an experimental approach with Carpobrotus edulis (L.) NE Br.

Determinants of plant invasiveness in clonal

species: an experimental approach with

Carpobrotus edulis (L.) N. E. Br. and

Alternanthera philoxeroides (Mart.) Griseb.

Rubén Portela Carballeira

Doctoral Thesis UDC / 2019

Director: Sergio Rodríguez Roiloa

Tutor: Rodolfo Barreiro Lozano

PhD Program in Marine Sciences, Technology and Management

SERGIO RODRÍGUEZ ROILOA y RODOLFO BARREIRO LOZANO, Profesor

Contratado Doctor Interino y Catedrático de Ecología respectivamente, del Departamento

de Biología de la Universidade de A Coruña

DECLARAN:

Que la siguiente memoria titulada “Determinants of plant invasiveness in clonal

species: an experimental approach with Carpobrotus edulis (L.) N. E. Br. and

Alternanthera philoxeroides (Mart.) Griseb.” presentada por Don RUBÉN PORTELA

CARBALLEIRA ha sido realizada bajo su dirección en el Departamento de Biología de

la Universidade de A Coruña dentro del Programa Oficial Internacional de Doctorado

DO*MAR Marine Science, Technology and Management regulado por el RD 99/2011, y

cumple con las condiciones exigidas para ser defendida y optar al grado de “Doctor

Internacional” ante el tribunal que lo deberá juzgar.

Y para que así conste a los efectos oportunos, firman la presente en A Coruña a 15 de

septiembre de 2019.

El director, El tutor,

Dr. Sergio Rodríguez Roiloa Dr. Rodolfo Barreiro Lozano

SERGIO RODRÍGUEZ ROILOA and RODOLFO BARREIRO LOZANO,

Associate Professor and Professor of Ecology, respectively, from the Department of

Ecology of the University of A Coruña

CERTIFY:

That the following report entitled “Determinants of plant invasiveness in clonal

species: an experimental approach with Carpobrotus edulis (L.) N. E. Br. and

Alternanthera philoxeroides (Mart.) Griseb.” written by Mister RUBÉN PORTELA

CARBALLEIRA has been prepared under their supervision in the Department of

Biology at the Science Faculty of the University of A Coruña, within the framework of

the Official PhD International Program DO*MAR Marine Science, Technology and

Management regulated by Royal Decree no. 99/2011 and it meets the requirements to be

defended and to aspire to the degree of “International PhD”.

And for any legal statement, the present document is signed in A Coruña, September 15,

2019.

The director, The tutor,

Dr. Sergio Rodríguez Roiloa Dr. Rodolfo Barreiro Lozano

“La suerte favorece solo a la mente preparada”

Isaac Asimov

A mis padres

Contents

Summaries 11

Preface (Spanish) 21

Extended summary (Spanish)

23

General introduction

41

SECTION I - Carpobrotus spp.

Introduction to Carpobrotus spp.

59

Chapter 1. Physiological integration buffers sand burial stress in the clonal

plant Carpobrotus edulis invading a coastal dune in NW Iberia

75

Chapter 2. Effects of clonal integration in the expansion of two alien

Carpobrotus species into a coastal dune system – a field experiment

93

Chapter 3. Biomass partitioning in response to resources availability: A

comparison between native and invaded ranges in the clonal invader

Carpobrotus edulis

111

Chapter 4. Importance of plasticity in response to soil nutrient content and

competitive ability in explaining invasiveness of the clonal Carpobrotus

edulis: a trans-continental study

131

SECTION II – Alternanthera philoxeroides

Introduction to Alternanthera philoxeroides

159

Chapter 5. Effects of physiological integration on defense strategies against

herbivory by the clonal plant Alternanthera philoxeroides

173

Chapter 6. Trans-generational effects in the clonal weed Alternanthera

philoxeroides

197

Chapter 7. A dynamic model-based framework to test the effectiveness of

biocontrol targeting a new plant invader – the case of Alternanthera

philoxeroides in the Iberian Peninsula

219

Conclusions 267

Acknowledgments (Spanish) 271

Supplementary material 273

11

Abstract (English)

Biological invasions are one of the main causes of global biodiversity loss. The reason

why only a few alien species become invasive has yet to be clarified. In this doctoral

thesis, a series of experiments have been conducted to elucidate the role played by

different traits associated with clonal reproduction in biological invasions. In chapters I

and II, field experiments were carried out to investigate the benefit of physiological

integration in Carpobrotus spp. Chapters III and IV delve into the selection of phenotypic

plasticity and the competitive ability of Carpobrotus spp. throughout the processes of

biological invasions. Chapter V focuses on the role of physiological integration in the

defensive response to real and simulated herbivory by the invasive plant Alternanthera

philoxeroides. Chapter VI evaluates the role of DNA methylation as an epigenetic

transmission mechanism of phenotypic plasticity for this species. Finally, in chapter VII

a dynamic simulation model for the biocontrol of A. philoxeroides is proposed, using the

insect Agasicles hygrophila in a model population located in Fisterra, Galicia (NW Spain).

13

Abstract (Spanish)

Las invasiones biológicas son una de las principales causas de pérdida de biodiversidad

a nivel global. El motivo por el cual algunas especies exóticas se conviertan en

invasoras mientras que otras no todavía no se ha esclarecido. En esta tesis doctoral se

han realizado una serie de experimentos para dilucidar el papel que juegan en las

invasiones biológicas diferentes rasgos asociados a la reproducción clonal de las plantas.

En los capítulos I y II se realizaron experimentos de campo para investigar el beneficio

de la integración fisiológica en Carpobrotus spp. Los capítulos III y IV ahondan en la

selección de la plasticidad fenotípica y la habilidad competitiva de Carpobrotus spp. A

lo largo de los procesos de invasiones biológicas. El capítulo V se centra en el papel de

la integración fisiológica en las respuestas defensivas frente a herbivoría, real y

simulada, de la planta invasora Alternanthera philoxeroides. El capítulo VI evalúa el

papel de la metilación del ADN como mecanismo epigenético de transmisión de la

plasticidad fenotípica para esta especie. Finalmente, en el capítulo VII se propone un

modelo de simulación dinámico para el biocontrol de A. philoxeroides empleando el

insecto Agasicles hygrophila en una población modelo localizada en Galicia.

15

Abstract (Galician)

As invasións biolóxicas son una das causas da perda global de biodiversidade. A razón

pola que só algunhas especies exóticas se volven invasoras aínda non foi esclarecida.

Nesta tese doutoral realizáronse unha serie de experimentos para dilucidar o papel que

desempeñan nas invasións biolóxicas diferentes atributos asociados coa reprodución

clonal. Nos capítulos I e II realizáronse experimentos de campo para investigar o

beneficio da integración fisiolóxica en Carpobrotus spp. Os capítulos III e IV afondan na

selección da plasticidade fenotípica e máis da capacidade competitiva de Carpobrotus

spp. ó longo dos procesos de invasións biolóxicas. O capítulo V céntrase no papel da

integración fisiolóxica nas respostas defensivas fronte á herbivoria, real e simulada, da

planta invasora Alternanthera philoxeroides. O capítulo VI avalía o papel da metilación

do ADN como mecanismo epixenético de transmisión da plasticidade fenotípica desta

especie. Finalmente, no capítulo VII proponse un modelo de simulación dinámica para o

biocontrol de A. philoxeroides, empregando o escaravello Agasicles hygrophila, nunha

poboación modelo localizada en Fisterra, Galicia.

17

Abstract (Portuguese)

As invasões biológicas são uma das principais causas da perda global de biodiversidade.

A razão pela qual apenas algumas espécies exóticas se tornam invasivas ainda não foi

esclarecida. Nesta tese de doutorado, uma série de experimentos foram feitos para

elucidar o papel desempenhado por diferentes características associadas à reprodução

clonal de plantas nas invasões biológicas. Nos Capítulos I e II, experimentos de campo

foram conduzidos para investigar o benefício da integração fisiológica em Carpobrotus

spp. Os capítulos III e IV inquirem na seleção da plasticidade fenotípica e da capacidade

competitiva de Carpobrotus spp. ao longo dos processos de invasões biológicas. Capítulo

V incide sobre o papel da integração fisiológica em respostas defensivas à herbivoria, real

e simulada, da planta invasora Alternanthera philoxeroides. O Capítulo VI avalia o papel

da metilação do DNA como um mecanismo epigenético de transmissão da plasticidade

fenotípica para esta espécie. Finalmente, no capítulo VII é proposto um modelo de

simulação dinâmica para o biocontrole de A. philoxeroides usando o inseto Agasicles

hygrophila numa população modelo localizada em Fisterra, Galiza (NW Espanha).

19

Abstract (Chinese)

生物入侵是全球生物多样性丧失的主要原因之一。导致少数外来物种成为入侵物

种的原因尚未澄清。在这篇博士论文中，一系列控制实验被开展，以阐明与克隆

繁殖相关的不同性状在生物入侵过程中所起的作用。在第一章和第二章中，进行

田间试验以研究多肉入侵植物 Carpobrotus spp. 生理整合的生态学价值。第三章和

第四章深入研究 Carpobrotus spp. 的表型可塑性和竞争能力在生物入侵过程中的适

应性选择。第五章重点讨论生理整合在入侵植物空心莲子草（Alternanthera

philoxeroides）抵抗真实和模拟食草环境中的作用。第六章评估 DNA 甲基化作为

空心莲子草表型可塑性的表观遗传传递机制的作用。最后，在第七章中，通过研

究昆虫莲草直胸跳甲(Agasicles hygrophila)与入侵加利西亚地区（西班牙西北部）

的空心莲子草种群的交互影响，提出一种用于生物防治空心莲子草的动态模拟模

型。

21

Prefacio

Los comienzos de esta tesis doctoral se remontan a la primavera del año 2015, cuando yo

me encontraba cursando el Máster en Ciencias, Tecnología y Gestión Ambiental en la

Universidade de A Coruña. Debía escoger un tema para mi TFM, que podía consistir bien

en realizar prácticas en una empresa o bien llevar a cabo un trabajo de investigación en la

universidad. Puesto que mi idea en aquella época era cursar el Máster en Ciencias Marinas

con posterioridad, me pareció buena idea realizar un trabajo de investigación sobre

ecotoxicología en organismos marinos. Con ese propósito me dirigí al despacho del

profesor Rodolfo Barreiro Lozano, ya que era quien impartía la asignatura de

ecotoxicología. Al no encontrarlo, me dirigí al despacho más cercano, el del profesor

Sergio Rodríguez Roiloa, para preguntarle si sabía cuándo estaría disponible el profesor

Barreiro. Al comentarle mi idea sobre el TFM, me preguntó si las plantas invasoras

clonales me parecían un campo de investigación interesante. Le respondí que sí.

Cuatro años después, con varios artículos científicos publicados y tras haber realizado

estancias de investigación en China, Brasil y Portugal, me lo siguen pareciendo.

