat types studied
Table 2 Two-way nested model results of the environmental variables (cover of litter, bare soil,
trees and perennial herbs) according to the habitat type and site using PERMANOVA
Source
Habitat
Site (Habitat)
Res
Total
*P < 0.10; **P < 0.05
df
5
12
18
35
MS
7443.6
4276.6
1272
F
1.74
3.36
P
0.0641*
0.0001**
Significant differences were found in environmental variables among habitat
types and site (Table 2). Linear regression analysis showed a positive relationship
between bare soil cover and A. altissima germination (Fig. 5).
85
35
F = 54.901, df = 1
R2 = 0.618, P < 0.0001
30
Germination (%)
25
20
15
10
5
0
0
20
40
60
80
100
Bare soil (%)
Figure 5 Linear regression analysis of A. altissima germination rate vs bare soil cover.
Statistical results are given
DISCUSSION
The comprehensive approach used in this study, using both laboratory and field
experiments and evaluating a wide array of factors influencing germination and early
establishment of A. altissima, highlights the importance of both genetic and
environmental factors in modulating A. altissima germination, as well as the existence
of interactions between these factors. Seed germination and early survival varied among
maternal sources, and were affected by environmental factors such as the site, the
percentage of bare soil and rainfall pulses. Furthermore, the interaction between genetic
and environmental factors influenced the early success of A. altissima, as demonstrated
by the contrasting responses found for each maternal source depending on temperature
(laboratory experiment) or other environmental conditions, such as those represented by
the different sites or habitat types tested (field experiment), which suggests that both
habitat preference and optimal environment conditions (and therefore the effect of
external factors) vary depending on the genetic source.
86
ENVIRONMENTAL FACTORS AFFECT A. ALTISSIMA GERMINATION AND
SURVIVAL
In accordance with previous studies (Kota et al. 2007; Vilà et al. 2008), we found a low
germination percentage under field conditions (although we repeated the field
experiment for three years, as no seeds germinated until the third year). The low
viability and germination recorded for this species by our work and by other studies are
at odds with the high degree of invasion of A. altissima in many habitats across the
globe (Kowarik and Säumel 2007). A possible explanation for these contrasting results
is the high fecundity of the species, with a single tree able to produce a large number of
seeds (since 325,000 samaras per tree; Little, 1974; Bory and Clair-Maczulajtys 1980).
We found that temperature had a significant effect on the germination of A.
altissima, with lower temperatures rendering higher germination rates (Fig. 1). This
result contrasts with previous studies (Little 1974, Graves 1990) and with our own
results in the field. Emergent pulses in the field were related to rainfall pulses but not to
temperature (Fig. 2). The most plausible explanation for these contrasting results is that
the temperature in the field fluctuated (both daily and seasonally) and is not fixed;
germination cannot therefore be related to a unique value of temperature. Conversely,
and in contrast with previous studies using the same species (Vilà et al. 2008), we found
significant differences in germination curves among the different habitat types tested,
which are related to the differences found according to the environmental variables
(Table 1). The habitat type with the lowest germination and early survival rates was
roadside (RS), which could be related to the scarce water availability and low fertility in
soils often found in these areas (Bochet and Garcia-Fayos 2004; García-Palacios et al.
2010). Roadsides are ecosystems commonly invaded by A. altissima under
Mediterranean conditions (Kowarik 1983; Danin 2000; Constán-Nava et al. 2007;
Traveset et al. 2008). In spite of our results, the high degree of invasion in these areas
may be due chiefly to the use of this species in roadside restoration and because these
ecosystems act as dispersal corridors, where a large number of seeds can easily reach
the area, aiding its invasion (Vilà and Pujadas, 2001; Kowarik and von deer Lippe 2006;
Kowarik and Säumel 2007; Kowarik and von deer Lippe 2011).
Our results also show that the effect of habitat type on the germination rate
changed depending on the seasons or the microenvironmental characteristics of each
site inside the habitat. The field experiment revealed that 61% of the variance in the
seed germination rate was explained by the percentage of bare soil in the plot. This
87
highlights how important areas with high bare soil cover are for A. altissima
germination, particularly in habitats with relatively low stress (such as open oak
forests), as reported previously for this and other invasive species (e.g. Burke and
Grime 1996; Bartuszevige et al. 2007; Kota et al. 2007). This dependence on bare soil
could be explained by the lower competition with other plants and the lower litter
presence in these areas, which may have both direct negative effects (a lower
availability of resources and a physical hindrance of seed emergence) and indirect
effects by increasing insect herbivory (Facelli and Pickett 1991; Facelli 1994).
Similarly, we found 2.3% of seeds were predated in old fields, possibly caused by
arthropods or small mammals (Facelli 1994; Ostfeld et al. 1997). The effect of the site
on A. altissima germination could be also related to microenvironmental factors not
included in this field study, such as soil moisture or light availability (the latter was
measured indirectly using vegetation cover, but was not statistically significant), which
have been found to be relevant to the success of A. altissima (Kota et al. 2007;
González-Muñoz et al. 2011).
GENETIC FACTORS INFLUENCE THE PERFORMANCE OF A. ALTISSIMA AND
MODULATE ITS RESPONSE TO ENVIRONMENTAL FACTORS
Although variations in seed germination between maternal sources have been found in
numerous species (e.g. Baskin and Baskin 1998), no previous studies have found any
effect of this factor on the germination of A. altissima (only on seed weight; Kota et al.
2007; Delgado et al. 2009). However, our results, both under field and laboratory
conditions, show that genetic factors do affect the germination ability of A. altissima.
These contrasting results might be influenced by the higher number of maternal sources
and seeds used in this study in comparison with previous ones (12 vs. 6 genetic
sources). The survival and growth rates of A. altissima are known to differ among
provenances (Feret 1985), which clearly suggests some genetic component affecting A.
altissima performance and adding confidence to our results.
Genetic effects are those resulting from the maternal tree and how it interacts
with the environment, as well as those from the different paternal contributions, as the
variation within the seeds could be ascribed to the fact that different seeds were started
by pollen from different trees (Roach and Wulff 1987; Baskin and Baskin 1998;
Bischoff et al. 2006). In this sense, environmental factors such as differences in the
availability of water, nutrients or light for the maternal tree could explain changes in
88
germination rates of seeds coming from the same maternal source, but collected in
different years (e.g. seeds from maternal sources 3 or 12; see Fig. 1 and Fig. 3A), but
may also be related to the paternal genetic contributions, which could differ from one
year to another. Knowledge of the genetic and genotypic diversity of A. altissima in the
Mediterranean area of Europe (Dallas et al. 2005) could help in this sense.
Significant differences were found in the survival curves of A. altissima
seedlings among habitat types and maternal sources. Our results suggest that, on the one
hand, habitat type per se could be influencing the survival of A. altissima due to the
differences in environmental factors, as discussed above. On the other hand, the
differences among maternal sources suggest that genetic factors also affect the survival
ability of the invasive species. Regardless of the maternal source, however, all seedlings
died, during either winter or summer, most likely due to the low winter temperatures
(Kowarik and Säumel 2007) or the summer drought that is typical in Mediterranean
environments (Maestre et al. 2001). This suggests that climatic conditions are more
important than the genetic component and that they could be the main limiting factor for
the invasive success of this species. However, this conclusion must be considered with
care, as only maternal individuals from one type of environment were used in this study,
and further research including sources from contrasting environments is needed to
confirm or reject this hypothesis.
CONCLUSIONS
In this study we used a comprehensive approach that included the effect of
environmental factors at different scales: 1) local temperature and rainfall, represented
by
the
climate
data,
2)
habitat-related
environmental
conditions
and
3)
microenvironmental conditions, represented by our “site” effect and the attributes
measured at each plot, genetic factors and the interaction among them on the invasive
success of A. altissima under Mediterranean conditions. Our study reveals that the
genetic component not only affected the performance of A. altissima, it also modulated
its response to environmental factors, which seemed to be the main drivers of
germination and early establishment for this species. These results may help to predict
the invasion success of A. altissima in contrasted habitats under Mediterranean
conditions, and therefore to better reduce the spread of this species in Mediterranean
ecosystems.
89
ACKNOWLEDGEMENTS
S. Soliveres and two anonymous reviewers provided helpful comments and
improvements on an early version of this manuscript. We also thank Language Centre
(University of Alicante) and C. Beans for improving the English of this manuscript. We
are grateful to the staff of the Font Roja Natural Park (Generalitat Valenciana), Alcoy
Council and J.L. Ferrándiz (landowner) for the permits provided and their collaboration.
We would also like to thank C. Constán, A. Constán, M.J. Nava, G. Plaza, E. Pastor, N.
Vizcaíno and the other collaborators who helped during the laboratory and fieldwork.
We thank J. Huesca for the technical support, M.J. Baeza and V.M. Santana for their
comments and M.J. Anderson for her statistical suggestions. This research and SCN
fellowship were supported by the projects GV06/029 founded by the Generalitat
Valenciana and the ESTRES (063/SGTB/2007/7.1) and RECUVES (077/RN08/04.1)
founded by the Spanish Ministry for the Environment. Font Roja Natura UA Scientific
Station (ECFRN UA), which depends on the Office of the Pro-Vice- Chancellorship for
Research, Development and Innovation (VIDI) of the University of Alicante, also
supported this research.
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Appendix 1: Effect of storage time
In February 2008, we randomly selected 100 seeds (five replicates of 20 seeds)
of each year of recollection (2005, 2006 and 2007), regardless of the maternal source of
these seeds. These non–stratified seeds were treated with a sodium hypochlorite
solution (2 %) for ten minutes to avoid fungal infection and re-hydrated with distilled
water during 24 hours. After the application of this treatment, we located the seeds on
Petri dishes in a growth chamber during 31 days under optimum conditions (20º C, 16/8
h light/dark photoperiod and moist filter paper; Little 1974, Graves 1990). Petri dishes
were randomly moved to avoid position effects in the germination chamber. The seeds
were examined daily and when the radicle was visible were considered germinated and
removed from the Petri dishes. Results are shown below.
100
Seed germination (%)
80
2005
2006
2007
60
40
20
0
0
5
10
15
20
25
30
35
Day
Figure 6 Germination curves of A. altissima seeds from 3 different years of recollection
(showed with different dots) after 31 days in a growth chamber. Data are mean percentage ± SE,
n = 4)
Germination rates were not significantly different between the three times of storage (P
> 0.05), being of 12 ± 2.5% in seeds of 2007, 19 ± 6.4% in seeds of 2006 and 20 ± 3.8%
in seeds of 2005.
95
Appendix 2: Regression tree
A regression tree (De’ath and Fabricius 2000; Crawley 2007) was performed to analyze
the effect of bare soil, perennial herbs, shrubs and trees cover, and litter depth on A.
altissima germination percentage. This method allows introducing correlated predictors
and detecting non-linear responses. The tree was pruned to improve parsimony using
10-fold cross-validation (De’ath and Fabricius 2000). This analysis was done using the
Tree package (Oksanen et al. 2005) for R software (R Development Core Team 2009).
Figure 7 Regression tree model for A. altissima germination. Split values for the predictor used
are shown in each branch. Terminal nodes show the mean value for each group of the response
variable introduced and the number of cases in each node (between parenthesis; n = 36 cases for
the tree). The general fit of the model (D2, percentage of variance explained by the model),
extracted from the null deviance (Deviance root), and the deviance of the final chosen tree after
10-fold cross-validation (Deviance tree) are shown
The regression tree assessing the effect of environmental variables explained ~55% of
the variance in seed germination, and pinpointed the percentage of bare soil as the best
predictor for seed germination.