23


1. Conceptos generales sobre las invasiones biológicas

Mientras que el flujo de especies es un fenómeno natural tan antiguo como la

propia vida en la Tierra, los seres humanos han sido capaces de acelerarlo hasta límites

que no eran posibles anteriormente (McNeely 2001; Meyerson & Mooney 2007). Hoy en

día, las especies pueden ser transportadas miles de kilómetros en cuestión de días u horas,

atravesando montañas y océanos. Sin embargo, no todas las especies que son introducidas

en un nuevo hábitat sobreviven, y no todas aquellas que sobreviven son capaces de

competir de forma exitosa con las especies nativas. Se denomina especie exótica o no-

nativa a aquella introducida fuera de su hábitat natural (Colautti & MacIsaac 2004).

Aquellas especies exóticas que son capaces de desplazar a las especies nativas se

consideran especies invasoras. Las invasiones biológicas son el proceso por el cual una

especie es introducida fuera de su hábitat natural debido a la acción voluntaria o

accidental del hombre, adaptándose al nuevo ambiente y proliferando de tal forma que

altera el ecosistema y desplaza a las especies nativas (Levine et al. 2003; Mack et al.

2000; Vitousek et al. 1996). Los dos factores principales que contribuyen a las invasiones

biológicas son la capacidad de las especies invasoras de expandirse y volverse dominantes

en el hábitat invadido (lo que se conoce como invasividad) y aquellas características del

hábitat que favorecen los procesos de invasión (invasibilidad) (Richardson & Pyšek

2006). Aquellos hábitats que han sido alterados por la acción del hombre son más

propensos a sufrir invasiones biológicas (Alpert et al. 2000).

El establecimiento y proliferación de especies invasoras puede modificar la

estabilidad y funcionalidad de las comunidades locales y desplazar a las especies nativas,

con la consiguiente degradación de los ecosistemas. Las invasiones biológicas son la

segunda causa principal de pérdida de biodiversidad a nivel global, solo por detrás de la

destrucción de hábitats (Mack et al. 2000; Vitousek et al. 1996). Esto se debe

principalmente a la capacidad competitiva de las especies invasoras, que da lugar en

muchas ocasiones a comunidades con una menor diversidad de especies e incluso

monoespecíficas. Una de las cuestiones clave en el estudio de las invasiones biológicas

es la de intentar determinar por qué algunas especies introducidas se convierten en


24

invasoras, mientras que otras permanecen como exóticas, sin causar impactos

significativos en el ecosistema (van Kleunen et al. 2011). A pesar de los importantes

avances conseguidos en este campo durante los últimos años, los procesos de invasión,

sus causas y consecuencias, todavía están lejos de estar completamente esclarecidos, y

representan por lo tanto un atractivo reto para la ciencia en general y la ecología en

particular. Determinar qué rasgos favorecen la expansión y dominancia de las especies

invasoras se ha convertido en uno de los principales retos dentro de la ecología moderna

(Pyšek & Richardson 2008; Richardson & Pyšek 2006; Thuiller et al. 2006; van Kleunen

et al. 2011).

2. El papel de la reproducción clonal en las invasiones biológicas

Una característica común de muchas especies vegetales y que ha sido asociada a las

invasiones biológicas es la reproducción asexual. Esta consiste en la reproducción a partir

de un único individuo parental dando lugar a descendientes genéticamente idénticos. Por

este motivo, la reproducción asexual también se denomina reproducción o crecimiento

clonal. El crecimiento clonal se caracteriza por la producción vegetativa de un número

indeterminado de descendientes, denominados rametos, dispuestos a intervalos más o

menos regulares sobre tallos modificados que crecen en superficie (estolones) o bajo la

superficie del suelo (rizomas) (Klimeš et al. 1997). El crecimiento clonal ha sido señalado

como una característica que podría contribuir a la invasividad de algunas especies (Liu et

al. 2006; Pyšek 1997; Roiloa et al. 2015; Song et al. 2013). Esta idea se basa en el hecho

de que muchas de las especies de plantas invasoras más exitosas presentan reproducción

clonal. En este sentido, el 67% de las especies exóticas más agresivas de Europa, el 47%

en Norteamérica, el 54% en Sudamérica y el 51% en Australia muestran crecimiento

clonal (Pyšek 1997). Esto ha sido avalado por estudios posteriores, en los que se ha

determinado que la clonalidad está positivamente correlacionada con la invasividad de

las especies exóticas en diferentes regiones (Liu et al. 2006; Shah et al. 2014). Por otra

parte, los estolones y rizomas juegan un importante papel como reservorio de agua y

carbohidratos, favoreciendo la supervivencia de las plantas en situaciones de estrés, o al

producirse la fragmentación del sistema clonal (Goulas et al. 2001; Stuefer & Huber 1999;

Suzuki & Stuefer 1999). Esto puede jugar un papel crucial en la colonización de nuevos

entornos por especies clonales invasoras, especialmente tras un proceso de fragmentación

(Dong et al. 2010, 2011, 2012; Konlechner et al. 2016; Lin et al. 2012).


25

La mayoría de especies que poseen reproducción clonal también pueden recurrir

de forma facultativa a la reproducción sexual (Yang & Kim 2016). Ambos tipos de

reproducción presentan distintas ventajas e inconvenientes (Lei 2010). Así, la ventaja más

evidente de la reproducción clonal es que puede llevarse a cabo a partir de un único

individuo, que eventualmente puede dar lugar a una población de clones (van der Merwe

et al. 2010). La parte negativa de esto es que apenas existe variabilidad genética entre los

individuos originados por reproducción clonal, mientras que la reproducción sexual

implica la recombinación genética de ambos progenitores, dando lugar a individuos con

combinaciones únicas de genes (Bürger 1999). Esto permite la actuación de la selección

natural a nivel de individuos, no de poblaciones enteras como sucedería en el caso de

clones, lo cual favorece la adaptación y supervivencia de las poblaciones que presentan

reproducción sexual. Por otra parte, cuando un individuo está bien adaptado a las

características de su entorno, la reproducción clonal es preferible a la sexual, ya que los

descendientes estarán igualmente bien adaptados a ese entorno (Otto 2009). Se evita así

el gasto de recursos en la producción de órganos sexuales (flores y frutos), que conllevan

un coste importante para las plantas (Griffiths & Bonser 2013; Roze 2012). Finalmente,

debe tenerse en cuenta que la reproducción sexual permite la dispersión de nuevos

individuos mediante semillas, que pueden atravesar largas distancias y están adaptadas

para sobrevivir a condiciones desfavorables, mientras que la reproducción clonal es

llevada a cabo generalmente por medio de tallos o raíces, lo cual otorga a la planta una

capacidad de dispersión bastante limitada (Vittoz & Engler 2007; von der Lippe &

Kowarik 2007).

Una de las características más interesantes asociada al crecimiento clonal de las

plantas es la capacidad para la integración fisiológica, también conocida como integración

clonal. La integración fisiológica permite el transporte de recursos y otras sustancias entre

los distintos individuos de un sistema clonal mientras estos permanezcan conectados

mediante estolones o rizomas (Price & Marshall 1999). La conexión entre los distintos

rametos de un sistema clonal puede mantenerse por un tiempo indefinido (Price &

Marshall 1999). El intervalo en el que las conexiones entre los rametos permanecen

funcionales varía considerablemente entre especies y se ve afectado por las condiciones

ambientales (Jónsdóttir & Watson 1997). En algunos casos la conexión deja de ser

funcional inmediatamente tras de la producción del nuevo rameto (Jónsdóttir & Watson

1997), mientras que en otros puede mantenerse durante años (Eriksson & Jerling 1990).


26

El intercambio de recursos y sustancias incluye agua, nutrientes, hormonas o

fotoasimilados, además de señales defensivas (Stuefer et al. 2004). Esta capacidad de

integración entre rametos mejora la supervivencia del sistema clonal, al aportar recursos

a aquellos rametos creciendo en condiciones desfavorables o sometidos a estrés (Hartnett

& Bazzaz 1983; Roiloa & Retuerto 2006; Saitoh et al. 2002). Estudios recientes han

demostrado la importancia de la integración clonal en la expansión de diversas especies

invasoras (Liu et al. 2009; Otfinowski & Kenkel 2008; Roiloa et al. 2014a, 2014b, 2016;

Wang et al. 2008). Un meta-análisis realizado por Song et al. (2013) encontró una

correlación entre la invasividad de diferentes especies exóticas y el beneficio de la

integración clonal para aquellos rametos que crecen en condiciones desfavorables,

poniendo de manifiesto la relación entre la integración clonal y la capacidad invasora

(Song et al. 2013). Otro fenómeno asociado a la integración clonal es la división de

trabajo entre los distintos individuos que forman el sistema clonal, es decir, la capacidad

de especializarse en distintas tareas (Stuefer et al. 1996). Esto resulta especialmente

ventajoso en ambientes con una distribución heterogénea de recursos (Hutchings &

Wijesinghe 1997; Roiloa et al. 2007). Así, los rametos se especializan en la obtención de

aquellos recursos que son más abundantes localmente, incrementando por lo tanto la

eficiencia en su obtención, y distribuyéndolos posteriormente entre los distintos módulos

del sistema clonal (Stuefer et al. 1998).

3. Superando las desventajas de la reproducción clonal

Cuando una especie es introducida fuera de su hábitat natural, debe adaptarse a

nuevas condiciones ambientales. Esto es particularmente importante en especies clonales,

debido a la falta de variabilidad genética antes mencionada. Por ello, la plasticidad

fenotípica es una característica que podría jugar un papel clave en las invasiones

biológicas de plantas clonales (Davidson et al. 2011; Keser et al. 2014; Richards et al.

2006). La plasticidad fenotípica es la capacidad de un único genotipo de producir

diferentes fenotipos dependiendo de las características del ambiente en el que se

desarrolla (Bradshaw 1965; Sultan 2000). Un ejemplo de esta plasticidad es la

distribución de la biomasa en aquellos órganos responsables de obtener el recurso más

limitante para el crecimiento de la planta (Gleeson & Tilman 1992; Hilbert 1990; Weiner

2004). Por ejemplo, cuando una planta crece en un ambiente pobre en nutrientes se espera

que priorice el desarrollo de raíces. Si la plasticidad fenotípica permite una mayor


27

eficacia biológica en distintos ambientes, se espera que sea fomentada por la selección

natural (Fusco & Minelli 2010; Ghalambor et al. 2007). Así pues, el estudio de la

plasticidad fenotípica de individuos de especies invasoras procedentes tanto del área

nativa como del área invadida puede ser un enfoque adecuado para conocer al papel que

este rasgo juega en las invasiones biológicas (van Kleunen et al. 2010). Asimismo, la

comparación de la plasticidad fenotípica entre diferentes especies invasoras y congéneres

no invasores ha mostrado una mayor respuesta plástica de las especies invasoras hacia la

disponibilidad de agua, diferentes niveles de nutrientes y diferentes concentraciones de

CO2, lo que refuerza la noción de que este rasgo contribuye significativamente a la

capacidad invasora (Funk 2008; Geng et al. 2006; Raizada et al. 2009; Wei et al. 2017).

Otro factor que puede ayudar a compensar la falta de variabilidad genética en

plantas clonales son los mecanismos epigenéticos. Se trata de una serie de modificaciones

químicas en el ADN que alteran su expresión, pero no la secuencia de nucleótidos (Gao

et al. 2010; Wolffe & Matzke 1999). Estas modificaciones pueden aumentar la

expresividad de un gen o bien silenciarlo completamente, evitando que se exprese.

Distintos estudios han señalado que la influencia de las condiciones ambientales en los

fenotipos podría estar mediada por este tipo de mecanismos (Bossdorf et al. 2010;

Hallgrímsson & Hall 2011; Verhoeven & Preite 2014). Los mecanismos de regulación

epigenética del ADN consisten en modificaciones reversibles, heredables y que pueden

ser alteradas de forma más flexible que la secuencia del genoma (Heard & Martienssen

2014; Martienssen & Colot 2001). Una modificación epigenética del ADN que ha sido

ampliamente documentada en diversas especies es la metilación de la citosina (y, de

forma menos extendida, de la adenina) (Bender 2004). Cuando la metilación se produce

en el segmento promotor de un gen, inhibe la transcripción del mismo. Así pues, los

mecanismos epigenéticos otorgan a aquellas especies con reproducción asexual una

alternativa eficiente a la recombinación genética como fuente de variabilidad. Se ha

sugerido que las regulaciones epigenéticas en la expresión génica permitirían el

establecimiento de especies invasoras a corto plazo (Pérez et al. 2006). Estudios recientes

han encontrado una correlación entre la variabilidad epigenética y fenotípica de especies

invasoras clonales agresivas con baja diversidad genética, indicando que los mecanismos

de regulación epigenética son una fuente alternativa de variabilidad (Gao et al. 2010;

Wang et al. 2019). Sin embargo, el papel que juegan los mecanismos epigenéticos en la

invasividad de las especies clonales no ha sido estudiado en profundidad.


28

4. Objetivos

Entender los mecanismos implicados en los procesos de invasiones biológicas es

clave para predecir futuros escenarios de invasión y para el diseño de estrategias eficientes

para el control y restauración de áreas invadidas. El objetivo general de esta tesis doctoral

es contribuir a determinar el papel que el crecimiento clonal, y distintos atributos

asociados a este, juegan en los procesos de las invasiones vegetales. Con dicho propósito,

se llevaron a cabo una serie de experimentos en los que se comprobó el beneficio que

suponían diversas características asociadas a la reproducción clonal en dos especies

invasoras, Carpobrotus edulis (L.) N. E. Br. y Alternanthera philoxeroides (Mart.)

Griseb. Ambas especies presentan reproducción clonal mediante estolones y son

consideradas invasoras agresivas, ya que causan graves alteraciones en los ecosistemas

en los que son introducidas. Con el fin de determinar la importancia del crecimiento clonal

en la capacidad invasora de las dos especies estudiadas se realizaron un total de siete

experimentos, incluyendo trabajos de campo, experimentos manipulativos en condiciones

controladas de jardín común, y trabajos de modelización. Los principales resultados se

describen a continuación:

Capítulo I

El desplazamiento de arena debido a la acción del viento es un fenómeno común

en ambientes desérticos y sistemas dunares costeros. Las plantas que crecen en este tipo

de hábitats deben estar adaptadas para sobrevivir a eventos de enterramiento parcial

(Brown 1997). Se trata de un factor severo de estrés, pues modifica parámetros claves

para la supervivencia vegetal, como la incidencia de luz, la humedad o la temperatura

(Baldwin & Maun 1983; Maun & Lapierre 1984). Aquellas plantas que presentan un

crecimiento postrado, al carecer de tallos ortotrópicos que les permitan alzar sus hojas

sobre la arena, requieren otros mecanismos para sobrevivir (Chen et al. 2010; Yu et al.

2002). Se realizó un experimento de campo con la especie clonal invasora C. edulis para

determinar el papel que la integración fisiológica juega en la respuesta de esta planta

frente al enterramiento parcial por arena. Fragmentos clonales compuestos por cuatro

rametos fueron colocados en una duna costera, donde los dos rametos apicales fueron

enterrados y la integración fisiológica con los rametos basales fue permitida o impedida.

El grupo de control consistió en fragmentos clonales no sometidos a enterramiento ni


29

fragmentación. El experimento tuvo una duración de 17 semanas y se realizó en la playa

de Seselle, Ares (A Coruña) en primavera del año 2018. Los resultados mostraron que la

integración fisiológica indujo una respuesta plástica no local en los rametos basales,

dependiendo de las condiciones experimentadas por sus rametos apicales. Así, cuando los

rametos apicales permanecieron sin enterrar, los resultados mostraron una división del

trabajo programada desde el punto de vista del desarrollo, con rametos basales

especializados en la adquisición de recursos a través de sus raíces, mientras que los

rametos apicales desarrollaron su parte aérea. Por el contrario, cuando los rametos

apicales fueron enterrados en la arena, los rametos basales cambiaron su patrón de

asignación de biomasa y aumentaron la producción de estructuras fotosintéticas. En

conclusión, la integración fisiológica permitió a los rametos apicales de C. edulis

sobrevivir al enterramiento, e impidió la pérdida de biomasa a nivel de los fragmentos

clonales al sufrir enterramiento, lo que puede tener importantes consecuencias para

entender el éxito invasor de esta especie clonal.

Capítulo II

La integración fisiológica resulta beneficiosa para los sistemas clonales, al

permitir el transporte de recursos desde aquellos rametos que ya se encuentran

establecidos hacia los rametos más jóvenes, promoviendo su desarrollo y así la expansión

del sistema clonal (Hartnett & Bazzaz 1983; Roiloa & Retuerto 2006). Se llevó a cabo un

experimento de campo con las especies invasoras C. edulis y Carpobrotus acinaciformis

(L.) L. Bolus, considerado por algunos autores como menos invasivo que C. edulis debido

a que su distribución en Europa es menor (Lambinon 1995; Suehs et al. 2001). En este

experimento se emplearon fragmentos clonales con dos individuos, teniendo el rameto

basal acceso a nutrientes al crecer sobre turba, mientras que el rameto apical se

desarrollaba sobre arena. La integración fisiológica entre ambos podía estar permitida o

impedida. El experimento tuvo una duración de 12 semanas y se realizó en la playa de

Seselle, Ares (A Coruña) en primavera-verano del año 2015. Los resultados del

experimento mostraron el beneficio que supone la integración fisiológica para ambas

especies de plantas clonales cuando estas se desarrollan en ambientes con una distribución

heterogénea de nutrientes. Asimismo, mientras que el beneficio derivado de la integración

clonal por sí mismo no explica las diferencias en la capacidad de invasión entre estas dos

especies exóticas, los resultados indican que la mayor invasividad de C. edulis podría


30

deberse a una mayor capacidad para amortiguar el efecto negativo de la fragmentación en

comparación con C. acinaciformis.

Publicado en: Portela, R. and S. R. Roiloa (2017). Effects of clonal integration in the

expansion of two alien Carpobrotus species into a coastal dune system – a field

experiment. Folia Geobotanica 52(3-4), 327-335. doi: 10.1007/s12224-016-9278-4

Capítulo III

La plasticidad fenotípica es otro rasgo que parece favorecer las invasiones

biológicas, pues permite a las plantas adaptarse a una amplia variedad de condiciones

ambientales (Davidson et al. 2011; Keser et al. 2014; Richards et al. 2006). Esto es

especialmente relevante en plantas clonales, al permitir compensar en parte la falta de

variabilidad genética entre individuos. Para comprobar si este rasgo juega un papel

relevante en los procesos de invasión de C. edulis, se emplearon individuos procedentes

del área nativa de la especie (Sudáfrica) y otros procedentes del área invadida (Península

Ibérica) y se comparó la respuesta plástica de ambas poblaciones al estar sometidos a la

escasez de diferentes recursos. Las plantas, compuestas en este experimento por rametos

individuales, fueron sometidas a tres tratamientos diferentes: escasez de agua, escasez de

nutrientes o escasez de luz. La respuesta frente a los tratamientos se comparó con un

grupo control que creció con abundancia de los tres recursos. El experimento fue

realizado en una cámara de crecimiento en el laboratorio de Ecología de la Universidade

da Coruña (UDC), y tuvo una duración de cinco semanas, llevándose a cabo en la

primavera del año 2016. Los resultados de este experimento mostraron que la distribución

de biomasa en respuesta a la disponibilidad de nutrientes en C. edulis difiere entre

poblaciones de áreas de nativas e invadidas. Así, las plantas procedentes del área invadida

tuvieron una mayor respuesta plástica frente a la escasez de nutrientes, consistente en un

mayor desarrollo de las raíces, apoyando la hipótesis de que este rasgo ha sufrido una

selección adaptativa durante el proceso de invasión. Esta respuesta plástica de forrajeo

puede contribuir a la optimización de la absorción de nutrientes por parte de las plantas

y, por lo tanto, podría considerarse como una estrategia de adaptación. Sin embargo, esta

respuesta no se observó en los otros tratamientos experimentales. La ausencia de

respuesta ante la escasez de agua podría deberse a que esta especie está bien adaptada al

estrés hídrico, por lo que la corta duración del experimento no permitió apreciar una


31

respuesta. En cuanto al estrés por falta de luz, se encontró una respuesta plástica por parte

de las plantas (menor desarrollo de las raíces), pero esta respuesta fue idéntica entre ambas

poblaciones. Comprender las implicaciones ecológicas de los cambios en la distribución

plástica de biomasa es importante para determinar los procesos de adaptación de las

plantas a nuevos entornos, y contribuye a desenmarañar los mecanismos subyacentes a la

capacidad invasora de las plantas.

Publicado en: Portela, R., Barreiro, R., & Roiloa, S. R. (2019). Biomass partitioning in

response to resources availability: a comparison between native and invaded ranges in the

clonal invader Carpobrotus edulis. Plant Species Biology 34(1), 11-18. doi:

10.1111/1442-1984.12228

Capítulo IV

Este experimento continúa el trabajo desarrollado en el capítulo III,

incrementando el número de poblaciones de C. edulis (una población del área nativa,

Sudáfrica, y tres poblaciones de áreas invadidas, Península Ibérica, California y

Australia), y utilizando otras especies del género Carpobrotus con distinto grado de

invasividad (C. acinaciformis, especie invasora en la Península Ibérica; Carpobrotus

chilensis (Molina) N. E. Br., especie exótica no invasora presente en California y

Carpobrotus virescens (Haw.) Schwantes, especie nativa de Australia), así como una

especie nativa de Europa, Ammophila arenaria (L.) Link (especie clonal que habita dunas

costeras, invasora en otros países). Por una parte, el experimento buscaba comparar el

grado de plasticidad fenotípica de las distintas especies de Carpobrotus, así como de las

distintas poblaciones de C. edulis, al enfrentarse a estrés por escasez de nutrientes. Por

otra parte, también se comparó la habilidad competitiva de las diferentes especies y

poblaciones de Carpobrotus. El experimento fue realizado en jardín común en la Facultad

de Ciencias de la UDC en la primavera del año 2017 y tuvo una duración de 20 semanas.

Los resultados de este experimento indican la presencia de una selección

adaptativa durante el proceso de invasión de C. edulis. Así, las poblaciones procedentes

de áreas no nativas mostraron un crecimiento significativamente mayor en respuesta a un

incremento de nutrientes que las poblaciones de la zona de distribución nativa. Sin

embargo, las diferencias detectadas en el crecimiento de las plantas no se transfirieron a

una mayor habilidad competitiva en poblaciones de áreas de distribución no nativas. Por


32

otra parte, se encontró un mayor beneficio de la adición de nutrientes, en términos de

aumento de la biomasa total, en C. edulis procedente de California que en el congénere

menos invasivo C. chilensis, lo que sugiere que la respuesta plástica al contenido de

nutrientes del suelo podría explicar las diferencias en la invasividad de ambas especies.