References (except those in the main text)
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package Version 1.6–10
96
APÉNDICE FOTOGRÁFICO
Foto 1 Semilla de A. altissima teñida
mediante test de Tetrazolio
Foto 2 Experimento con semillas de A.
altissima bajo condiciones controladas
en cámara de germinación
Foto 3 Parcela de 1 × 1 m para el
estudio de la germinación de A.
altissima en campo (en pinar de solana)
Foto 4 Plántula de A. altissima
germinada en experimento de campo
97
98
CAPITULO 3
Direct and indirect effects of invasion by the
alien tree Ailanthus altissima on riparian plant
communities and ecosystem multifunctionality3
3
Manuscrito enviado
Autores: Soraya Constán-Nava, Santiago Soliveres, Rubén Torices, Lluís Serra, Andreu Bonet
99
100
ABSTRACT
The effects of invasive species on biodiversity are well-known and their effect on
separate ecosystem functions and soil attributes has gained interest more recently.
However, most of the existing studies focus on species richness, ignoring better
indicators of biodiversity and better predictors of ecosystem functioning such as the
diversity of evolutionary histories or phylodiversity. Moreover, and in spite of the well
known relationship between them, no previous study has separated the direct effect of
alien plants on multiple ecosystem functions simultaneously (multifunctionality) from
the indirect effect mediated by the effect of alien plants on biodiversity. This latter issue
is deeply necessary for improving management plans and developing effective
conservation strategies for natural ecosystems undergoing alien plant invasions. We
aimed to analyze the direct and indirect effects, mediated or not by its effect on
biodiversity, of Ailanthus altissima -an invasive tree that reduces species diversity and
some ecosystem functions in natural ecosystems worldwide- on ecosystem
multifunctionality of riparian habitats under Mediterranean climate. For doing this, we
measured vegetation attributes (species richness and cover, phylodiversity), soil
functions (plant biomass and soil enzyme activities) and soil attributes (pH, EC, OM, P)
in plots infested by A. altissima and in control (non invaded) ones. Our results show that
plant species richness, phylodiversity and multifunctionality were reduced under the
presence of A. altissima. The effect of the alien plant on multifunctionality was indirect,
mediated mainly by its effect on phylodiversity and, to a less extent, on species
richness. Species composition and plant cover, but not soil attributes, were also affected
by the invasive plant. We provide an easy methodology to tease apart direct and indirect
effect of alien species on ecosystem multifunctionality by using observational data. This
may help to center restoration and conservation efforts either on the elimination of
invasive species and introduction of native ones, in case of existing solely indirect
effects, such as in our case, or to include soil restoration measures to recover ecosystem
services and functions in case of existing both direct and indirect effects of the alien
species on ecosystem multifunctionality.
101
102
INTRODUCTION
I
nvasive plant species are a major threaten for biodiversity and ecosystem functioning
in many ecosystems around the world (Vitousek et al. 1997; Liao et al. 2008; Vilà et
al. 2011). The last years have seen an increasing interest on the effects of the invasive
species on species composition, community structure or ecosystem functioning of
invaded habitats (Williamson 1998, 2001; Simberloff et al. 2003; Weber 2003).
Previous research shows that plant invasions affect the diversity of plants (Richardson
et al. 1989; Vilà et al. 2011; Hejda et al. 2009) or animals (Maerz et al. 2005; Watling et
al. 2011), modify soil properties (Vilà et al. 2006; Truscott et al. 2008), and alter
nutrient cycles (Vitousek and Walker 1989; Ehrenfeld 2003; Castro-Díez et al. 2009), or
native microbial communities and their associated ecosystem processes (Wolfe and
Klironomos 2005; van der Putten et al. 2007; Weidenhamer and Callaway 2010).
Previous studies focus separately on the effect of invasive plants either on
biodiversity or on ecosystem functioning, ignoring the well-known and tight
relationship between both ecosystem attributes. Many ecosystem functions are known to
increase with higher species richness (e.g. Tilman et al. 1997; Zavaleta et al. 2010;
Maestre et al. in press); and, therefore, a reduction of diversity or alteration of the
composition and structure of vegetation -both effects documented in invaded
ecosystems- can result in an indirect decline of ecosystem functioning (Hooper et al.
2005; Richardson et al. 2007). The presence of invasive species can also directly alter
ecosystem functions by several mechanisms such as releasing allelopathic compounds,
altering the nutrient balance, or by soil salinisation and impoverishment (e.g. Vitousek
1986; Callaway and Aschehoug 2000; Levine et al. 2003), to name a few. Thus, we are
still far from a full mechanistic understanding on the nature of the effect of alien plants
on ecosystems, since the directness or indirectness of their effect on ecosystem
functioning, mediated or not by their effect on biodiversity, is poorly understood.
Knowing the nature of the effects of invasive species on ecosystem functions and
services may help to improve management plans and restoration strategies for those
ecosystems undergoing invasion by alien species. In this sense, if invasive species
produces indirect effects in ecosystem functioning, management plans should focus
mainly on removing the invasive species and introducing, or enhancing the
establishment of, native ones. However, if invasive species produce both direct and
103
indirect effects, management plans should include, additionally, actions for restoring
other ecosystem properties, such as soil fertility and functions.
Most of the existing studies about the effect of invasive plants on biodiversity
focus just on the number of species as a measure of biodiversity ignoring better
indicators of biodiversity and better predictors of ecosystem functioning, such as the
diversity of evolutionary histories or phylogenetic diversity (Forest et al. 2007;
Maherali and Klironomos 2007). Numerous and important ecological traits are
preserved through evolutionary times and, therefore, phylogenetic diversity (hereafter
phylodiversity) is thought to be a good surrogate of the diversity of functional forms
present in a given community (Prinzing et al. 2001; Webb et al. 2002). Species
assemblages containing phylogenetically diverse species are more likely to be able to
provide high levels of different ecosystem functions and services than those
communities formed by functionally similar ones either by complementarity in the use
of resources, interspecific positive interactions or the existence of trade-offs in
important ecological traits featuring different species (Cardinale et al. 2007; Zavaleta et
al. 2010). Indeed, phylodiversity may be a better indicator of biodiversity since it is
more related to important ecosystem functions than other more commonly used
measurements, such as species richness; and therefore, can provide critical information
to understand the functional effects of biodiversity loss (Forest et al. 2007; Maherali and
Klironomos 2007; Cadotte et al. 2008).
While the study of the ecosystem functions and the services they provide has
been a major topic in ecology during the last years (reviewed in Hooper et al. 2005),
most of these studies focus on single functions or processes. However, the effect of a
given factor (i.e. invasive species) on ecosystem functioning must be studied
considering together multiple ecosystem functions (multifunctionality; Hector and
Bagchi 2007; Zavaleta et al. 2010; Maestre et al. in press). Studying multiple ecosystem
functions simultaneously is crucial since different ecosystem processes or services may
be affected by a given factor, but in different directions (Zavaleta et al. 2010).
Therefore, analyzing them separately could result in contradictory results and
misleading conclusions. Invasive species, for example, could increase C sequestration
and plant productivity, and reduce soil erosion in the midterm, due to their high growth
rates (Liao et al. 2008; Vilà et al. 2011). However, they may also simultaneously reduce
nutrient cycling, soil fertility or ecosystem resilience due to their elevated nutrient
uptake, the reduction in species diversity or the releasing of allelopathic compounds
104
(Liao et al. 2008; Weidenhamer and Callaway 2010). These contradictory effects make
necessary to study the effect of alien plants on multiple ecosystem functions
simultaneously to avoid misleading conclusions on the net effect of these species on
natural ecosystems. This is especially relevant considering the reduction of species
richness commonly found in ecosystems undergoing alien species invasion (Vilá et al.
2011). Since it is known that the minimum number of species to maintain ecosystem
functioning increase with the number of functions considered (Hector and Bagchi 2007;
Zavaleta et al. 2010), the indirect impact of alien species on ecosystem functioning
could be, therefore, much higher than previously suggested by studies focused in single
functions or processes.
We aimed to analyze the direct and indirect effects, mediated or not by its effect
on biodiversity, of the invasive tree Ailanthus altissima on ecosystem functionality of
riparian habitats under Mediterranean climate. We measured vegetation and soil
attributes, together with measurements of different ecosystem functions in invaded and
non invaded riparian forests by the exotic species. Our main hypotheses were 1) A.
altissima reduces not only the species richness, but also the phylodiversity of neighbor
vegetation, 2) The effects of A. altissima on vegetation diversity indirectly reduce
ecosystem
multifunctionality,
and
3)
A.
altissima
reduction
on
ecosystem
multifunctionality is not only indirect, mediated by its effect on plant diversity, but also
directly by its known effects on nutrient cycles and soil properties.
MATERIAL AND METHODS
STUDY AREA
The study was conducted in riparian forests of water courses and streams
localized in the Site of European Community Importance (Directive 92/43/CEE) SCI
Serres de Mariola i el Carrascar de la Font Roja, in southeast Spain (X 713000, Y
4288000 UTM). This area is one of the ones with highest diversity of the Mediterranean
Basin, considered as Mediterranean Hotspot and one of the 25 Earth Hotspot (Myers et
al. 2000; Médail and Quézel 1999). Furthermore, this area is part of one of the refugee
areas for flora after climatic change in the Mediterranean since Pleistocene (Serra et al.
2003; Médail and Diadema 2009), so the presence of invasive species could cause high
negative effects. The climate is Mediterranean, with mean annual precipitation and
temperature of 647 mm and 14.7 ºC, respetively (Bocairent meteorological station
105
located in the study area, at 641 m.a.s.l.; data from period 1985-2006 for temperature
and 1966-2006 for precipitation; Rivas Martínez et al. 2007). Soils are xerorthents on
limestone (Soil Survey Staff 2006), with the presence of impermeable clays. Riparian
habitats are dynamic and complex presenting regular floods which favor species
movement (Forman and Godron 1986; Pyšek and Prach 1994). These regular floods,
together with human disturbances, make these areas vulnerable to invasion by exotic
plants (Hood and Naiman 2000; Holmes et al. 2005; Pyšek et al. 2010). Indeed, riparian
areas act as seed transport corridors that promote the expansion of invasive species
(Thébaud and Debussed 1991; Pyšek and Prach 1993; 1994; Säumel and Kowarik
2010). Vegetation in riparian ecosystems is especially important for numerous
ecological processes, such as providing habitat for wildlife, stabilizing riverbanks,
filtering sediments and nutrients into stream and influencing soil properties (e.g.
Forman and Godron 1986; Décamps 1993; Tabacchi et al. 2000). Therefore, the
presence of invasive species in such ecosystems is likely to alter ecosystem functions
indirectly by reducing the high levels of plant diversity they hold. Moreover, invasive
plants are also likely to alter ecosystem multifunctionality directly by the bank
destabilization, lowering of water tables, salinisation of soil and change channel
capacity for flood flow (e.g. Mooney and Drake 1986; Hulme and Bremner 2006).
Ailanthus altissima (Mill.) Swingle (Simaroubaceae), our target species, is a
deciduous tree from China and North Vietnam which has became in an invasive species
around the world since the 18th century due to its use as ornamental and landscape
restoration (Kowarik and Säumel 2007). It has occupied numerous ecosystems in the
Mediterranean Basin, as disturbed urban areas, old fields, pine and oaks forests, or
riparian communities (Kowarik 1983; Constán-Nava et al. 2007; Kowarik and Säumel
2007). A. altissima is known to affect ecosystem functioning and vegetation
composition, structure and dynamics (Lawrence et al. 1991; Vilà et al. 2006; Motard et
al. 2011). Previous research has shown that A. altissima reduces plant diversity (Vilà et
al. 2006; Motard et al. 2011), alters N cycling (Castro-Díez et al. 2009), slows down
litter decomposition, or increases the total N, organic C, C/N ratio and pH in the soils it
colonizes (Vilà et al. 2006; Gómez-Aparicio and Canham 2008; Godoy et al. 2010; but
see Castro-Díez et al. 2011). The importance of plants for multiple ecosystem functions
and the sensitivity to invasion of riparian ecosystems, together with the well-known
reduction in both diversity and several ecosystem functions promoted by A. altissima,
106
make this study area and species a suitable target to study both direct and indirect
effects of invasive plants on ecosystem multifunctionality.