Además, contrariamente a nuestra hipótesis, también se encontró un mayor beneficio de

la adición de nutrientes en C. acinaciformis respecto a C. edulis procedente de la

Península Ibérica. Un mayor estudio de la relevancia de la habilidad competitiva en los

hábitats costeros invadidos por C. edulis será necesario para dilucidar el papel que este

rasgo desempeña en las invasiones biológicas de esta especie.

Capítulo V

La integración fisiológica, además de posibilitar el transporte de recursos entre

individuos del sistema clonal, también juega un papel en los mecanismos defensivos de

las plantas frente a la herbivoría, al permitir la transmisión de señales que promueven

respuestas defensivas inducidas (Chen et al. 2011; Stuefer et al. 2004). Este tipo de

respuestas defensivas inducidas son beneficiosas para las plantas, ya que evitan la

producción innecesaria de compuestos químicos defensivos, los cuales suponen un gasto

considerable de recursos (Agrawal 2000; Karban & Baldwin 1989). En este capítulo se

llevó a cabo un experimento con la planta invasora Alternanthera philoxeroides, que fue

sometida a herbivoría por parte de un depredador especialista que daña las hojas de la

planta (Agasicles hygrophila Selman and Vogt), un depredador generalista que se

alimenta de savia (Planococcus minor Maskell) y tres tratamientos de herbivoría

simulada. Los tratamientos de herbivoría simulada implicaban la eliminación de tejido

foliar de las plantas, la aplicación exógena de ácido jasmónico y la aplicación conjunta

de ambos. El ácido jasmónico es una fitohormona que juega un importante papel como

molécula señalizadora en procesos defensivos (Cipollini & Sipe 2001; van Kleunen et al.

2004). Los diferentes tratamientos fueron aplicados en la parte apical de un sistema

clonal, que estaba integrado fisiológicamente con la parte basal, la cual no sufrió ningún

tipo de daño. Para testar el efecto de la integración clonal se incluyó un control en la que

los rametos basales y apicales permanecieron desconectados. De esta forma, los objetivos

del experimento eran estudiar los mecanismos de la respuesta defensiva de la planta frente

a la herbivoría y el papel que la integración clonal juega en esa respuesta. El experimento

fue realizado durante una estancia de investigación en la Beijing Forestry University


33

(BFU) en Beijing (China) en el año 2016. Se trató de un experimento en invernadero con

una duración de cinco semanas (la duración estuvo limitada por los daños provocados por

el depredador especialista, el cual causa daños masivos en hojas y tallos).

Al analizar los compuestos químicos defensivos producidos por las plantas, no se

encontró un incremento en fenoles o taninos entre los tratamientos de herbivoría y el

control. Los daños en las hojas producidos por el depredador especialista provocaron en

las plantas la misma respuesta que el tratamiento de herbivoría simulada que implicaba

daños en las hojas junto con la aplicación del ácido jasmónico. Esta respuesta consistió

en un aumento de la biomasa de raíces en la parte basal de los sistemas clonales y solo

fue posible cuando la integración fisiológica se mantuvo a lo largo del experimento.

Además, dicha respuesta fue positiva para las plantas en el caso del tratamiento de

herbivoría simulada, aunque los daños producidos por el depredador fueron tan masivos

que no se apreció ningún beneficio en ese tratamiento. Así, los resultados del experimento

muestran que el ácido jasmónico juega un papel en la respuesta compensatoria de A.

philoxeroides a la herbivoría, y que esta no consiste en la producción de compuestos

químicos defensivos, sino en un cambio no local en la distribución de la biomasa,

orientado a compensar las pérdidas de superficie foliar en los rametos afectados por la

herbivoría. Este experimento destaca la importancia de la integración fisiológica en las

respuestas defensivas frente a la herbivoría por parte de las plantas clonales.

Publicado en: Portela, R., Dong, B. C., Yu, F. H., Barreiro, R., & Roiloa, S. R. (2019).

Effects of physiological integration on defense strategies against herbivory by the clonal

plant Alternanthera philoxeroides. Journal of Plant Ecology 12(4), 662-672. doi:

10.1093/jpe/rtz004

Capítulo VI

Uno de los efectos negativos de la reproducción clonal es la falta de variabilidad

genética entre los individuos. Así, por ejemplo, se ha descrito que apenas existe

variabilidad entre las diferentes poblaciones de A. philoxeroides en China, pese a que la

planta lleva más de 50 años presente en el país (Wang et al. 2005). Para compensar esta

falta de variabilidad genética, se ha propuesto que las modificaciones epigenéticas del

ADN podrían jugar un papel importante en la adaptación de las plantas clonales a cambios

ambientales (Gao et al. 2010; González et al. 2017). En este capítulo se llevó a cabo un


34

experimento para comprobar el papel que juegan los mecanismos epigenéticos en la

plasticidad fenotípica intergeneracional de la planta invasora A. philoxeroides. El

experimento fue realizado durante una estancia de investigación en la Universidade

Federal de São Carlos (UFSCar) en São Carlos (Brasil) en el año 2017. Se emplearon dos

poblaciones del área nativa de la especie (Brasil) y una población de área no nativa

procedente de Fisterra, A Coruña (Península Ibérica), donde esta planta invasora ha sido

introducida recientemente (Romero & Amigo 2015). El experimento comprendió dos

generaciones de las plantas, obtenidas de forma vegetativa. Así, en la primera generación

la mitad de las plantas crecieron en condiciones de altos nutrientes y la otra mitad en

condiciones de bajos nutrientes, mientras que todas las plantas de la segunda generación

crecieron en condiciones de altos nutrientes. Para estudiar los mecanismos epigenéticos

implicados en la transmisión de la plasticidad fenotípica, se aplicó un agente demetilante

en las plantas de la generación parental, la 5-azacytidina. Se trata de un compuesto que

elimina la metilación en el ADN, un mecanismo implicado en la transmisión epigenética

de caracteres en A. philoxeroides (Gao et al. 2010). El experimento fue realizado en un

invernadero de la UFSCar y tuvo una duración total de 28 semanas. Los resultados

mostraron que las condiciones ambientales experimentadas por las plantas de la primera

generación afectaron a las plantas de la segunda generación, lo que se conoce como efecto

transgeneracional. Curiosamente, en las poblaciones de la zona de distribución nativa se

vieron afectadas las variables asociadas al crecimiento (número de rametos, biomasa del

tallo, biomasa radicular y biomasa total), mientras que en la población de la zona de

distribución no nativa se vio alterada a la distribución de la biomasa entre estructuras

aéreas y subterráneas. El efecto transgeneracional observado en las poblaciones de la

zona de distribución nativa puede ser debido a un efecto de "cuchara de plata" (es decir,

una ventaja debida al acceso a mejores recursos durante una etapa temprana del desarrollo

del sistema clonal), mientras que los cambios observados en las plantas procedentes de la

zona de distribución no nativa parecen estar regulados por la metilación del ADN. Este

experimento destaca la importancia de los efectos transgeneracionales en el crecimiento

de una planta clonal invasora, lo que podría ayudar a entender los mecanismos

subyacentes a su invasividad.

Publicado en: Portela, R., Dong, B. C., Yu, F. H., Barreiro, R., Roiloa, S. R., & Silva

Matos, D. M. (2019). Trans-generational effects in the clonal invader Alternanthera

philoxeroides. Journal of Plant Ecology. doi: 10.1093/jpe/rtz04

23

Capítulo VII

En este capítulo final de la tesis se ha elaborado un modelo de simulación

dinámico a partir de una revisión bibliográfica sobre las características biológicas de la

planta A. philoxeroides y el insecto A. hygrophila, depredador especialista empleado

como biocontrol de esta planta invasora en diferentes países (Buckingham 1996; Lin et

al. 1984; Sainty et al. 1998). El éxito de los diversos programas de biocontrol ha sido

mixto (con la erradicación exitosa de la planta en el sur de EEUU pero no en China o

Australia, por ejemplo), dependiendo principalmente de la temperatura de la región, ya

que tanto la planta como el insecto son nativos de climas tropicales y no toleran los

inviernos fríos. Sin embargo, mientras que la planta tiene la capacidad de rebrotar a partir

de sus raíces al año siguiente, el insecto desaparece por completo. La interacción entre

planta e insecto es compleja, puesto que el insecto no solo se alimenta exclusivamente de

la planta, sino que deposita sus huevos en las hojas y realiza la fase de pupa en el interior

de los tallos huecos. Por lo tanto, en caso de desaparecer la planta, el insecto seguirá el

mismo destino, eliminando el riesgo de que cause a su vez una invasión biológica. El

modelo elaborado permite, por una parte, simular el desarrollo de la población de A.

philoxeroides localizada en Fisterra, A Coruña (Península Ibérica) a lo largo de un

periodo de 10 años. También permite simular la aplicación de diversos tratamientos de

biocontrol, pudiendo seleccionarse la época del año en la que se realiza la liberación del

insecto, el número de insectos liberados o el número de liberaciones que se realizan

durante el periodo de simulación, para determinar la estrategia óptima. Este modelo es

aplicable a otras poblaciones de A. philoxeroides, con la única condición de conocer las

temperaturas mínimas y máximas locales a lo largo del año, así como la extensión

ocupada por la planta. Las hipótesis del trabajo eran la posibilidad de supervivencia del

insecto en la localidad de Fisterra y la posibilidad de erradicar completamente la

población de A. philoxeroides. Las simulaciones con el modelo obtenido muestran que el

biocontrol de la planta en Fisterra podría ser eficaz si el insecto es liberado con frecuencia

suficiente a lo largo de varios años, alcanzándose la erradicación total de la población, y

que la supervivencia del insecto sería posible si la cantidad de individuos liberada al

comienzo del año es lo suficientemente grande. Este trabajo fue realizado durante una

estancia de investigación en la Universidade de Trás-os-Montes e Alto Douro (UTAD)

en Vila Real (Portugal), en el año 2018.


24

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41


1. General concepts about biological invasions

Human activities allow species to overcome natural barriers that prevent their

expansion (e.g. mountain ranges, deserts, or large bodies of water), reaching regions

where they were not previously present. Although the transport of species due to human

action has been occurring since the beginning of agriculture and animal breeding, this

trend has accelerated enormously during the last centuries due to globalization (McNeely

2001; Meyerson & Mooney 2007). Those species introduced by human action outside its

natural distribution range are considered exotic species, also named introduced or non-

native species (Colautti & MacIsaac 2004). Most of these introductions occur

accidentally, although there are also cases of intentional introductions (e.g. species with

ornamental purposes or with economical use). This applies to adult individuals as well

as seeds or any part of an organism with the capacity to survive and reproduce, what is

known as a propagule. Once the exotic species has been introduced, its survival depends

to a great extent on its proximity to altered areas, where native species are subjected to

stress conditions (Keeley 2004; Richardson et al. 1992). At this point the exotic species

is called sub-sporadic or adventive species, and the population tends to disappear (Frank

& McCoy 1995). The arrival of new individuals to the population, known as propagule

pressure, is crucial for it to survive during this stage (Lockwood et al. 2005). Propagule

pressure is not only a required step in the arrival of a species to a new geographical area,

but it has been described as an important factor that contributes to the invasiveness of the

species, by increasing the genetic variability of the populations (Simberloff 2009). If the

population reaches a sufficient size to survive without depending on the arrival of new

individuals, it achieves the status of naturalized species (Richardson et al. 2000).

Once stablished in the new area, if the species is not only able to self-sustain over

several life cycles, but also to produce reproductive offspring in large numbers and at a

considerable distance from the parental population/site of introduction, it is considered

an invasive species (Richardson et al. 2011; Wilcox & Turpin 2009). Invasive species

have the ability to displace local species wherever they arrive, thus altering the ecosystem

structure. According to some authors, a plant species is considered invasive if in less than


42

50 years it has displaced 100 meters from the original population, if its reproduction is by

seeds; or more than 6 meters in 3 years, if it has vegetative reproduction through roots,

rhizomes or stolons (Richardson et al. 2000). However, according to other authors, an

impact of the exotic species on native biological diversity is required before it can be

considered invasive (IUCN 2000). According to the tens rule, approximately 10% of

species successfully take consecutive steps of the invasion process: about 10% of species

transported beyond their native range will be released or escape in the wild (they reach

the status of adventive species); about 10% of these introduced species will be able to

establish themselves in the wild (naturalized species); and about 10% of the established

species will become invasive (Williamson & Fitter 1996). However, it is enough for one

or two invasive species to be present in a habitat to seriously affect it (Frankel et al. 1995).

The attributes that favor the ability of a plant to become invasive are directly related to

its ability to reproduce, to grow rapidly from its germination to the reproductive stage,

and its adaptability to the environmental conditions of the new habitat (Richards et al.

2006; van Kleunen et al. 2010b). Biological invasions consist of a sequential process of

introduction, establishment and expansion of exotic species in geographical areas where

they were previously not present, culminating with the displacement of native species

(Fig. 1) (Blackburn et al. 2011; Vilà et al. 2008).

There are several reasons why biological invasions are problematic. The main

effect of the proliferation of invasive species in an ecosystem is the displacement of native

species, particularly endemic species that have low competitive capacity, with the

consequent risk of extinction (Cronck & Fuller 2001; Elton 1958). Island habitats are

especially sensitive to biological invasions, since evolutionary processes on islands occur

mainly under intraspecific competition, with native species susceptible to be outcompeted

by introduced species (Fernández-Palacios 2004; Lowe et al. 2004). Besides inducing the

displacement and extinction of native species, some invasive species are able to alter the

geomorphological characteristics of the habitat (Hilton et al. 2005), as well as its water

regime (Weber 2004), and incorporate allelopathic compounds into the soil (Callaway &

Aschehoug 2000). The damage produced in the ecosystems can be irreversible and have

serious economic consequences (Mack et al. 2000; McNeely 2001). In addition to causing

a degradation of landscape value, exotic plants that proliferate on agricultural land reduce

crop yields (Mehmood et al. 2017). It has also been pointed out that some exotic species

may increase the incidence of allergic responses (Laaidi et al. 2003; Weber 2004). The


43

costs associated with the control and elimination of invasive species, which in some cases

are so high that the total eradication of populations is unaffordable, should also be taken

into account (Pimentel et al. 2005). For these reasons, legislation on invasive species

around the world highlights the importance of prevention and early management of

biological invasions (e.g. EU Regulation 1143/2014) (Pallewatta et al. 2002).

Figure 1. Schematic model of the different phases of a biological invasion. Modified from Roiloa et al.

2015 with permission from the author.

2. Unraveling the causes of biological invasions

The mechanisms by which biological invasions occur have been studied

extensively during the last decades, although a universal explanation of the causes of this

phenomenon has not yet been found (Catford et al. 2009). Habitat disturbance, either

natural or caused by human action, favors the proliferation of invasive species in the


44

ecosystem (Alpert et al. 2000; Rose & Hermanutz 2004). Traditionally, two different

approaches have been used for the study of biological invasions: the study of the

invasiveness of species (those characteristics that allow them to overcome each of the

barriers of the invasion process) and the invasiveness of ecosystems (the properties of an

ecosystem that determine its vulnerability to biological invasions) (Lonsdale 1999;

Richardson & Pyšek 2006). Some factors that have been suggested as potential

determinants of invasiveness are the introduction history, species traits and ecological

and evolutionary processes that occur after the introduction (van Kleunen et al. 2010b).

The most widely tested and accepted mechanism for explaining biological

invasions is the enemy release hypothesis (ERH), which states that plant species, once

introduced outside their native distribution range, experience a decrease in herbivory and

pathogen pressure (Keane & Crawley 2002). Subsequent events are described by the shift

defense hypothesis (SDH) and by the evolution of increased competitive ability

hypothesis (EICA). According to EICA hypothesis, selection will favor genotypes with

improved competitive abilities in the absence of predators, thus increasing vegetative

growth or reproductive efforts depending on which is more important for success in a

particular new environment (Blossey & Nötzold 1995). However, if the plant has less

competitors in the invasive range and competitive ability involves traits that have a fitness

cost, then selection might act against it (Bossdorf et al. 2004). The SDH was proposed as

a more realistic explanation than EICA about the partial release from herbivores in the

invasive range. (Müller-Schärer et al. 2004). The SDH predicts that the level of defenses

against specialists will decrease as the level of defenses against generalists increase.

Qualitative toxins (e.g. terpenes or alkaloids), occurring in relatively low quantities, act

mainly against generalist herbivores, while quantitative compounds (e.g. tannins and

phenols), occurring in higher concentrations, act against specialist herbivores (Parker &

Hay 2005; Rhoades & Cates 1976). It is expected that the concentration of qualitative

defenses will increase in plants of the introduced range, while quantitative defenses with

high costs of production will decrease (Doorduin & Vrieling 2011; Joshi & Vrieling

2005). Besides, after a process of adaptive selection favoring plants that spend limited

resources on defenses, specialized herbivores are expected to show improved actuation

against the plants of the introduced populations (Blossey & Nötzold 1995; Joshi &

Vrieling 2005), which can be useful in planning biological control strategies. In spite of

these important interactions, the role of native herbivore species in shaping plant


45

invasiveness has been greatly overlooked. Different works have found results that support

both the ERH and EICA hypotheses (Daehler & Strong 1997; DeWalt et al. 2004; Manea

et al. 2019; Uesugi & Kessler 2013). Another proposed explanation for the abundance of

invasive species in their new habitat is the novel weapons hypothesis (NWH), according

to which some invasive species possess biochemical compounds that are not previously

present in the habitat where they are introduced, hence acting as biochemical weapons

(i.e. allelopathic compounds) which alter plant-predator relationships, or plant–soil

microbial interactions, giving a great advantage to the invasive plant (Callaway &

Ridenour 2004). However, little evidence of this hypothesis has been found so far (Lind

& Parker 2010).

In addition to those hypotheses that explain biological invasions by focusing on

the invasiveness of exotic species, others focus on the invasibility of habitats. Habitat

disturbance, either natural or caused by human action, favors the proliferation of invasive

species in the ecosystem (Alpert et al. 2000; Rose & Hermanutz 2004). Moreover,

according to the invasional meltdown hypothesis (IMH), the presence of invasive

species in an ecosystem facilitates the establishment of other exotic species (Mack 2003;

Simberloff & Holle 1999). In this sense, the biotic resistance hypothesis (BRH), also

known as diversity-invasibility hypothesis, describes a correlation between biological

diversity and invasiveness, with the most diverse ecosystems also being those most

resistant to invasions (Levine et al. 2004). Related to the BRH is the vacant niche

hypothesis (VNH), according to which exotic species can be successfully established in

those communities that have a vacant niche, as there are resources that are not being

exploited (Hierro et al. 2005). Also in this line has been proposed the island

susceptibility hypothesis (ISH), which describes how island ecosystems are more

susceptible to suffering a biological invasion, generally having less diversity than

continental ecosystems (Pyšek & Richardson 2006). Finally, the fluctuating resource

availability hypothesis (FRH) predicts that a temporal fluctuation in the resources of an

ecosystem may make it susceptible to biological invasion, if sufficient propagule pressure

coincides with the scarcity period (Colautti et al. 2006; Richardson & Pyšek 2006).


46

3. The role of clonal growth in biological invasions

A common feature of many plant species, that has been associated with biological

invasions, is asexual reproduction. Asexual plant reproduction is ubiquitous, occurring in

many different taxonomic groups (Klimeš et al. 1997; Tiffney & Nicklas 1985). Asexual

reproduction is also known as clonal reproduction, clonal growth or vegetative

propagation. It consists in the production of an indeterminate number of genetically

identical descendants, called ramets, arranged at more or less regular intervals on

modified stems that grow over or under the surface of the soil (stolons or rhizomes,

respectively). Once established, ramets can survive independently, or remain connected

to the clonal system for an indefinite amount of time (Price & Marshall 1999). The

interval in which the connections between ramets remain functional varies considerably

between species and is affected by environmental conditions (Jónsdóttir & Watson 1997).

In some cases the connection ceases to be functional immediately after the production of

the new ramet (Jónsdóttir & Watson 1997), while in others it can be maintained for years

(Eriksson & Jerling 1990). Stolons and rhizomes also play an important role as a reservoir

of water and carbohydrates, favoring the survival of plants in situations of stress, or in

case of fragmentation of the clonal system (Goulas et al. 2001; Stuefer & Huber 1999;

Suzuki & Stuefer 1999). Rhizomes are better suited for the storage of resources than

stolons, and also have a greater longevity, being able to remain connected to the clonal

system long after its corresponding aerial structures have disappeared (Stuefer 1998).

Clonal growth has been pointed as a characteristic that could contribute to plant

invasiveness (Liu et al. 2006; Pyšek 1997; Roiloa et al. 2015; Song et al. 2013). This idea

is based in the rationality that many of the most successful invasive plant species present

clonal propagation. In this sense, 67% of the most aggressive exotic species in Europe,

47% in North America, 54% in South America and 51% in Australia show clonal growth

(Pyšek 1997). This has been endorsed by posterior studies, in which it has been

determined that clonality is positively correlated with the invasiveness of clonal species

in different regions (Liu et al. 2006; Shah et al. 2014). Recent studies have showed the

importance of clonal integration in the expansion clonal invaders (Otfinowski & Kenkel

2008; Roiloa et al. 2014a, 2014b, 2016; Wang et al. 2008; Yu et al. 2009). A meta-

analysis conducted by Song et al. (2013) found a correlation between invasiveness of

different exotic species and the benefit of clonal integration for recipient ramets growing

in unfavorable conditions, highlighting the relationship between clonal integration and


47

invasiveness (Song et al. 2013). However, this results were not reflected as a benefit for

the whole clone. On the other hand, the role of stolons and rhizomes as storage organs

could play a crucial role in the colonization of new environments by invasive clonal

species, most notably after a process of fragmentation (Dong et al. 2010, 2011, 2012;

Konlechner et al. 2016; Lin et al. 2012).