EXPERIMENTAL DESIGN
We established ten 10 m × 10 m control plots (without A. altissima) and ten 10 m × 10
m plots infested by the invasive species (cover of A. altissima >85%) in July 2008 and
2009. We randomly placed forty 0.5 × 0.5 m quadrats within each plot and we visually
estimated plant cover separated by species. To complete the sampling of the species
richness, we covered the whole plot then to look for other species present in the plot but
not detected within the quadrats. In each plot, we also measured leaf area index (LAI)
using a LAI 2000 plant canopy analyzer (Li-Cor, Inc., Lincoln, NE). In half of the
sampled quadrats (n = 20), we also estimated litter depth at three different sampling
points within each quadrat, and the cover of litter, bare soil and rock. In five of the
sampled quadrats, we removed all the aerial plant biomass of both shrub and herb strata
and transported it to the laboratory where was oven-dried until constant weight at 50 ºC.
Tree biomass was not removed because of the important ecological role and
conservation value that riparian trees hold in the study area. Nevertheless, no important
implications in our conclusions are expected since we did not find significant
differences in tree cover (as estimated with LAI at heights > 1.5 m), but only on tree
species composition, between invaded and not invaded plots. In each plot, we took a
composed sample from five different soil cores of 30 cm depth. Soil samples were
oven-dried during 72 h at 50 ºC in the laboratory and sieved to obtain the fraction < 2
mm, which was used for the analyses described below.
ANALYSIS OF SOIL FUNCTIONS AND ATTRIBUTES
Our measurements of important ecosystem functions covered surrogates of both the C
and P cycle (β-glucosidase and acid phosphatase soil enzymatic activities) together with
the ability of plants for C fixation (biomass measurements, described above). Both
enzyme activities were estimated by using the methodology described in Tabatabai
(1982; β-glucosidase) and Tabatabai and Bremner (1969; acid phosphatase). The
functional variables measured in this study are known to be strongly related to key
ecosystem processes, such as ecosystem productivity (plant biomass; Tilman 1988;
Flombaum and Sala 2008), Carbon cycling (β-glucosidase; Tabatabai 1982) and
107
Phosphorus cycling (phosphatase; e.g. Sinsabaugh et al. 2008, Maestre and Puche
2009).
Apart from the ecosystem functioning variables (see above), structural soil
variables were measured. We analyzed pH, electric conductivity (EC), organic matter
and available phosphorus as key soil attributes. Electric conductivity (EC) and pH were
measured in 1: 2.5 mass: volume soil and water suspension (CMA 1973). Organic
matter (OM) was measured by the Walkley–Black acid digestion method (CMA 1973).
Available phosphorus, a good surrogate of P availability in soils for plants and microbes
(Bardgett 2005), was analyzed by NaHCO3 extraction (Watanabe and Olsen 1965). This
component was included as structural soil variable because many of the reactions that
govern the availability of P are geochemical, rather than biological (i.e. in calcareous
soils, such as the ones studied here, phosphates combine with calcium becoming
unavailable for plants; Bardgett 2005).
MEASUREMENT OF THE EVOLUTIONARY RELATIONSHIPS
We assembled a phylogenetic tree for the 115 species included in this study using
Phylomatic function available at Phylocom 4.1 software (Webb et al. 2008). All the
families in our dataset matched the family names of the angiosperm megatree used in
Phylomatic (R20100701.new), which was based on the APG III phylogenetic
classification of flowering plant orders and families (Angiosperm Phylogeny Group
2009). Phylogenetic relationships were further resolved based on data from various
published molecular phylogenies (Apiaceae [Downie et al. 2000], Asteraceae [Funk et
al. 2009], Dipsacales [Winkworth et al. 2008], Lamiids [Bell et al. 2010], Fabaceae
[Steele et al. 2010], Poaceae [Bouchenak-Khelladi et al. 2008], Rosaceae [Potter et al.
2007]).
After assembling the phylogenetic tree, we adjusted its branch lengths with the
help of the Phylocom BLADJ algorithm, which fixes the age of internal nodes based on
clade age estimates, whereas undated internal nodes in the phylogeny are spaced evenly
(Webb et al. 2008). Thus, BLADJ is a simple tool that fixes the root node of a
phylogeny at a specified age, as well as other nodes for which age estimates are
available. It sets all other branch lengths by placing the nodes evenly between dated
nodes, as well as between dated nodes and terminals (of Age 0). We search for
divergence time estimations in the TimeTree database (Hedges et al. 2006;
http://www.timetree.org). TimeTree uses a tree-based (hierarchical) system to identify
108
all published molecular time estimates bearing on the divergence of two chosen taxa
(e.g. species), compute summary statistics, and present the results. We mainly used this
database to fix the ages of internal nodes on our phylogenetic hypothesis, completing
TimeTree results with other published sources when this database did not provide any
date (Lavin et al. 2005; Besnard et al. 2009; Bremer and Eriksson 2009, Bell et al. 2010;
Bouchenak-Khelladi et al. 2010; Torices 2010; Wang et al. 2010). The procedure
described above resulted in the fixation of 80 nodes (representing more than 70% of
internal nodes of our tree).
DATA REDUCTION
We avoided conducting an elevated number of analyses derived from the high number
of response variables measured by summarizing our dataset using different
methodologies. First, we organized soil variables not directly related with ecosystem
functioning (pH, EC, OM, litter cover, and available phosphorus) and reduced them to a
single synthetic variable by using principal component analysis (PCA). Prior to the
PCA, we standardized (sensu Anderson et al. 2008) the five variables, substracting the
mean and dividing by its standard deviation. With this a priori data transformation, we
homogenized units for all the variables and therefore reduced the higher influence of a
given variable just by having different units (i.e. parts per million instead of percentage;
Anderson et al. 2008). The first axis of the PCA conducted explained 85.2% of the
variation in the data (eigenvalue = 58) and therefore this first axis was used to infer the
effects of A. altissima on soil attributes (hereafter, we will refer to this first PCA axis as
“soil attributes”). This first axis was highly related to available phosphorus (eigenvector
= 0.718) and litter cover (eigenvector = -0.695), but not to pH, EC or OM (eigenvectors
= -0.044, -0.008 and 0.012, respectively).
Secondly, to summarize variables related to ecosystem functioning, we
calculated a multifunctionality index (M), recently proposed by Maestre et al. in press
(see below). The use of this index help us to improve two aspects of our results: 1)
reduces the three variables (biomass and β-glucosidase and acid phosphatase activities)
to a single one, therefore reducing multiple testing, and 2) allows us to assess for the
effect of A. altissima on the ability of the ecosystem to maintain multiple functions
simultaneously, rather than conducting separate analyses for each function. The
multifunctionality index (M) used was based on the calculation of the Z-scores of the
three functions measured. The multifunctionality index for each plot was the average Z109
score for these three functions. The use of Z-scores allows measuring all functions on a
common scale, regardless of their units, has good statistical properties (normal
distribution, mean and variance not related), and is highly correlated with other
multifunctionality indices proposed in the literature (Maestre et al. in press).
Apart from calculating the number of species in each plot, to allow for
comparisons with other studies, we also calculated the phylodiversity of such plots. A
substantial number of indices have been proposed to evaluate different aspects of
phylodiversity (see Helmus et al. 2007; Vamosi et al. 2009; Pausas and Verdú 2010; for
recent reviews). Among them, the most used and accepted are the Mean Phylogenetic
Distance (MPD), the Mean Nearest Taxon Distance (MNTD), the Phylogenetic Species
Variability (PSV) and the Phylogenetic Species Evenness (PSE; see Webb et al. 2002;
Helmus et al. 2007; Kraft et al. 2007). In our case, all of them were strongly correlated
(r > 0.5 in all cases) and yielded similar results regarding the effect of A. altissima
invasion on them and their relationships with ecosystem multifunctionality. For
simplicity, we only show here the results regarding PSE. We selected this index because
it incorporates both phylogenetic and species abundance information, its variability is
less sensitive to sample size and does not depend on species richness (Helmus et al.
2007). The maximum value of PSE is 1, corresponding with a community formed by
evolutionarily independent species (the so-called “star phylogeny”) and with equal
abundances of those species (complete evenness). Since PSE can be more related to
evenness in the community than to its phylogenetic pattern (Helmus et al. 2007), we
also briefly comment on the results of the Phylogenetic Species Variability index (PSV)
to aid with the interpretation.
STATISTICAL ANALYSES
To assess for the differences in plant species composition and cover between invaded
and non invaded plots, ordination methods were carried out. Based on the length of the
main gradient in the data, estimated by an indirect gradient analysis (Detrended
Correspondences Analysis, DCA: ter Braak and Smilauer 2002) redundancy analysis
(RDA) was identified as appropriate method. Analyses of DCA and RDA were carried
out using CANOCO for Windows v 4.5 (ter Braak and Smilauer 2002). We included
species cover, and presence/absence of A. altissima, litter depth, cover of litter, cover of
bare soil and cover of rock as environmental variables. A forward selection procedure by
means of a Monte Carlo permutation test was carried out in order to test the significance of the
110
environmental variables. Significance of just the first and then all canonical axes were tested by
the distribution-free Monte Carlo test (999 permutations).
We analyzed the effect of A. altissima invasion on species richness,
phylodiversity (PSE), soil attributes, and ecosystem multifunctionality (M) separately.
For doing this, we used ANOVAs with A. altissima invasion as a fixed factor with two
levels (control and invaded). We ran separated models for each of the four ecosystem
attributes described above, which were introduced as response variables. We also
evaluated the relationship between M (dependent variable) and either species richness or
PSE (predictors) by using linear regressions. Apart from looking at the indirect effect of
A. altissima on multifunctionality (M), mediated by its effect on species richness or
phylodiversity (the ANOVA analysis described above), we aimed to assess for the
direct effect of A. altissima on such variable. For doing this, we used the residuals of the
linear regressions between either species richness or PSE and multifunctionality,
described above, as a response variable for an ANOVA with A. altissima invasion as a
fixed factor. The use of the residuals of the linear regressions between species richness
or PSE and M allowed us to account for the direct effect of A. altissima on M
independently of the effect of the invasive species on either the species richness or the
phylodiversity, respectively.
Although we justified above the rationale for the inclusion of each measured
variable as either structural or functional, we acknowledge that other of the measured
variables could be considered as functional rather than structural (e.g. phosphorus
availability or soil organic matter; Maestre et al. in press). For this reason, we
recalculated the M index using also phosphorus availability as a functional variable and
the results gathered were very similar (see results below). All the variables used
accomplished analyses assumptions and no transformations were needed. All the
analyses were carried out using SPSS 13.0 for Windows (Chicago, Illinois, USA),
excepting the Principal Component Analysis (PCA). The latter was conducted by using
Primer v. 6 statistical package for Windows (PRIMER-E Ltd., Plymouth Marine
Laboratory, UK). We calculated PSV and PSE indexes using the Picante package
(Kembell et al. 2010) for R version 2.10.1 (R Development Core Team 2009).
111
RESULTS
Species composition of understory and plant cover in riparian communities was affected
by invasion of A. altissima and litter accumulation (RDA axes I and II speciesenvironmental variables correlation: 0.98 and 0.77, respectively; cumulative percentage
variance 94.5 %) (Table 1).
Table 1 Results of the forward selection procedure on selected environmental variables using
Monte Carlo Permutation Test of the RDA analysis on species cover
Variable
λA
P
F
Presence/absence of A. altissima
0.65
33.25
0.001
% litter
0.08
5.13
0.004
litter depth
0.03
0.122
1.74
% stone
0.02
0.190
1.49
% bare soil
0.01
0.742
0.53
The invasion of A. altissima significantly reduced species richness (25 ± 2 vs 15
± 2 plant species in control vs invaded plots, respectively; Table 2) and phylodiversity
(PSE = 0.29 ± 0.02 vs 0.14 ± 0.02 in control vs invaded plots, respectively). The latter
results show an increase in phylogenetic clustering on the invaded communities. Since
results from the PSE index were almost identical to those of the PSV index (results not
shown), this suggests that the results found are driven by the effects of A. altissima on
the phylodiversity rather than on the evenness among understory species. These
negative effects of A. altissima on richness and phylodiversity also extended to
ecosystem multifunctionality (M index = 0.29 ± 0.08 vs -0.29 ± 0.15 in control vs
invaded plots, respectively) but not to soil attributes, which were unaffected by A.
altissima invasion (Table 2).
Both species richness and phylodiversity were significantly and positively
related to ecosystem multifunctionality (Fig. 1). In our case, phylodiversity was a better
indicator of multifunctionality than species richness (25% of the variance explained vs
17% explained by species richness). The latter results were robust regardless of the
phylodiversity indicator used (e.g. PSV vs M: R2 = 0.22; P = 0.038; MPD vs M: R2 =
0.24; P = 0.028). The ANOVAs conducted using A. altissima invasion as a fixed factor
and the residuals of the linear regressions between species richness or phylodiversity
and ecosystem multifunctionality revealed that the negative effect of A. altissima was
caused indirectly, mediated by its effect on phylodiversity and, to a less extent, species
112
richness (Table 2). When filtering the effect of species richness on M, the effect of A.
altissima on the latter was importantly reduced (a three-fold reduction in the F when
DIRECT EFFECT
Table 2 Summary of A. altissima invasion effects on the different ecosystem attributes
measured. The F-statistic and the P-value for A. altissima invasion effect, and also the amount
of variance (R2) explained by the model are shown. Since some of the variables used were
derived from the simplification of multiple data (i.e. soil attributes and multifunctionality; see
main text), an interpretation of the results has been added for clarity. “Direct” and “indirect
effect” labels mean that the raw data or the residuals from linear regressions were used in the
analyses, respectively. In the latter case, the predictor variable used in the linear regression
(with multifunctionality as a response) is shown between parentheses. Soil attributes = First axis
of the Principal Components Analysis performed with soil pH, EC, OM, available P and litter
cover. Phylodiversity = Results from the Phylogenetic Species Evenness index used.
Multifunctionality = M index constructed by averaging the Z-scores of the three functional
variables used (glucosidase and phosphatase enzymatic activities, plant biomass)
Ecosystem attribute
F1,18
R2
P-value
Interpretation
Soil attributes
0.85
0.05
0.369
No effect on soil pH, EC, OM
or available P
Species richness
12.7
0.41
0.002
A. altissima reduces species
richness and phylodiversity
Phylodiversity (PSE)
29.4
0.62
<0.0001
A. altissima reduces ecosystem
Multifunctionality
EFFECT
INDIRECT
Multifunctionality
(species richness)
Multifunctionality
(PSE)
11.3
0.39
0.004
3.23
0.15
0.089
multifunctionality. However,
this effect is mediated mainly
by its effect on phylodiversity
and, to a less extent, on
1.36
0.07
0.26
species richness
analyzing residuals vs raw data; Table 2) and turned from highly to marginally
significant. This change was more important when filtering by the effect of
phylodiversity on M. In the latter case, the effect of A. altissima on M turned to not
significant at all when analyzing the residuals and the F was reduced almost 10 times
(Table 2). When using the M index including P availability as a functional variable, the
results were very similar (effect of A. altissima on the raw index: F1,18 = 4.14; P = 0.057;
R2 = 0.2; relationship between M and species richness: R2 = 0.16; P = 0.077;
relationship between M and phylodiversity: R2 = 0.19; P = 0.054).
113
1,0
A
0,8
0,6
2
R = 0.17; P = 0.071
0,4
0,2
0,0
Multifunctionality index
-0,2
-0,4
-0,6
-0,8
-1,0
0
5
10
15
20
25
30
35
40
Species richness
1,0
B
0,8
0,6
0,4
2
R = 0.25; P = 0.026
0,2
0,0
-0,2
-0,4
Control
Invaded
-0,6
-0,8
-1,0
0,0
0,1
0,2
0,3
0,4
Phylodiversity (PSE)
Figure 1 Relationships between species richness (A) or phylodiversity (B) and ecosystem
multifunctionality. Results from the linear regressions are shown in each panel (n = 20). To aid
interpretation, the particular relationships for control and invaded plots are, for species richness
vs multifunctionality, R2 = 0.39; P = 0.071 and R2 = 0.09; P = ns, respectively. For the
relationship between phylodiversity and multifunctionality results were: R2 = 0.49; P = 0.025
for control and R2 = 0.06; P = ns for invaded plots.
114
DISCUSSION
The novelty of our study settles in that is one of the few existing ones addressing
the effect of invasive species on other measures of biodiversity than species richness,
and on multiple ecosystem functions simultaneously. To our knowledge, this is also the
first study separating direct and indirect (mediated by its effect on biodiversity) effects
of invasive plants on ecosystem functioning. Our main findings were that plant species
richness and phylodiversity, together with ecosystem multifunctionality, were reduced
under
the
presence
of
Ailanthus
altissima.
The
reduction
of
ecosystem
multifunctionality was indirectly mediated by the decline in phylodiversity and, to a less
extent, in species richness. This suggests that mainly with the removal of the invasive
species and the reintroduction of native ones, with no other actions needed, important
functions and services could be re-established. We speculate that a more rapid increase
of phylodiversity, and therefore on ecosystem multifunctionality, would be found if the
reintroduction of native species would include those taxa more distantly related with the
remaining ones. Among these functions, and at the light of previous research (Strauss et
al. 2006; Diez et al. 2008; Zavaleta et al. 2010) is likely that more phylodiverse
communities might prevent future invasions.
Our results support the first hypothesis since A. altissima reduced not only
species richness, but also the phylodiversity of neighbor vegetation. Moreover, we
found an alteration of species cover and composition in invaded plots. Although the
studies regarding the effect of invasive species on biodiversity show some controversy
(e.g. Sax and Gaines 2003; Meffin et al. 2010), our results are in the lines of those
developed in riparian habitats (e.g. Gerber et al. 2008; Gaertner et al. 2009; Davies
2011) or with A. altissima as their target species (Vilà et al. 2006; Motard et al. 2011).
A more detailed view of our results suggests an even more important reduction of A.
altissima on plant biodiversity than the one suggested by looking just at species richness
or phylodiversity. For example, we found other invasive species, such as Robinia
pseudoacacia L., in the plots invaded by A. altissima but not in the control ones. It is
known that the presence of invasive species accelerates invasions of other alien species
and amplify their effects on native communities (invasional meltdown; see Simberloff
and Von Holle 1999; Richardson et al. 2000; Pyšek and Richardson 2010). Moreover,
among the species only present in the control plots, we found rare species of high
conservation interest, such as Cephalanthera damasonium (Mill.) Druce (protected
115
species with one of the best populations in the area, Decreto 70/2009; Serra et al.,
2006). Other species of interest were found out, but near, of the sampled plots
(Himantoglossum hircinum (L.) Spreng., Populus canescens (Aiton) Sm., Potamogeton
coloratus Hornem. or Zannichellia peltata Bertol.); since they were near to plots
invaded by Ailanthus altissima, increasing the area invaded by A. altissima could
represent a potential threaten for these species too. These results highlight the
importance of preventing the invasion of exotic plants in habitats with conservation
interest, as the riparian forests studied here. Restoration measures including A. altissima
removal could favor the colonization of native species and help to preserve richness
(especially in case of endangered species).
Our study is one of the few existing ones that show how phylodiversity is a
better indicator of ecosystem functioning than other more commonly used measured of
biodiversity, such as species richness (Forest et al. 2007; Maherali and Klironomos
2007; Cadotte et al. 2008) and, to our knowledge, is the first one showing this
relationship using multiple ecosystem functions simultaneously. We found a reduction
of phylodiversity in invaded plots, which could have negative effects on multiple
ecosystem functions and services. Thus, our results highlight that management plans,
apart from removing A. altissima, should include strategies for the reintroduction of the
lost native species, especially of those taxa more distantly related to the remaining ones
and protected native species. On the other hand, the entrance of the invasive species
may be influenced by the phylogenetic structure of the host community (Vacher et al.
2010), with alien species more distantly related to the host community more likely to
end as invasive ones in such communities (Strauss et al. 2006; Diez et al. 2008).
Although it has not been studied before, joining the results of both lines of research
suggest that conserving more phylogenetically diverse species assemblages should
increase the resilience of such communities to future invasions. Communities assembled
by more functionally diverse species (i.e more phylodiverse) are more likely to 1)
sustain higher levels of ecosystem functions such as resilience and resistance to plant
invasions (Zavaleta et al. 2010) and 2) include among these species some taxa closely
related to possible invaders, therefore preventing them to establish and invade the host
communities (Strauss et al. 2006; Diez et al. 2008).
Our results also sustain the second hypothesis since the effects of A. altissima on
vegetation diversity indirectly reduced ecosystem multifunctionality. Although overall
our results clearly show a positive relationship between either species richness or
116
phylodiversity and ecosystem multifunctionality, we found that this relationship
changed when considering invaded or control plots separately (Fig. 1). There was a
negative relationship between multifunctionality and phylodiversity or richness when
we included only non invaded plots (blue dots in Fig. 1) and there was no relation when
including only invaded ones (red dots). The most plausible explanation for these results
that we might think of is that the high species richness or phylodiversity of the noninvaded plots could include functionally redundant species and a higher soil nutrient
uptake. It must be noted that most of the measured functions included in this study are
mostly related to nutrient cycling in soils and plant productivity, ignoring a number of
other important ecosystem functions. These particular functions may be favored by
some particular species, showing elevated growth rates or complementarity in the use of
resources (Cardinale et al. 2007), making the rest of species functionally redundant for
these particular functions but increasing more complex interspecific networks thus
enhancing coexistence and facilitating biodiversity maintenance (Bascompte et al.
2006). However, although less diverse assemblages may be able to increase the
performance of one particular function or process, more diverse communities are
necessary to sustain relatively high levels of multiple ecosystem functions
simultaneously (Zavaleta et al. 2010). The lack of relationship in the invaded plots
seems more likely driven by a high statistical noise in our data, probably derived from
different times since A. altissima invasion (Vilà and Ibañez 2011), or the presence of
other invasive species (Appendix 1). Regardless of the particular mechanisms behind
the differential biodiversity-functioning relationship between invaded and non-invaded
plots, we clearly show a reduction of ecosystem multifunctionality with the presence of
A. altissima. Although A. altissima has a rapid growth (Zasada and Little 2002), and
may increase the soil fertility and C fixation (Vilà et al. 2006; Gómez-Aparicio and
Canham 2008), our results show a net negative effect of A. altissima on the productivity
of herbs and shrubs and on C and P cycling.
Conversely, and contradicting our third hypothesis, we found that the effect of A.
altissima on multifunctionality was not direct, but indirectly mediated by its effect on
diversity. The reduction of plant diversity may cause important changes in plant
productivity and nutrient storage and cycling (Tilman et al. 1997; Hooper and Vitousek
1998; Cardinale et al. 2007), these were the functions covered with our measurements
and this may explain the overwhelming importance of these indirect effects in our
study. Previous experimental studies, however, has shown direct effects of A. altissima
117
on several ecosystem functions and soil attributes (litter decomposition and pH, Godoy
et al. 2010; N cycle, Castro-Díez et al. 2011); evidences also found with observational
studies (pH, organic C, C/N ratio, Vilà et al. 2006; Gómez-Aparicio and Canham 2008).