Most of the species that present clonal reproduction can also resort to sexual

reproduction (Yang & Kim 2016). Both types of reproduction have different advantages

and drawbacks (Lei 2010). Thus, the most obvious advantage of clonal reproduction is

that it can be carried out from a single individual, which can eventually lead to a

population of clones, what is known as a genet (van der Merwe et al. 2010). The downside

of this is that there is hardly any genetic variability among the individuals originated by

clonal reproduction, whereas sexual reproduction implies the genetic recombination of

both parents, giving rise to individuals with unique combinations of genes (Bürger 1999).

This allows the action of natural selection at the level of single individuals, not entire

populations, which favors the adaptation and survival of populations. When an individual

is well adapted to the characteristics of their environment, clonal reproduction is

preferable to sexual reproduction, since the descendants will also be well adapted to that

environment (Otto 2009). Accordingly, the expenditure of resources in sexual organ

production (flowers and fruits), which entail a significant cost for plants, is avoided (Roze

2012). On the other hand, sexual reproduction permits the offspring of maladapted

individuals in stressful environments to acquire adaptive alleles (Griffiths & Bonser

2013). Finally, it should be taken into account that sexual reproduction allows the

dispersal of new individuals by seeds, which can cross long distances and are adapted to

survive unfavorable conditions, while clonal reproduction is usually done by stems or

roots, which gives the plant a fairly limited dispersal capacity (Vittoz & Engler 2007; von

der Lippe & Kowarik 2007).

As long as the different ramets of a clonal system remain connected, they maintain

the ability to exchange resources and other compounds through their stolons or rhizomes,

due to a process known as physiological integration or clonal integration (Fig. 2) (Price

& Marshall 1999). This exchange of substances include water, nutrients, hormones or

photoassimilates, as well as defensive signals (Stuefer et al. 2004). This capacity of

integration between ramets improves the survival of the whole clonal system, by

providing resources to those ramets growing in unfavorable conditions or subjected to


48

stress (Hartnett & Bazzaz 1983; Roiloa & Retuerto 2006; Saitoh et al. 2002). Clonal

integration allows the survival of plants in highly disturbed environments, and even their

establishment in areas that would be unsuitable for a non-clonal plant (Alpert & Mooney

1986). In addition, the subsidy of resources to apical ramets, bypassing the need of

biomass allocation to the development of roots, allows the rapid expansion of the clonal

system, thus efficiently colonizing the surrounding space (Roiloa et al. 2010). The

relationship between physiological integration and biological invasions has been

suggested in works with different invasive species (Pyšek 1997; Song et al. 2013; Wang

et al. 2008).

Another phenomenon associated with clonal integration is the division of labor

among different individuals of the clonal system (Stuefer et al. 1996). Division of labor

comprises two aspects: the individual specialization of tasks, with the consequent

improvement in the performance of some tasks to the detriment of others, and the

cooperation between potentially independent units of a modular system, consisting in the

exchange of resources among them (Fig. 2) (Alpert & Stuefer 1997). Two types of

division of labor have been described in clonal plants: environmentally-induced division

of labor, which is a plastic response to uneven access to resources by different individuals

of the clonal system; and developmentally-programmed division of labor among

individuals with different ontogeny (Stuefer 1998). Environmentally-induced division of

labor is especially advantageous in environments with heterogeneous distribution of

resources (Hutchings & Wijesinghe 1997; Roiloa et al. 2007). The specialization of

ramets in obtaining those resources to which they have easy access allows an optimal

distribution of the biomass along the clonal system (Stuefer et al. 1998). There is a certain

similarity between environmentally-induced division of work in clonal plants and

economic models, since in both cases the goal is to maximize the benefits. According to

the principle of supply and demand, production is maximized if resources are taken where

they are most abundant (i.e., resources can be acquired at the lowest cost) and used where

they are most scarce (Bloom et al. 1985; Rapport & Turner 1977). On the other hand,

developmentally-programmed division of labor has been reported in species which grow

in open environments in which light is not a limiting resource, but nutrients are, like sand

dunes (e.g. Carex bigelowii and C. arenaria) (Jónsdóttir & Watson 1997). Thus, by the

establishment of multiple rooting ramets over a large area, plants increase the number of

sampling points for the critical resource. Also, since light is not a limiting factor in open


49

habitats, a few specialized modules are sufficient to provide the clonal system with

photoassimilates. Overall, both types of intra-clonal division of labor grant benefits in

terms of fitness-related traits, such as higher biomass and clonal offspring.

Figure 2. Schematic representation of the mechanisms of physiological integration (left) and division of

labor (right). Modified from Roiloa et al. 2015 with permission from the author.

4. Overcoming the disadvantages of clonal growth

When a species is introduced outside its natural habitat, it must adapt to new

environmental conditions (Pérez et al. 2006). This is particularly important in clonal

species, due to the abovementioned lack of genetic variability. Therefore, phenotypic

plasticity is a characteristic that could play a key role in the biological invasions of clonal

plants (Davidson et al. 2011; Keser et al. 2014; Richards et al. 2006). Phenotypic

plasticity is the ability of a single genotype to produce different phenotypes depending on

the characteristics of the environment (Bradshaw 1965; Sultan 2000). A widely described

case of phenotypic plasticity is the allocation of biomass in those structures responsible

for obtaining the most limiting resource for plant growth (Gleeson & Tilman 1992;

Hilbert 1990; Weiner 2004). Namely, low levels of belowground resources (nutrients and


50

water) will result in a proportional increase of root biomass, whereas low light levels will

induce a decrease in the proportional root growth and increasing the proportional

production of leaves. When plasticity allows greater biological efficacy across different

environments, it is expected to be favored by natural selection (Fusco & Minelli 2010;

Ghalambor et al. 2007). Comparison of the phenotypic plasticity of individuals of

invasive species from both the native area and the invaded area has been described as an

adequate approach to understand the role that this trait plays in biological invasions (van

Kleunen et al. 2010a). Also, the comparison of plasticity between different invasive

species and non-invasive congeners has shown a greater plastic response of the invasive

species towards the water availability, different nutrient levels and different

concentrations of CO2, which reinforces the notion that this trait significantly contributes

to plant invasiveness (Funk 2008; Geng et al. 2006; Raizada et al. 2009; Wei et al. 2017).

Another factor that can compensate for the lack of genetic variability in clonal

plants is epigenetic regulation of DNA. It consists on a variety of chemical modifications

in the DNA that alter its expression, but not the nucleotide sequence (Gao et al. 2010;

Wolffe & Matzke 1999). These modifications can increase the expressivity of a gene or

silence it completely, preventing it from being expressed. Different studies have indicated

that the influence of environmental conditions on phenotypes could be mediated by this

mechanisms (Bossdorf et al. 2010; Hallgrímsson & Hall 2011; Herman & Sultan, 2016;

Verhoeven & Preite 2014). Epigenetic DNA regulation comprises reversible, inheritable

modifications that can be altered more flexibly than the genome sequence (Heard &

Martienssen 2014; Martienssen & Colot 2001). An epigenetic modification of DNA that

has been widely documented in various species is the methylation of cytosine (and, less

extensively, of adenine) (Bender 2004; Bossdorf et al. 2008). When methylation occurs

in the promoter segment of a gene, it inhibits transcription thereof. Thus, epigenetic

mechanisms give those species with asexual reproduction an efficient alternative to

genetic recombination as a source of variability. It has even been suggested that

epigenetic regulations in gene expression would allow the establishment of invaders in

the short term (Pérez et al. 2006). Recent studies have found a correlation between

epigenetic and phenotypic variation of aggressive clonal invaders with low genetic

variability, indicating that epigenetic mechanisms are an alternative source of variability

(Gao et al. 2010; Wang et al. 2019). However, the role of trans-generational effects in the

invasiveness of clonal species has been generally overlooked.


51

5. Objectives

Understanding the mechanisms involved in biological invasions is key to predict

future invasion scenarios and to design efficient strategies for the control and restoration

of invaded areas. The general objective of this doctoral thesis is to contribute to determine

the role that clonal plant growth, and different attributes associated with it, play in

biological invasions. With this aim, a series of experiments were carried out in which the

benefit of various characteristics associated with clonal reproduction was tested in two

invasive species, Carpobrotus edulis (L.) N. E. Br. and Alternanthera philoxeroides

(Mart.) Griseb. Both species present clonal reproduction through stolons and are

considered aggressive invaders, since they cause serious alterations in the ecosystems

they invade. In order to determine the existence of adaptive selection processes

throughout the invasion, some of these experiments were conducted comparatively in

populations from the native and invaded areas. With this aim, a field experiment was

conducted in a coastal dune to test the benefits of clonal integration in C. edulis in

situations of partial burial due to sand (Chapter I), and another field experiment was

performed to determine the benefits of clonal integration in C. edulis and their co-

occurring congener Carpobrotus acinaciformis (L.) L. Bolus when growing in conditions

of heterogeneous distribution of nutrients (Chapter II). An experiment was also carried

out with populations of native (South Africa) and invaded (Iberian Peninsula) ranges of

C. edulis to study the phenotypic plasticity shown by individuals when subjected to

conditions of scarcity of water, light or nutrients (Chapter III). The objective of this work

was to test if patterns in biomass partitioning in response to resource shortages differ

between populations from the native and invaded range. Continuing this line of research,

another experiment was performed with four different populations of C. edulis (South

Africa, Iberian Peninsula, California and Australia), as well as other congener species

differing in their invasiveness status (C. acinaciformis from the Iberian Peninsula,

Carpobrotus chilensis (Molina) N. E. Br. from California and Carpobrotus virescens

(Haw.) Schwantes from Australia). The objective in this experiment was to compare the

biomass allocation pattern of Carpobrotus spp. in response to nutrient scarcity between

native and invaded ranges, as well as between congeners with different degrees of

invasiveness. The same approach was also used to study the competitive ability of

Carpobrotus spp. (Chapter IV). As for A. philoxeroides, an experiment with real and

simulated herbivory treatments was carried out to test the role that clonal integration plays


52

in the defensive response of the plant, as well as the role of jasmonic acid in the induced

defensive mechanisms of this species (Chapter V). On the other hand, an experiment was

carried out to study the trans-generational effect of phenotypic plasticity response to

nutrient scarcity in populations of A. philoxeroides from their native (Brazil) and invasive

(Iberian Peninsula) range, and whether this effect is mediated by an epigenetic regulatory

mechanism, DNA demethylation (Chapter VI). Finally, a dynamic simulation model was

elaborated recreating the life cycle of A. philoxeroides, the prey-predator relationship

with a biocontrol agent of this invasive species (Agasicles hygrophila Selman & Vogt)

and the optimization of a cost-effective biocontrol for the elimination of the plant in a

population located in Fisterra (NW Spain) (Chapter VII).

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Section I

Carpobrotus spp.

Carpobrotus spp.

61

1. Biology and native distribution

The genus Carpobrotus (Phylum Magnoliophyta, Class Magnoliopsida, Order

Caryophyllales, Family Aizoaceae, Subfamily Ruschioideae) comprises around 20-25

species of succulent perennial plants, native to South America (Chile), South Africa, New

Zealand and Australia (Campoy et al. 2018; Wisura & Glen 1993). The different species

of the genus present clonal growth through prostrated rhizomes, 3-angled leaves, solitary

flowers on erect stems (whose colors are characteristic for each species, varying between

white, yellow, pink or purple) and indehiscent fleshy fruits (Hartmann 2002). Plants of

this genus grow mainly in coastal areas with Mediterranean or temperate climates, both

in sand dunes and in rock cliffs, although they can also be found in interior areas with

sandy substrates (Hartmann 2002). They are well adapted to dry environments and are

resistant to fire, since they accumulate water in their tissues (Pierce 1994). In conditions

of stress due to drought or salinity, Carpobrotus spp. can induce crassulacean acid

metabolism (CAM), resulting in a high water use efficiency by uptaking CO2 trough

stomata at night, thus reducing water loss via transpiration (Earnshaw et al. 1987).

Carpobrotus edulis (L.) N.E. Br and Carpobrotus acinaciformis (L.) L. Bolus are

two species of the genus Carpobrotus, native to the Cape Region in South Africa (Wisura

& Glen 1993). Both species develop an extensive monopodial system and present radial

growth with a structure of nodes and internodes, forming dense mats by the production

of apical ramets. Individuals within the clonal system remain physiologically integrated

by stolon connections, allowing the plants to spread and colonize the surrounding area

(Roiloa et al. 2010; Wisura & Glen, 1993). Both species are monoecious, having

hermaphrodite flowers. Petals are 25-30mm long and flowers are 45-55mm wide (Fig. 1).

According to Wisura and Glen (1993), C. edulis is the only species of the genus with

yellow flowers, while C. acinaciformis has pink flowers (Fig. 2). However, since flowers

of C. edulis may vary on its color (from yellow to pink), leaf equilaterality has been

suggested as a useful trait for discerning Carpobrotus taxa. C. edulis presents an

equilateral leaf-cross section, while C. acinaciformis has an isosceles leaf cross-section

(Gonçalves 1990). The leaves of C. edulis are 6-13 cm long, straight or very slightly

curved, with rough brown teeth along the bottom ridge (Keighery 2014), while leaves of

C. acinaciformis are curved (Wisura & Glen 1993).

Section I

62

Figure 1. Reproduction of herbarium sheet of Carpobrotus edulis. Royal Botanical Garden of Madrid.

Caption: (a) habit; (b) longitudinal section of flower; (c) cross section of receptacle; (d) stamens; (e) seeds

with funicle.

It has been found that C. edulis is capable of hybridize with other species of the

genus Carpobrotus, both in the native area of the species (Wisura & Glen 1993) and in

the introduction areas (Suehs et al. 2004; Vilà et al. 1998; Waycott 2016), as well as with

a species of other genus of the family Aizoaceae, Disphyma crassifolium (L.) L. Bolus

(Chinnock 1972). The hybrid of C. edulis and C. acinaciformis is referred as C. affine

acinaciformis (Suehs et al. 2004). According to previous studies, this hybrid would have

Carpobrotus spp.

63

greater vigor than C. edulis in those areas where the plant has been introduced, showing

a higher phenotypic plasticity and germination capacity in a wide range of habitats and

environmental conditions (Podda et al. 2018; Traveset et al. 2008). Hybridization of C.

edulis with the native species C. rossii (Haw.) Schwantes in Australia and the naturalized

non-invasive species C. chilensis (Molina) N.E. Br in California involves the genetic

contamination of native populations and the alteration of their ecological functions

(Waycott 2016).

Figure 2. Detail on flowers and leaf cross-section of C. edulis (top) and C. aff. acinaciformis (bottom).

1. Invaded range

Several species of Carpobrotus have been introduced outside their native range,

notably C. edulis and C. acinaciformis, causing biological invasions in different countries

(D'Antonio 1993; Robert et al. 2013; Traveset et al. 2008). The reasons for these

introductions have been varied. On the one hand, the general appearance of Carpobrotus

Section I

64

spp. and their showy flowers make these plants fitting for ornamental purposes (Sanz-

Elorza et al. 2004). On the other hand, its ability to form a dense tapestry covering large

extensions and its rapid growth have encouraged its use to stabilize land embankments or

mobile coastal dunes (Campos et al. 2004). Moreover, its fire resistance makes

Carpobtorus spp. useful for create firebreaks around inhabited areas (Pierce, 1994),

which could be an interesting application if not for the aggressive clonal growth and

potential invasiveness of these plants outside its native range. Once introduced in the new

areas, plants expand to coastal and disturbed environments, including disused farmlands

and burnt areas (D'Antonio et al. 1993; Kuebbing et al. 2014).

The species of Carpobrotus currently naturalized in Europe are C. edulis and C.

acinaciformis, as well as their hybrid, C. affine acinaciformis (Suehs et al. 2004).

However, the unclear discerning between both species and the hybrid makes it difficult

to establish the exact identity of all populations, as well as to determine which species

correspond to each registered introduction. The first introductions of Carpobrotus spp. in

Europe go back to the seventeenth century in Belgium (where the plants did not survive)

and England (Robert et al. 2013). The plants later reached France and expanded along the

Mediterranean coast during the 19th century. The first mention in Spain occurred in the

NW, in Baiona, in 1892 (Lázaro-Ibiza 1900). Naturalized populations of Carpobrotus

spp. can be found nowadays in Germany (Washburn & Frankie 1985), France (Vilà et al.

2006), UK (Robert et al. 2013), Spain (Sanz-Elorza et al. 2004), Portugal (Marchante et

al. 2014), Italy (Carranza et al. 2010), Croatia (Stancic et al. 2008) and Greece

(Arianoutsou et al. 2010) (Fig. 4). There also exist biological invasions of C. edulis in

other countries, such as the US (D'Antonio, 1993), Australia (Collins & Scott, 1982) or

New Zealand (Chinnock, 1972) (Fig. 3).

2. Invasiveness

The main causes of the invasiveness of Carpobrotus spp. are its rapid growth, its

ability to outcompete native species and its morphological plasticity (D'Antonio & Mahall

1991; Traveset et al. 2008). Both C. edulis and C. acinaciformis are capable of performing

either sexual or asexual reproduction, in the native range as well as in the introduced

areas. Both species have fleshy fruits and their seeds are dispersed by birds or small

mammals (Novoa et al. 2012). However, it is clonal reproduction that gives a huge

Figure 3. Distribution map of C. edulis worldwide. Map obtained from GBIF in May, 2019.

Figure 4. Distribution map of C. edulis in Europe. Map obtained from GBIF in May, 2019.

Carpobrotus spp.

67

advantage over native species. Carpobrotus spp. form dense mats in which no other

species is able to sprout, thus acquiring all the light and nutrients of the area occupied by

the clonal system (Campos et al. 2004; Maltez-Mouro et al. 2010). Clonal growth allows

Carpobrotus spp. to grow hanging over rock cliffs, or in mobile sand dunes, preventing

the plants from being completely buried by sand displacement due to wind.

In addition, physiological integration trough stolons allows the transport of

resources, mainly water and nutrients, among the different individuals of the clonal

system (Pitelka & Ashmun 1985; Price & Marshall 1999). This is particularly useful in

coastal dunes, where the nutrients are distributed according to a heterogeneous spatial

gradient and the amount of water is limited (Alpert & Mooney 1996). Furthermore,

division of labor among members of the clonal system has been described in Carpobrotus

spp. (Portela & Roiloa 2017; Roiloa et al. 2014). This trait allows older ramets, located

in the basal position of the clonal systems, to specialize in the acquisition of water and

nutrients, so that the apical branches allocate its biomass to the production of aerial

structures, thus enhancing the expansion of the clonal system. Previous works have found

an adaptive selection of this trait during the invasion process, suggesting that it

contributes to the invasiveness of C. edulis (Roiloa et al. 2016).

3. Ecological impact

Carpobrotus spp. are capable of altering in their favor the biotic and abiotic

conditions of their environment (Molinari et al. 2007). As aforementioned, they

successfully outcompete native species, altering the structure of ecosystems (Fig. 5). Both

species richness and diversity diminish due to invasions by Carpobrotus spp.

(Badalamenti et al. 2016; Jucker et al. 2013). This also facilitates the entrance of other

ruderal plants into the habitat (Santoro et al. 2012). In addition, Carpobrotus spp.

negatively affects the pollination of native species, because its huge and colorful flowers

attract pollinators (Vilà et al. 2009). These habitat disturbances can, in turn, affect the

associated animal communities (Galán 2008). Carpobrotus spp. invasions have the

potential to alter soil properties, varying the concentration of organic matter and the pH

of the substrate mainly due to the accumulation of necromass in a substantially greater

amount than that produced by native species (Novoa et al. 2014; Santoro et al. 2011). The

variation of the pH affects the availability of nutrients, since soil acidification leads to a

Section I

68

reduction in the availability of Ca and Mg, as well as to the inhibition of nitrification

(Conser & Connor 2009). This change in soil characteristics can be maintained for several

years after the elimination of Carpobrotus spp., hindering the recolonization by native

species and favoring subsequent invasions by other ruderal species (Novoa et al. 2013).

Figure 5. Carpobrotus sp. invading a dune system in O Grove, Pontevedra (NW Spain) in competition with

the native species Honckenya peploides (L.) Ehrh.

4. Legal status

Due to its highly invasive condition and the danger it poses to the native flora, C.

edulis has been classified as a noxious weed in several countries, forbidden to be released

in the environment. In Europe, this is the case of Spain (Royal Decree no. 630/2013, 2nd

August), Portugal (Royal Decree no. 565/99, 21st December), UK (Schedule 9 to the

Wildlife and Countryside Act 1981), Ireland (Section 52 of The Wildlife Amendment Act

2000) and Italy (regional law of the 6th of April 2000, no. 56, in Tuscany). Also, in the

US C. edulis has been included in the list of invasive alien species by the California

Invasive Plant Council (http://www.cal-ipc.org/).

http://www.cal-ipc.org/

Carpobrotus spp.

69

5. Management

The mechanical control of Carpobrotus spp. can be performed manually, since

roots of ice plants are shallow enough to allow them to be easily pulled out of the sand

(Carta et al. 2004). Thin mats can be rolled out as a carpet, without need of using

machinery. Tractors should only be used with mats of considerably size, since the use of

machinery can disturb the whole dune and damage endangered native species. The

amount of water in the tissues of the plants increases their weight and the effort required

for its transport and disposal. One solution to this problem is to cover the shrubs with a

tarpaulin until they dry, or to dry them after they have been torn off (Chenot et al. 2018).

Precautions should be made to avoid outbreaks from the removed plant material, and it is

advisable to restore the native vegetation after the removal of Carpobrotus spp., since

they have a poor performance when growing in the shade of other vegetation (Sanz-Elorza

et al. 2004). Overall, while mechanical control methods can lead to the successful

eradication of Carpobrotus spp., they are time and manpower consuming, and require

biomass management (Carta et al. 2004). Chemical control, on the other hand, can be

successfully achieved using glyphosate (Fagúndez & Barrada 2007). A surfactant can be

useful for breaking the plant's cuticle and enhance the effect of the herbicide. However,

since glyphosate has a broad-spectrum effect, chemical control of Carpobrotus spp.

should be avoided when native plants are present.

The biological control of Carpobrotus spp. has been proposed as an alternative to

the use of herbicides and laborious mechanical control. Outside its native area,

Carpobrotus spp. have no predators that limit its expansion (Maltez-Mouro et al. 2010).