These contrasting results could be explained by the differences in the target habitat or
the soil functions between previous research and our study, or because we failed in
interpreting A. altissima as a cause rather than a consequence of the differences between
invaded and non-invaded plots (Levine et al. 2003). Although it is known that habitat
characteristics can influence on A. altissima invasion (e.g. Traveset et al. 2008;
Constán-Nava and Bonet under review), this latter explanation is unlikely. Firstly,
habitat characteristics were very similar among plots since we only worked in riparian
communities, we did not find differences in important soil attributes, and management
and disturbance levels of these plots has been exactly the same since the study area has
been a protected area for over 100 years. Secondly, A. altissima competiveness with
native species by means of allelopathic and herbicide substances (Heisey 1990, 1996;
Lawrence et al. 1991), or by its extended root system (Motard et al. 2011) may take
precedence over the initial habitat characteristics. Therefore, the difference in the degree
of invasion between control and invaded plots are more likely to be caused by dispersal
limitation than by initial differences among the plots. Invaded plots were closer than
control ones to roadsides, a well-known dispersal corridor for this species (Kowarik and
von der Lippe 2006; 2011), supporting this statement. All these evidences suggest that
the reduction found in invaded plots on species richness, plant cover, phylodiversity and
ecosystem multifunctionality are a consequence, and not a cause, of the presence of A.
altissima in such plots. Nevertheless, the contrasting results found with previous studies
call for future research including manipulative approaches and the measurement of
multiple ecosystem functions simultaneously to conclude if the invasive species alters
directly or indirectly ecosystem functioning.
CONCLUSIONS
Our study is the first one showing how phylodiversity and ecosystem multifunctionality
were reduced under the presence of an invasive species. Plots infested by Ailanthus
altissima showed lower phylogenetic diversity and species richness, which indirectly
decreased ecosystem multifunctionality. Our study highlights the importance of taking
into account both direct and indirect effects of invasive species for improving
118
management strategies in natural ecosystems. Moreover, we provide an easy
methodology to tease apart these direct and indirect effects by using observational data.
This may help to focus restoration and conservation efforts either on the elimination of
invasive species and introduction of native ones when indirect effects prevail, such as in
our case; or, alternatively, to include soil reclaiming or other restoration practices aimed
to recover the loss of ecosystem services and functions in case of existing both direct
and indirect effects of the alien species on ecosystem multifunctionality.
ACKNOWLEDGEMENTS
We thank to Carrascal de la Font Roja and Sierra de Mariola Natural Parks staff, MJ.
Nava, A. Constán, E. Pastor, A. Dávila and the rest of collaborators for their help in the
fieldwork. F.T. Maestre, Y. Valiñani, A. Sanz and P. Alonso helped with the enzymatic
assays. This research and SCN PhD fellowship were supported by the ESTRES Project
(063/SGTB/2007/7.1) and RECUVES Project (077/RN08/04.1) founded by the Spanish
Ministerio de Medio Ambiente. Font Roja Natura UA Scientific Station (ECFRN UA),
depending on the Pro-Vice-Chancellorship for Research, Development and Innovation
(VIDI) of the University of Alicante, supported also this research. RT was partially
supported by a postdoctoral scholarship from the Spanish Ministerio de Educación
(BVA 2010-0375).
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125
Appendix 1: list of species in control (non invaded) and invaded by A. altissima in
Mediterranean riparian forests
Control
Species
Life form*
Abundance§
Adiantum capillus-veneris L.
Agrimonia eupatoria L. subsp. eupatoria
Allium sphaerocephalon L.
Andryala integrifolia L.
Arctium minus (Hill) Bernh.
Aristolochia paucinervis Pomel
Asparagus acutifolius L.
Asperula aristata subsp. scabra (J. and C. Presl) Nyman
Bituminaria bituminosa (L.) C. H. Stirt.
Brachypodium phoenicoides Roem. and Schult.
Buglossoides arvensis (L.) I. M. Johnst.
Carex flacca Schreb. subsp. serrulata (Biv.) Greuter
Carex mairii Coss. and Germ.
Catananche caerulea L.
Celtis australis L.
Centaurea aspera L. subsp. aspera
Cephalanthera damasonium (Mill.) Druce
Cichorium intybus L.
Cirsium monspessulanum (L.) Hill. subsp. ferox (Coss.) Talavera
Cistus albidus L.
Convolvulus arvensis L.
Corylus avellana L.
Crataegus monogyna Jacq.
Daphne gnidium L.
Daucus carota L. subsp. carota
Dittrichia viscosa (L.) Greuter
Epilobium sp.
Equisetum ramosissimum Desf.
Euphorbia sp.
Festuca arundinacea Schreb. subsp. fenas (Lag.) Arcang.
Ficus carica L.
Foeniculum vulgare Mill. subsp. piperitum (Ucria) Cout.
Fraxinus ornus L.
Genista scorpius (L.) DC.
Hedera helix L.
Helichrysum italicum subsp. serotinum (Boiss.) P. Fourn.
Hypericum perforatum L.
Juglans regia L.
Juncus subnodulosus Schrank
Marrubium vulgare L.
Matthiola fruticosa (L.) Maire
Medicago minima (L.) L.
Medicago suffruticosa Ramond ex DC.
Mentha suaveolens subsp. suaveolens (L.) Hudson
Mercurialis tomentosa L.
Piptatherum miliaceum (L.) Coss. subsp. miliaceum
Origanum vulgare L. subsp. virens (Hoffmanns. and Link) Bonnier and Layens
G
H
G
T/H
H
G
C/N
C/H
C/H
H/G
T
G
H
H
M
H
G
H
H
N
G
M
Me
N
H
C
H
G
C
M
C
M
M
M
C
C
C
CC
C
C
M
M
E
CC
R
C
C
C
CC
R
C
C
CC
CC
H
Me
H
M
N
P
C
H
M
G
H
C
T
C/M
H
C
H
H
M
C
C
M
C
C
C
C
M
M
C
C
C
C
C
M
C
M
126
CC
Osyris alba L.
N
C
Pallenis spinosa (L.) Cass.
H
C
Pinus halepensis Mill.
M
CC
Plantago lanceolata L.
H
C
Polygala monspeliaca L.
T
M
Populus nigra L.
M
CC
Potentilla reptans L.
H
C
Prunus domestica L.
M
R
Prunus spinosa L.
N/Me
M
Pulicaria dysenterica (L.) Bernh.
H
C
Quercus coccifera L.
Me/N
CC
Quercus ilex L. subsp. ballota (Desf.) Samp.
M
C
Rhamnus alaternus L.
N/Me
CC
Rosa canina L.
P
R
Rosmarinus officinalis L.
N
CC
Rubia peregrina L. subsp. peregrina
P
C
Rubus ulmifolius Schott
P
CC
Salix atrocinerea Brot.
Me
M
Salvia verbenaca L.
H
M
Sambucus nigra L.
Me
M
Scabiosa atropurpurea L.
H
M
Scirpus holoschoenus L.
G
CC
Scrophularia auriculata L. subsp. valentina (Rouy) Ortega Oliv., Serra, Herrero
H
C
& Muñoz Garm.
Silybum marianum (L.) Gaertn.
H
M
Smilax aspera L.
P
C
Solanum dulcamara L.
P
M
Sonchus maritimus L. subsp. aquatilis (Pourr.) Nyman
H
C
Torilis sp.
T
Trachelium caeruleum L.
H
M
Trifolium repens L.
H
C
Trifolium sp.
Ulex parviflorus Pourr.
N
C
Ulmus minor Mill.
M
C
Verbascum sp.
H
Viburnum tinus L.
Me
M
Vicia sp.
Viola alba Besser
H
M
*C: chamaephyte, P: phanerophyte, G: geophyte, H: hemicryptophyte, M: macrophanerophyte,
Me: mesophanerophyte, N: nanophanerohyte, T: therophyte
§ CC: very common, C: common, E: exotic, M: medium abundance, R: rare, RR: very rare
127
Invaded by A. altissima
Species
Life form*
Abundance§
Agrostis stolonifera L.
Ailanthus altissima (Mill.) Swingle
Allium sphaerocephalon L.
Asparagus acutifolius L.
Asperula aristata subsp. scabra (J. and C. Presl) Nyman
Avena sp.
Bituminaria bituminosa (L.) C. H. Stirt.
Brachypodium phoenicoides Roem. and Schult.
Brachypodium retusum (Pers.) Beauv.
Bromus diandrus Roth
Bromus sp.
Bryonia dioica Jacq.
Buglossoides arvensis (L.) I. M. Johnst.
Carex flacca Schreb.
Celtis australis L.
Cichorium intybus L.
Convolvulus arvensis L.
Conyza sumatrensis (Retz.) E. Walker
Crataegus monogyna Jacq.
Cynoglossum cheirifolium L.
Daphne gnidium L.
Daucus carota L. subsp. carota
Equisetum ramosissimum Desf.
Eryngium campestre L.
Ficus carica L.
Foeniculum vulgare Mill. subsp. piperitum (Ucria) Cout.
Galium aparine L.
Galium spurium L.
Geranium rotundifolium L.
Hedera helix L.
Humulus lupulus L.
Juncus subnodulosus Schrank
Lonicera etrusca G. Santi
Marrubium vulgare L.
Medicago sativa L.
Mentha sp.
Mentha suaveolens subsp. suaveolens (L.) Hudson
Oxalis corniculata L.
Pallenis spinosa (L.) Cass.
Parietaria judaica L.
Piptatherum miliaceum (L.) Coss. subsp. miliaceum
Plantago lanceolata L.
Polygala monspeliaca L.
Populus nigra L.
Prunus spinosa L.
Quercus coccifera L.
Quercus ilex L. subsp. ballota (Desf.) Samp.
Rhamnus alaternus L.
Robinia pseudoacacia L.
Rosa canina L.
Rubia peregrina L. subsp. peregrina
Rubus ulmifolius Schott
G
M/Me
G
C/N
C/H
C
E
C
C
C
C/H
H/G
H/G
T
C
CC
CC
C
H
T
G
M
H
G
T
Me
H
N
H
G
H
Me
H
T
T
T
P
P
G
P
H
H
M
C
C
E
C
CC
M
C
C
C
CC
CC
CC
C
C
C
M
CC
C
RR
M
M
C
CC
H
H
H
H
H
H
T
M
N/Me
Me/N
M
N/Me
M
P
P
P
C
C
C
CC
C
C
M
CC
M
CC
C
CC
E
R
C
CC
128
Sambucus nigra L.
Me
M
Silybum marianum (L.) Gaertn.
H
M
Smilax aspera L.
P
C
Solanum dulcamara L.
P
M
Torilis arvensis (Huds.) Link subsp. neglecta (Spreng.) Thell.
T
C
Torilis sp.
T
Trifolium sp.
Ulmus minor Mill.
M
C
Urtica urens L.
T
CC
Viburnum tinus L.
Me
M
Vicia sp.
*C: chamaephyte, P: phanerophyte, G: geophyte, H: hemicryptophyte, M: macrophanerophyte,
Me: mesophanerophyte, N: nanophanerohyte, T: therophyte
§CC: very common, C: common, E: exotic, M: medium abundance, R: rare, RR: very rare
129
Appendix 2: Evolutionary relationships between the sampled species
130
APÉNDICE FOTOGRÁFICO
Foto
1
Bosque
de
ribera
mediterráneo presente en el área de
estudio
Foto 2 Presencia de A. altissima en
bosque de ribera dentro del área de
estudio
131
132
CAPITULO 4
Long-term
Ailanthus
control
of
the
altissima:
invasive
Insights
tree
from
Mediterranean protected forests4
4
Manuscrito publicado en: Forest Ecology and Management 206 (6), pp 1058-1064
Autores: Soraya Constán-Nava, Andreu Bonet, Estrella Pastor, Maria José Lledó
133
134
ABSTRACT
Ailanthus altissima is an invasive tree species which has colonized numerous
ecosystems and affected ecosystem processes worldwide. Despite its importance as an
invasive species and the high economic costs incurred from its spread, there is a lack of
long-term management planning for its control. Although mechanical disturbance is
commonly applied, the effect that this treatment may have exhausting its resprouting
ability and also its joint effect with phytochemical treatments are poorly understood,
especially in Mediterranean environments. We tested three treatments (plus a control)
aimed to reduce A. altissima growth in Mediterranean forests throughout 5 years of
study. The treatments (one cut stump, double cut stump and cut stump with glyphosate
application) were repeated annually. General plant performance (biomass, height and
resprout-type density) was measured yearly during the study. Water potential and
stomatal conductance were also measured at the end of the study to evaluate particular
ecophysiological factors which might affect the response of A. altissima to assayed
treatments, together with leaf area index. Our results show that only the cut stump with
glyphosate application treatment was able to reduce the long-term growth and spread of
A. altissima. The treatments applied favoured collar sprout growth in response to
disturbance events (treatments) opposite to the control, where new sprouts grew mainly
from the root. Treated resprouts displayed ecophysiological changes depending on the
assayed treatment. To our knowledge, this is the only study testing the long-term effect
of both physical disturbance and phytochemical application on A. altissima growth. Our
study further refines our knowledge on the effects of repeating both commonly and
newly used treatments, improving our management techniques to reduce the presence
and growth of this invasive tree.
135
136
INTRODUCTION
A
ilanthus altissima (Mill.) Swingle (Simaroubaceae) is a deciduous tree native to
China and North of Vietnam that has developed into an invasive species
expanding on all continents except Antarctica (Kowarik and Säumel 2007). It is a
significant invasive plant in the Mediterranean Basin, where it mainly occurs on
disturbed urban sites, oldfields, and along roadsides (Kowarik and Säumel 2007). This
expansion has been caused by its use in roadside restoration and as an ornamental
species (Kowarik and Säumel 2007). A. altissima exhibits rapid establishment and
growth, with a high rate of sexual reproduction (Little 1974; Bory and ClairMaczulaijtys 1980) and the ability to resprout rapidly and form dense clonal stands after
disturbance (Kowarik 1995; Kowarik and Säumel 2007). A. altissima has strong
competitive effects on other species due to the presence of allelopathic compounds
(Heisey 1990, 1996; Heisey and Heisey 2003; Carter and Fredericksen 2007), which
affects ecosystem functioning and vegetation dynamics and structure (Lawrence et al.
1991; Vilà et al. 2006). Its control is therefore desired, particularly in protected areas
(Castroviejo et al. 2003; Meggaro and Vilà 2002).
The main method commonly used to eliminate A. altissima is mechanical
removal by means of the complete removal of aerial biomass once a year (hereafter onecut stump treatment; Hoshovsky 1988; Hunter 2000). This is problematic, however, due
to the plant’s tendency to resprout following disturbance (Hoshovsky 1988; Bory et al.
1991). The tendency to resprout is a highly efficient strategy in response to the loss of
above-ground biomass following disturbance (Midgley 1996; Bellingham and Sparrow
2000; Bond and Midgley 2001). While one cut stump treatment once a year may be an
unsuccessful control method, continued long–term stump cutting, especially after this
species spends some of its reserves during its annual growth pulse, could conceivably
lead to the depletion of the reserves in this resprouting species and therefore
compromise the chances of regeneration (Vilà and Terradas 1995; Canadell and LópezSoria 1998). Although never tested, mechanical treatments applied twice per year
(hereafter double cut stump treatment) during the growing period could therefore
improve the success of control methods, particularly in the long–term success.
Alternatively, recent research in temperate environments has identified the fact that the
joint application of mechanical and chemical (herbicide) treatments (hereafter cut stump
with herbicide application treatment) is more effective at eradicating this species than
137
mechanical treatment alone (Meloche and Murphy 2006). These joint methods have not,
however, been tested under Mediterranean conditions. The mid– and long–term effects
of repeated application of these treatments on this resprouting species are virtually
unknown, and therefore studies to assess the long–term prognosis for control are needed
in order to test the usefulness of these techniques.
The invasion of A. altissima into Mediterranean ecosystems has occurred despite
the fact that it has not evolved within a climate of persistent summer droughts and
unpredictable soil water, characteristics of this Mediterranean climate. Theoretically,
therefore, A. altissima should be unable to outcompete native woody species, which
have evolved structural and physiological mechanisms to cope with these environmental
constraints (Levitt 1980). However, the water-saving mechanism (i.e., simultaneous
leaves water loss and root hydraulic conductance reduction) found on seedlings of A.
altissima has been suggested to be related to its wide expansion in Mediterranean areas
(Trifilò et al. 2004). Control treatments applied to reduce this species presence may
affect resource uptake by the new resprouts after perturbation to overcompensate
biomass loss. This would lead to soil water depletion and a derived increase in
competition with native vegetation. Alternatively, control treatments could negatively
affect the physiology of the invasive species resprouts (Meloche and Murphy 2006),
causing changes in the response of A. altissima to drought due to stomatal alterations.
Also potential positive feedback effects between the direct (biomass loss, poisonous
effect) and indirect (reduced water use efficiency, loss of competitive ability against
native plants) negative effects on A. altissima performance, could reduce the
competitive ability of this species against native vegetation because of a reduction of its
performance and occupied area (Huston 1999) and therefore lead to more successful
control of this species in Mediterranean environments. Because of these possible
complex and at times counter-intuitive effects, it is important to increase our knowledge
about the traits that allow A. altissima to invade Mediterranean environments -in
particular its water use strategy- and how control treatments affect these adaptations and
therefore its competitive ability against native species. This is critical if we are to
understand its potential effects on Mediterranean ecosystems, how it competes with
native woody plants, and how we can work towards preventing it from becoming a
dominant woody invader.
In this paper we aimed to determine the best strategy to reduce the long–term
invasion of A. altissima at Mediterranean forests. We tested the effects of three factors
138
combining mechanical and chemical treatments (one cut stump per year, double cut
stump per year, and one cut stump with herbicide application per year) on the growth of
Ailanthus altissima. The double cut stump treatment was considered an easier
alternative to herbicide application as the study area is as protected area used for
conservation. Our main hypotheses were: i) The one cut stump treatment per year will
be the least efficient treatment, because of an increase of A. altissima resprout growth
and survival during the following years (Bory et al. 1991; Meloche and Murphy 2006).
This increase could implicate ecophysiological changes in the species which allow it to
grow and maintain itself in the plant community, increasing its competitive ability
against native species (e.g., high stomatal conductance), ii) The double cut stump
treatment per year and one cut stump with herbicide application treatment per year will
be the most efficient methods, because they will affect negatively to the invasive
species. The first will reduce aboveground growth of resprouts as a result of a reduction
of root reserves as demonstrated in native resprouting species (Vilà and Terradas 1995;
Canadell and López-Soria 1998). The second will reduce plant performance and will
produce morphological alterations caused by the herbicide. The reduction of A.
altissima resprouts in both treatments will be jointed to ecophysiological changes which
will affect negatively to the species. This will reduce its competitive ability with native
species.
MATERIAL AND METHODS
STUDY SITE
The study was carried out in Carrascal de la Font Roja Natural Park in the northwest of
Alicante Province (SE Spain). Elevations range from 600 to 1356 m above sea level.
Soils are limestone with the presence of impermeable clays. The climate is
Mediterranean and is characterised by cold and wet winters, and a marked summer
drought. Mean monthly temperatures range from 4.4 ºC in January to 24.9 ºC in July.
During the five year study period, total annual rainfall ranged from 199.3 mm to 636
mm. The Natural Park includes different ecosystems such as deciduous forests (with
Quercus faginea Lam., Fraxinus ornus L., Acer granatense Boiss.), holm oak forests
(Quercus ilex subsp. ballota (Desf.) Samp.), Aleppo pine forests (Pinus halepensis
Mill.) and scrublands (with Genista scorpius (L.) DC., Ulex parviflorus Pourr., Cistus
albidus L., Cistus clusii Dunal, Rosmarinus officinalis L., Quercus coccifera L.,
139
Daphne gnidium L). Ailanthus altissima was first introduced at the Natural Park as an
ornamental garden plant and as a stabilising species used in motorway slope restoration
(Constán-Nava et al. 2007). Later A. altissima began to invade different vegetation
areas, infesting mainly pine forests and scrubland, but its presence is also important in
areas of conservation interest, such as oak and river forests (Constán-Nava et al. 2007).
EXPERIMENTAL DESIGN
In summer of 2005, we randomly selected 12 populations of Ailanthus altissima (see a
detailed description in Table 1), which were located along the roads crossing the natural
park in the north-facing slope. The distance between nearest populations ranged from
0.2 km to 2.8 km. All populations came from resprouts because of the repeated
application of annual stump cutting on them during more than 10 years. Biomass
removal stopped two years before the study.
Three treatments (plus a control, i.e., no herbicide or mechanical treatment) were
applied in a completely randomised design (n = 3) across the 12 populations, selecting 3
populations for each treatment. The different treatments applied were: one cut stump,
double cut stump, and one cut stump with herbicide (glyphosate) application (see
below). Each treatment was repeated annually on the overall shoots composing each
population, during the 2005-2008 period. In each of these populations, we randomly
selected three 2 m × 2 m subplots to conduct the monitoring explained below. Selection
of subplots in each treated (and control) populations was done avoiding border effect
and other soil heterogeneity factors (e.g. stones) and to ensure that all resprouts in the
treated area derived from roots that have no green shoot to supply carbohydrates
immediately after cut.
One cut stump treatment (hereafter 1CT) consisted of cutting all the resprouts in
July. Double cut stump treatment (hereafter 2CT) included the one cut treatment plus a
second cut stump on September, when A. altissima resprouts after the first cut stump.
Cut stump with glyphosate application treatment (hereafter CHT) was based on
applying the 1CT treatment combined with an immediate application of glyphosate on
the cut section of the resprout. We used a paintbrush to prevent any impact on
surrounding vegetation. Glyphosate was used because of its effectiveness on cut stumps
and because its use is allowed in the Natural Park under their management plans. The
glyphosate treatment was applied late in the growing season (July) when the root system
is most affected by the herbicide (Hoshovsky 1988).
140
Table 1 Characteristics of A. altissima populations assigned to the different treatments of the
experiment. Ai: invaded area; RCD: root collar diameter (mean ± ES)
Treatment Population Ai (m2)
Control
1CT
2CT
CHT
1
2
3
1
2
3
1
2
3
1
2
3
286.1
147
927
45.8
59.3
92.4
55.9
910.6
858.8
62.5
910.6
2058.3
Total number
RCD (mm)
of shoots
951
11.7 ± 0.2
314
15.2 ± 0.5
1043
7.7 ± 0.2
86
13.7 ± 1.5
181
16.8 ± 1.0
214
9.1 ± 0.6
130
24.1 ± 1.7
275
9.8 ± 0.3
2412
6.3 ± 0.1
72
11.7 ± 0.9
275
11.5 ± 0.6
1066
9.5 ± 0.2
Vegetation
Oak forest
Pine forest
Pine forest
Oak forest
Pine forest
Pine forest
Pine forest
Pine forest
Pine forest
Pine forest
Pine forest
Pine forest
Altitude
(m a.s.l.)
1050
710
750
1040
900
710
950
800
910
920
800
740
PLANT PERFORMANCE SURVEY
During the five years of the study, we measured height and root collar diameter
(hereafter RCD) of all resprouts in the subplots. We used allometric relationships
among biomass, height and RCD of resprouts growing under natural conditions (n =
100) to estimate biomass on our plots. The RCD was strongly related to dry weight
(biomass = 0.0508 × RCD2.8427, R2 = 0.96, P < 0.001), so was used as a field–based
surrogate of plant biomass. The density of resprouts growing on the 2 m × 2 m subplots
was measured each year following application of the necessary treatments. As
disturbance events affect the origin of the new sprouts (Del Tredici 2001), three
resprout types were defined: stem (adventitious resprout on a cut section the previous
year), collar resprout (growing in a lateral of a previous year cut section) and root
resprout (growing in a distance ≥ 5 cm from a cut section the previous year)
(Hoshovsky 1988; Bory et al. 1991), in order to test the effects of applied treatments on
each resprout type.
In the last year of the study (2009), we measured leaf area index (LAI) in each
subplot in summer using a LAI 2000 plant canopy analyzer (Li-Cor, Inc., Lincoln, NE).
Leaf water potential was also determined on freshly cut leaves of five resprouts of each
treatment (plus control) at pre-dawn (Ψpd) and midday (Ψmd) of spring (May) and
summer (July) of 2009 using a Schölander chamber (Soil Moisture Equipment Corp.,
Santa Barbara, California, USA). Stomatal conductance to water vapour (gs) was carried
out on five different resprouts (five leaves per resprout) in the morning and midday of
spring (May) and summer (July) with a portable infrared gas analyser (LI-6400, LI141
COR Inc., Lincoln, Nebraska, USA). gs measurements were made in full sunlight,
recording data after a period of stabilization when the coefficient of variation was < 5%.
STATISTICAL ANALYSES
Effect of treatments on A. altissima biomass, height, RCD, density of different resprout
types and total resprout density, Ψ and gs were analyzed by repeated measures analysis
of variance (RM ANOVA), using as input the mean value from the three subplots of
each population (n = 3). In all RM ANOVA models performed, Greenhouse-Geisser
correction was used when data did not satisfy sphericity assumptions. In these models
“treatment” was consider as the between-subjects factor and “time” (sampling date) as
the within-subjects factor. In Ψ and gs models, “hour of the day” was considered as the
within-subjects factor. Interactions between time and treatment were found in all
variables (see below). As these interactions may lead to misinterpretations in the effect
of applied treatments (Quinn and Keough 2002), separated analysis of variance
(ANOVA) were performed for each sampling date in order to disentangle the effect of
each treatment upon these variables at each sampling date. Tukey´s HSD post-hoc test
was used to test for significant differences among treatments. Data were log10–
transformed, when necessary, to satisfy ANOVA assumptions. Analyses of variance
were performed to assess for differences on biomass, height, and total resprout density
between populations prior to the application of the treatments (2005). In all cases, no
differences were found between populations prior to the application of treatments (P >
0.05). Differences in LAI between treatments were tested with the non-parametric
Kruskal-Wallis test. All statistical analyses were performed using SPSS v.15 (SPSS
Inc., Chicago, IL, USA).
RESULTS
PLANT GROWTH SURVEY
The combined herbicide plus mechanical treatment (CHT) reduced biomass, RCD and
height by about 90% one year after application (2006) and over the following three
years (Table 2). No other treatments differed from the control. The 2CT treatment
reduced plant height by more than 50% from the second year (2007) onwards and the
1CT treatment reduced plant height by 60%, but only during the last two years.
142
The density of shoots of A. altissima did not differ significantly between
treatments, despite the 1CT and 2CT tended to increase gradually over time compared
with the control and CHT tended to reduce the number of shoots (Table 2).
Table 2 Biomass, RCD, height and density of A. altissima (mean ± ES, n = 3) for each
treatment and year of study (ANOVA and Tukey´s test, P < 0.05). Treatments are: Control;
1CT: one cut stump treatment; 2CT: double cut stump treatment; CHT: cut stump with
glyphosate application treatment. Different letters indicate significant differences between
treatments
Treatment
2005
Biomass (gr m-2 )
Control
1318 ±603 a
1CT
1879 ±1205 a
2CT
2262 ±1825 a
CHT
828 ±246 a
RCD (mm)
Control
11.7 ±2.2 a
1CT
14.8 ±4.0 a
2CT
14.5 ±3.6 a
CHT
12.8 ±1.1 a
Height (m)
Control
1.2 ±0.3 a
1CT
1.3 ±0.3 a
2CT
1.4 ±0.4 a
CHT
1.0 ±0.2 a
Density (num m-2)
Control
8 ±2 a
1CT
4 ±0 a
2CT
8 ±3 a
CHT
5 ±1 a
2006
2007
2008
2009
2422 ±1042 a
283 ±85 ab
341 ±144 ab
36 ±25 b
3330 ±1511 a
802 ±463 ab
414 ±163 ab
115 ±42 b
3129 ±1366 a
235 ±81 ab
534 ±55 a
45 ±30 b
3560 ±1656 a
160 ±69 a
206 ±101 a
15 ±15 b
14.3 ±3.2 a
7.4 ±1.3 ab
7.7 ±0.5 ab
4.4 ±0.9 b
15.5 ±3.4 a
7.6 ±1.8 ab
7.1 ±0.2 ab
4.5 ±0.3 b
15.7 ±3.1 a
5.7 ±1.6 bc
6.9 ±0.7 ab
2.4 ±0.7 c
17.3 ±4.3 a
6.2 ±2.1 a
6.0 ±0.3 a
1.4 ±0.5 b
1.3 ±0.4 a
0.6 ±0.1 ab
0.6 ±0.2 ab
0.2 ±0.0 b
1.5 ±0.3 a
0.7 ±0.2 ab
0.6 ±0.0 b
0.3 ±0.0 b
1.4 ±0.3 a
0.5 ±0.1 b
0.5 ±0.0 b
0.1 ±0.0 b
1.6 ±0.5 a
0.5 ±0.1 b
0.4 ±0.0 b
0.07 ±0.0 b
8 ±2 a
8 ±1 a
10 ±3 a
3 ±1 a
6 ±1 a
11 ±2 a
13 ±5 a
6 ±3 a
7 ±1 a
14 ±4 a
16 ±6 a
5 ±2 a
7 ±1 a
12 ±6 a
11 ±4 a
2 ±1 a
Temporal changes differed in relation to treatment on stem resprouts type
density (Time × Treatment interaction: F9,24 = 2.82, P = 0.02; Table 3). 1CT and 2CT
had significantly more stem resprouts than control and CHT which were reduced over
time. There were more collar resprouts on the 1CT and 2CT treatments than the control
(F3,8 = 6.41, P = 0.01, Table 3), but the combined (CHT) treatment did not differ from
any of the treatments. The density of collar resprouts also changed significantly over
time (F3,24 = 7.04, P < 0.001, Table 3), but not time × treatment interaction was found.
There were no significant differences in density of root resprouts over time between
treatments. The LAI of trees in the combined (CHT) treatment was significantly less
than that of trees in the other treatments (Fig. 1).
143
Table 3 Resprout density of different types (mean ± ES. n = 3) by treatment from the period
study (ANOVA and post hoc Tukey´s test, P < 0.05). Treatments are: Control; 1CT: one cut
stump treatment; 2CT: double cut stump treatment; CHT: cut stump with glyphosate application
treatment. Different letters indicate significant differences between treatments
Resprout type density
(num m-2)
Stem
Collar
Root
Treatment
2006
2007
2008
2009
Control
1CT
2CT
CHT
Control
1CT
2CT
CHT
Control
1CT
2CT
CHT
0 ±0 a
3 ±1 b
2 ±1 b
0 ±0 a
0 ±0 a
4 ±1 b
5 ±1 b
2 ±1 ab
0 ±0 a
1 ±1 a
3 ±2 a
1 ±1 a
0 ±0 a
2 ±1 ab
3 ±1 b
0.4 ±0 ab
0 ±0 a
6 ±1 ab
6 ±3 b
2 ±1 ab
0 ±0 a
3 ±1 a
4 ±2 a
3 ±3 a
0 ±0 a
0 ±0 a
1 ±1 a
0 ±0 a
0 ±0 a
10 ±2 ab
11 ±5 b
5 ±2 ab
1 ±0 a
3 ±2 a
4 ±1 a
0 ±0 a
0 ±0 a
0.2 ±0 b
0 ±0 a
0 ±0 a
0 ±0 a
7 ±1 b
7 ±2 b
2 ±1 ab
0 ±0 a
1 ±1 a
4 ±2 a
0 ±0 a
6
a
Leaf area index
5
4
a
3
2
a
1
b
0
Control
1CT
2CT
CHT
Treatment
Figure 1 Leaf area index (LAI; mean ± SE. n = 3) of A. altissima. Treatments are: Control;
1CT: one cut stump treatment; 2CT: double cut stump treatment; CHT: cut stump with
glyphosate application treatment. Different letters indicate significant differences (KruscalWallis test)
ECOPHYSIOLOGICAL MEASUREMENTS
Ailanthus altissima showed marked seasonal changes, both in the morning and at
midday and relative to the treatments in Ψ (RM ANOVA: Time, F1,16 = 37.34, P <
0.001, Moment of the day, F1,16 = 345.47, P < 0.001, Treatment, F3,16 = 9.63, P =
0.001). Also, A. altissima showed changes on gs (RM ANOVA: Time, F1,14 = 167.64, P
144
< 0.001, Moment of the day, F3,42 = 94.97, P < 0.001, Treatment, F3,14 = 21.12, P <
0.001). In May, Ψpd of resprouts under the 1CT and CHT treatments were lower than
the 2CT, but none was significantly different from the control (Table 4). gs in the
morning was higher for CHT resprouts than the other treatments (F3,16 = 81.62, P <
0.001, Table 4). At midday, Ψ of the control resprouts was lower than the other
treatments which had all declined substantially. All treated resprouts showed higher gs
at midday than control. In July, Ψ
pd
declined in the 1CT resprouts, but the opposite
response was found on the other treatments (Table 4). There were no differences in Ψmd
between treatments. 1CT and 2CT showed lower gs in the morning than control and
CHT (Table 4). All treatments substantially reduced gs at midday, excepting CHT
resprouts.
Table 4 Leaf water potential (Ψ) and leaf conductance to water vapour (gs) of each treatment
measured in the morning and midday on May and July (mean ± ES). Different letters indicate
significant differences between treatments (ANOVA and post hoc Tukey´s test. n = 5).
Treatments are: Control; 1CT: one cut stump treatment; 2CT: double cut stump treatment; CHT:
cut stump with glyphosate application treatment
Ψ (-MPa)
Control
1CT
2CT
CHT
gs (mmol m-2 s-1)
Control
1CT
2CT
CHT
May
Morning
Midday
July
Morning
Midday
0.5 ± 0.03 ab
0.5 ± 0.02 b
0.4 ± 0.04 a
0.6 ± 0.01 b
2.1 ± 0.2 a
1.4 ± 0.1 b
1.2 ± 0.2 b
1.6 ± 0.1 b
0.6 ± 0.04 a
1.2 ± 0.2 b
0.8 ± 0.1 a
0.6 ± 0.02 a
2.3 ± 0.1 a
2.1 ± 0.2 a
2.0 ± 0.2 a
1.8 ± 0.1 a
155.1 ± 18.3 a
72.3 ± 5.6 b
169.4 ± 10.3 a
417.1 ± 25.3 c
45.8 ± 7.9 a
222.8 ± 31.4 c
120.0 ± 6.6 b
194.7 ± 17.6 c
122.6 ± 16.02 a
55.5 ± 13.5 b
57.7 ± 5.8 b
152.9 ± 22.7 a
5.9 ± 1.0 a
33.8 ± 6.7 ab
4.7 ± 0.2 a
71.7 ± 20.3 b
DISCUSSION
Long–term control of A. altissima resprouting was efficient when using the combined
herbicide and cutting treatment (CHT), mainly because of its reduction on aboveground
growth. Treatments only based on cut stump did not reduce the invasive species
presence significantly. Despite root removal is a commonly used technique to reduce
growth of resprouting trees, we did not consider this technique in our study because of
its high cost and predictable impact on soil surface, besides the use of heavy machinery
is limited by the natural park management plans and therefore this methodology is not
145
applicable in the study area. Overall, our results indicate that CHT repeated annually
reduced biomass of A. altissima performance at long-term (this treatment reduced the
invasive tree biomass, leaf index area and also affected its ecophysiological traits).
These effects could reduce its competitive ability with native species as suggested by
the fact that natural recolonization by native species was observed after five years of
treatment application.
EFFECTS OF THE TREATMENTS ON ABOVEGROUND GROWTH
In our study, neither single nor double cut stump treatments (1CT, 2CT) significantly
reduced resprout density, as evident in a number of other studies developed at temperate
areas (Bory et al. 1991; Burch and Zedaker 2003; Meloche and Murphy 2006). These
treatments did not reduce biomass, though resprout height was reduced after three years
of repeated annual cutting. A trade-off between stem height and stem number has been
described in sprouting species following disturbance (Midgley 1996; Kruger et al. 1997;
Vesk et al. 2004), consistent with the response we observed in A. altissima. Thus the
probability of survival of this invasive species is greater given its tendency to
recovering lost biomass through resprouting (Malanson and Trabaud 1988; Bellingham
and Sparrow 2000; Bond and Midgley 2001). Rainfall could have been responsible for
the species tendency to resprout in this study (Riba 1998), particularly the high spring
rainfall during 2007-2008 in the study site.
In our study, the combined herbicide and cutting treatment (CHT) failed to
reduce resprout density, in contrast with previous studies (Meloche and Murphy 2006).
However, although results were non-significant, this treatment finally removed
resprouts from more than half of the subplots by repeated application of the CHT
treatment. This reduced plant height and biomass, which suggests that more prolonged,
persistent application of herbicide and stem cutting may result in total control, though
further studies are needed to test this formally.
The resprouting response of A. altissima could be influenced by topographic
conditions, such as aspect, which could influence both water and nutrient availability
(López-Soria and Castell 1992; Gracia 2000). The north-facing slope on which our
subplots were located could explain resprouting in the CHT treatments in spite of
marked reductions in biomass. Only the CHT treatment reduced LAI, through reduced
biomass and resprout height. The reduction in A. altissima growth under this treatment
could promote the persistence and colonization of native species over invasive species.
146
For example, we found natural recovery of native species such as Thymus vulgaris L.
subsp. vulgaris, Brachipodium retusum (Pers.) P. Beauv., Cistus albidus L., or seedlings
of Quercus ilex L., Viburnum tinus L.and Pinus halepensis Mill in plots undergoing this
treatment but not the rest of assayed techniques(S. Constán-Nava, pers. observ.). This
could indicate a reduction in competition for light and the production of allelopathic
compounds through lower plant biomass and LAI.
All three treatments altered plant resource allocation. The majority of new
resprouts in the control treatment were root resprouts, which promote the colonization
of new ground. This is a common behaviour in sprouting species, and could involve the
development of adventitious roots that become autonomous from the main parent plant
(Del Tredici 2001). However, all disturbance treatments favoured the growth of collar
resprouts, which is a common response to frequent disturbance (Del Tredici 2001). The
sprouting collar is an active specialized organ of regeneration and rejuvenation,
resulting in the proliferation of a large numbers of sprouts (Bory et al. 1991; Del Tredici
2001). In contrast, the low number of stem resprouts growth could be due to an
inefficient mechanism for resprouting species, comparing with collar type which comes
from active organ.
EFFECTS ON ECOPHYSIOLOGICAL TRAITS
Rehydration and stomatal conductance of control plants were higher in spring than
summer which indicates higher water availability in spring. Adaptations consisting on
the progressive stomatal closure in response to increasing summer drought are known
from seedlings of this species (Trifilò et al. 2004), and may be a mechanism for
improving water use efficiency under Mediterranean climate. Our results corroborate
that A. altissima present high plasticity to Mediterranean drought by means of a watersaving strategy (Levitt 1980). This strategy allows growing and developing the invasive
species during drought periods (Levitt 1980). The high leaf water potential at pre-dawn
found on A. altissima plants, mainly during summer drought, could indicate a better
rehydration of the invasive species than native ones. For example, Fraxinus ornus, a
native deciduous forests tree, showed more negative leaf water potential values (-2.56 ±
0.56 MPa) than A. altissima (-0.6 ± 0.04 MPa) under the same environmental conditions
(S. Constán-Nava, unpublished data). This high rehydration of A. altissima could allow
for better growth and improve its competitive abilities against native species under
drought.
147
Substantial ecophysiological changes were detected on treated resprouts. All
treated resprouts showed high rates of stomatal conductance during spring, especially at
midday. High stomatal conductance could help the invasive species to recover and grow
after disturbance events. This could explain the lack of biomass reduction under 1CT
and 2CT treatments. The regulation of stomatal conductance is not only affected by
changes in environmental factors, but also endogenous effects are important (Schulze
1986; Meinzer et al. 2001). High stomatal conductance were registered on CHT
resprouts, maybe caused by internal effects derived from negative effects on leaf
morphology produced by the herbicide (S. Constán-Nava, pers. observ.). The stomatal
conductance of resprouts of all assayed treatments were higher than in control plants
due to a better hydration (with Ψmd > -1.6 MPa). A plausible explanation for this result
is the difference on size. That is, control plants showed hydration related with its tree
structure, in contrast with the young resprout structure of treated plants. In summer
midday, all treated resprouts tended to close the stomata, such as control plants,
excepting CHT resprouts. The latter had high stomatal conductance despite being at
summer drought. This could damage its water status and reduce its overall performance,
which agree with the rest of our results on CHT resprouts.
CONCLUSIONS
Despite numerous methods to control the invasive tree Ailanthus altissima that have
been previously tested (e.g. Meloche and Murphy 2006), none of them has been tested
or monitored along long-term periods and under Mediterranean conditions. We tested
three different treatments on resprout type, resprout density, and plant growth of A.
altissima over a five year period. Our results indicated that joint cut stump and herbicide
application is the only effective treatment in the long-term to reduce A. altissima in
Mediterranean forests. Contrary to our expectations, the double cut stump treatment did
not reduce the invasive species growth, and had similar results than the single stump
treatment.
Passive management on A. altissima resprouts should consider all negative
ecological effects of the species on ecosystem diversity and function. Methods
involving the physical removal of stumps (cutting) are ineffective, even where stumps
were cut again the following years (our study shows results from five years of
monitoring, but this particular treatment has been applied in the study area during 10
148
years without any success in A. altissima eradication). Although the combination of
mechanical and chemical treatments did not reduce the number of resprouts over a fiveyear period, they did reduce resprout biomass, height and leaf area index. A combined
herbicide cutting treatment should be included in plans of management for
Mediterranean protected areas and the technique should be monitored over the longterm to assess for its real success and to ensure native species recolonization and
ecosystem recovery. Since stump removal and glyphosate application reduces the
competitive ability of the invasive species, it should be used together with species
afforestation restoration programmes to improve the success of A. altissima control.
ACKNOWLEDGEMENTS
D. Eldridge, S. Soliveres and two anonymous reviewers provided helpful comments and
improvements on earlier versions of this manuscript. Also thank Font Roja Natural Park
personnel (in special, Abraham Santonja†) and all collaborators and volunteers for their
help in the fieldwork. This research and SCN PhD fellowship were supported by the
project
(GV06/029)
founded
by
Generalitat
Valenciana,
ESTRES
Project
(063/SGTB/2007/7.1) and RECUVES Project (077/RN08/04.1) founded by the Spanish
Ministerio de Medio Ambiente, and BAHIRA CICYT project (CGL2008-03649/BTE)
founded by the Spanish Ministerio de Ciencia y Tecnología. Font Roja Natura UA
Scientific Station (ECFRN UA), depending on the Office of the Vice President for
Research, Development and Innovation (VIDI) of the University of Alicante, supported
also this research.
149
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151
APÉNDICE FOTOGRÁFICO
Foto 1 Parcela de 2 × 2 m para
determinar la mejor estrategia de
control sobre A. altissima
Foto 2 Aplicación de herbicida sobre
tocones recién cortados de A.
altissima
Foto 3 Población de A. altissima tras
4 años de tratamiento de dos
desbroces anuales
Foto 4 Parcela sin A. altissima tras 4
años de tratamiento de desbroce y
herbicida
152
AGRADECIMIENTOS
Quiero ser breve pues han sido unos cuantos añitos y no quiero enrollarme más, ni mirar
hacia detrás para evitar emociones,… así que, mis agradecimientos van a A. Bonet
(director de tesis), a la Estación Científica Font Roja Natura (Vicerrectorado de
Investigación, Desarrollo e Innovación, Universidad de Alicante), a tod@s mis amig@s
(cada uno de ellos saben quienes son ☺), a los colaboradores, y a toda la gente buena
que he tenido el placer de conocer y de tratar estos años. A mi familia (mascotas
incluidas), por TODO. A Santi, por estar siempre ahí y por hacer mejor cada segundo
de mi VIDA.
Así pues, INFINITAS GRACIAS a tod@s.
Proyectos
GV06/029. Generalitat Valenciana. España
Recuves (077/RN08/04.1). Ministerio de Medio Ambiente. España
Estrés (063/SGTB/2007/7.1). Ministerio de Medio Ambiente. España
Entidades colaboradoras
Conselleria d’Infraestructures, Territori i Medi Ambient
CEMACAM Font Roja-Alcoy
Gerencia de Medi Ambient. Alcoy
153
154
AFILIACIÓN DE LOS COAUTORES
Andreu Bonet Jornet
Departamento de Ecología, Universidad de Alicante, 03080, Alicante, Spain.
Instituto Multidisciplinar para el Estudio del Medio Ramón Margalef. Universidad de
Alicante, 03080 Alicante, Spain. E-mail: andreu@ua.es
Santiago Soliveres Codina
Área de Biodiversidad y Conservación, Departamento de Biología y Geología, Escuela
Superior de Ciencias Experimentales y Tecnología, Universidad Rey Juan Carlos,
28933 Móstoles, Spain. E-mail: santiago.soliveres@urjc.es
Lluís Serra Laliga
Generalitat Valenciana, Conselleria d’Infraestructures, Territori i Medi Ambient, SS.
TT. d’Alacant. C/Churruca nº 29, 03071 Alacant, Spain. E-mail: serralaliga@yahoo.es
Rubén Torices Blanco
Centro de Ecologia Funcional. Departamento de Ciências da Vida. Universidade de
Coimbra. 3001–455. Coimbra, Portugal. E-mail: rubentorices@gmail.com
155
Estrella Pastor Llorca
Estación Científica Font Roja Natura UA. Universidad de Alicante, s/n 03801 Alcoy,
Spain. E-mail: estrella.pastor@ua.es
María José Lledó Solbes
Departamento de Ecología, Universidad de Alicante, 03080, Alicante, Spain. E-mail:
mj.lledo@ua.es
156
Reunido el Tribunal que suscribe en el día de la fecha, acordó otorgar, por
Tesis Doctoral de Don/Doña.
Alicante,
de
la calificación de
de
.
El Secretario,
El Presidente
157
a la
.
158
UNIVERSIDAD DE ALICANTE
Comisión de Doctorado
La presente Tesis de Don/Doña.
ha sido
registrada con el nº
del registro de entrada correspondiente.
Alicante,
de
de
.
El Encargado del Registro,
159