Most of the animals that feed on ice plants only ingest their fruits, spreading the seeds

(Novoa et al. 2012). The fungus Sclerotinia sclerotiorum (Phylum Ascomycota, class

Discomycetes, order Helotiales, family Sclerotiniaceae) has been proposed as a possible

biocontrol agent of Carpobrotus spp. It is a cosmopolitan fungus that infects more than

500 plant species worldwide (Saharan & Mehta 2008). Its main advantage is that it is

already present in many habitats where Carpobrotus spp. have been introduced.

Another candidate as a biocontrol agent is the cottony pigface scale, Pulvinariella

mesembryanthemi (Phylum Arthropoda, class Insecta, order Hemiptera, family

Coccidae). P. mesembryanthemi is a specialized predator, which feeds exclusively on ice

plants (Miller et al. 2005; Washburn & Frankie 1985) and has been introduced along with

the plants in many of the invaded areas (Fig, 6). In optimal conditions and with high

Section I

70

densities of Carpobrotus spp., this insect is capable of causing massive damage to the

plants (Donaldson et al. 1978). A greenhouse experiment comprising both biocontrol

agents showed that S. sclerotiorum did not affect the growth of C. edulis in the long term,

whereas P. mesembryanthemi decreased plant growth (Vieites-Blanco et al. 2019). The

joint use of both species had good results, killing half of the plants in one year and

considerably reducing the growth of the surviving plants. However, the performance of

the insect in field conditions as a biocontrol agent is still unknown.

Figure 6. Detail of P. mesembryanthemi with ovisac over C. edulis (University of A Coruña, May 2017) (left)

and S. sclerotiorum causing an impact on Phaseolus vulgaris L. (right).

Carpobrotus spp.

71

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Wisura, W., & Glen, H. F. (1993). The South African species of Carpobrotus (Mesembryanthema–

Aizoaceae). Contributions from the Bolus Herbarium, 15, 76-107.

Chapter I

Physiological integration buffers sand burial stress in the clonal plant

Carpobrotus edulis invading a coastal dune in NW Iberia

Rubén Portela1, Rodolfo Barreiro1, Sergio R. Roiloa1

1BioCost Group, Biology Department, Faculty of Science, Universidade da Coruña, A

Coruña 15071, Spain.

Chapter I

77

Abstract

Sand burial represents one of the most common stresses for plant performance in coastal

sand dunes, dramatically reducing the capture of light by leaves, and consequently

conducting to a carbon balance collapse. Physiological integration is one of the most

remarkable traits associated with clonal growth, and allows clonal plants to perform as

cooperative systems, favoring their colonization in stressful patches that could be difficult

for non-clonal plants. We conducted a field experiment in a coastal sand dune in NW

Iberia in which apical (relatively young) ramets of Carpobrotus edulis were either not

subjected to sand burial or subjected to burial in sand to a depth of 90% ramet height and

were either connected to or disconnected from basal (relatively young) ramets not

subjected to sand burial. Results supported our main hypothesis that physiological

integration benefits apical ramets that suffer from sand burial. Interestingly, our results

showed that physiological integration induced a non-local plastic response in basal ramets

that was dependent on the conditions experienced by their apical ramets. Thus, when

apical ramets remained unburied a developmentally programmed division of labor was

found, with basal ramets specialized in the acquisition of soil-based resources while apical

ramets specialized in aboveground expansion. On the contrary, when apical ramets were

subjected to sand burial, basal ramets changed their biomass allocation pattern and

increased the production of photosynthetic structures, alleviating the light stress suffered

by their connected apical ramets. Our results are pioneering in reveal that physiological

integration allows the invasive C. edulis to withstand sand burial when colonizing coastal

dunes, which can have important consequences to understand the invasive success of this

clonal species in coastal systems.

Keywords: biomass partitioning; Carpobrotus edulis; coastal sand dune; clonal

integration; division of labour; modular plasticity; sand burial.

Section I – Carpobrotus spp.

78

1. Introduction

Sand dune ecosystems are fragile habitats that contain both endemic and

endangered species, being a principal objective for biodiversity conservation (Maun

1998, 2009). However, sand dunes represent stressful environments for plant growth, and

plants on sand dunes have to develop different strategies to cope with the harsh conditions

such as low water content, shortage of essential nutrients (Maun 1998), and suffering

from wind erosion (Yu et al. 2008; Maun 2009; Fan et al. 2018). Horizontal sand

movement driven by wind frequently results in sand burial of plants inhabiting sand dunes

(Maun 1998; Yu et al. 2002, 2004; Fan et al. 2018). Such situation represents one of the

most common stresses affecting plant performance in sand dunes, dramatically reducing

the capture of light by leaves, and leading to a carbon balance collapse (Maun & Lapierre,

1984; Maun 1996, 1998; Yu et al. 2002, 2004; Fan et al. 2018).

Sand dunes are frequently occupied by plants capable of clonal growth (Maun &

Lapierre, 1984; Yu et al. 2002, 2004; Fan et al. 2018). Clonal growth is characterized by

the vegetative production of genetically identical units, named ramets. These ramets can

remain physically connected by stem internodes for a variable period of time, and are also

capable of an independent existence after disconnection (Klimes et al. 1997; Price &

Marshall 1999). This type of reproduction allows clonal plants to produce a large network

of interconnected ramets with the capacity for a fast horizontal expansion and the

susceptibility to experience environmental heterogeneity (Oborny & Cain 1997; Oborny

2019). Physiological integration via physical connection between ramets within a clonal

system allows translocation of resources and other substances among individuals. It has

been well documented for a variety of conditions and species that physiological

integration generally allows the transport of essential resources from ramets growing in

patches with high resource availability to ramets in patches with resource scarcity, or

from already stablished older ramets to developing new ramets. Thus, physiological

integration allows clonal plants to act as cooperative systems, favoring the colonization

of stressful patches where survival would be difficult for non-clonal plants or no-

integrated ramets (e.g. Hartnett & Bazzaz 1983; Slade and Hutchings 1987; Alpert 1999;

Saitoh et al. 2002; Roiloa & Retuerto 2006).

Wind frequency and intensity, as well as degree of soil moisture and presence of

vegetation cover or other physical barriers, determine a spatially heterogeneous pattern

of sand burial in coastal dunes (Maun 2009). Thus, is expected that sand burial affect

Chapter I

79

some modules within a clonal system while other parts of the clone remain unburied.

Also, events of disturbance are frequent under natural conditions, leading to the

fragmentation of clonal systems into portions of different size, which consequently

prevents physiological integration between the different fragments (Stuefer & Huber

1999). The present study was conducted under field conditions in a coastal sand dune to

determine the importance of physiological integration to buffer the negative impact of

sand burial in the invasive species Carpobrotus edulis. With this propose connected

(physiological integration is allowed) and disconnected (physiological integration is

prevented) apical ramets of C. edulis were subjected to sand burial, while the basal parts

of the clone remained unburied. While previous experiments have explicitly studied this,

as far as we know this is the first study with this aim conducted in a natural coastal sand

dune, and using as a model species an aggressive invasive plant. Specifically, we

predicted that (1) connection between ramets will allow physiological integration, and

consequently would increase plant growth in apical ramets due to the support received

from basal ramets; and that (2) the benefit of integration on plant growth would be

especially important for buried ramets, as the transport of essential resources from

unburied ramets would be essential to buffer the stressful conditions imposed by burial

on apical ramets.

2. Material and methods

2.1.Study species

Carpobrotus edulis (L.) N.E. Br (Aizoaceae), commonly known as ice plant, is

native to the Cape Region in South Africa. It is a succulent clonal species with prostrate

stems that develop into extensive systems with radial growth (Wisura & Glen 1993).

Clonal growth enables C. edulis to form dense mats by the production of offspring ramets

that remain physiologically integrated by stolon connections, allowing the plant to spread

horizontally and effectively colonize the surrounding area (Roiloa et al. 2010; Portela &

Roiloa 2017). C. edulis invades coastal ecosystems with Mediterranean-type climates

around the world, causing a negative impact on the diversity of the native flora

(D'Antonio and Mahall 1991; D'Antonio 1993; Traveset et al. 2008; Campoy et al. 2018).


80

2.2.Experimental design

Thirty-six similar-sized un-rooted clonal fragments of C. edulis were collected in

three coastal locations in Galicia (NW Iberia): Estai Cape (42º11'10''N, 8º48'55''W), O

Grove (42º28'18''N, 8º51'25''W) and A Coruña (43º22'53''N, 8º24'37''W). Locations were

at least 30 km apart, and clonal fragments within each location were collected at least 10

m apart from each other to increase genetic representation. Each fragment consisted of

the youngest four ramets along a stem, so that all fragments were at a similar size and

developmental stage. The two older ramets are hereinafter referred to as basal ramets, and

the two younger ramets are hereinafter referred to as apical ramets. Plants were

transplanted into a natural coastal sand dune system in Seselle (Ares, NW Iberia;

43º25'45''N, 08º13'37''W) where the experiment was carried out. The experiment

comprised two crossed factors: connection between basal and apical ramets (connected

vs. disconnected) and sand burial (buried vs. unburied). In the connection treatment, the

basal and apical ramets of the clonal fragments remained connected (physiological

integration was allowed). In the disconnection treatment, the stem internode connecting

the basal and apical ramets was disconnected by cutting the stem internode halfway

between the basal and apical ramets (physiological integration was impeded). In the

unburied treatment, neither the basal nor the apical ramets were buried in sand. In the

buried treatment, the basal ramets were not buried in sand, whereas the apical ramets were

buried in sand to a depth of 90% of their height. Apical ramets were not completely buried

in order to avoid their death when the connection to the basal ramets was severed. Burial

treatment was maintained throughout the experiment by periodical reconditioning of the

sand covering the plants. Both stem connection and sand burial intended to mimic natural

conditions of coastal sand dunes inhabited by C. edulis, where strong disturbance can

break the connection between ramets and intense frequent wind can bury part of the clone.

All the apical ramets were positioned towards the ocean to avoid potential bias due to

orientation effect. This arrangement mimicked the natural sand burial of apical ramets, as

the predominant wind blows from ocean to land, making apical ramets to be more

susceptible to burial by sand. Plants from different treatments were placed interspersed to

avoid confounding effects of position within the dune. Clonal fragments from each of the

three locations sampled in the field were equally represented and randomly assigned to

each combination of connection by burial treatments. The experiment was initiated on

Mach 6, 2018 and continued for 120 days.

Chapter I

81

2.3. Measurements

Ramets were weighed to measure fresh mass at the start of the experiment. Initial

fresh mass of basal and apical ramets in the connected treatment was estimated from

clonal fragments mass, using the proportional weight of basal (55.6%) and apical ramets

(44.4%) within each clonal fragment in the disconnected replicates (n = 18). At the end

of the experiment final fresh mass of the apical part and the basal part was measured. The

relative growth rate of fresh mass (RGR, g·g-1·day-1) was calculated as RGR = (lnFWt2 –

lnFWt1) / (t2 – t1), where t2- t1 is the duration (days) of the experiment, and FWt1 and FWt2

are fresh weight (g) at the start and end of the experiment respectively. RGR was

calculated separately for the basal part, apical part and whole clone fragment (basal +

apical part). Afterwards, the basal and apical parts were separated into shoot and roots,

dried at 70°C for 72 h and weighed. Total dry mass (shoot mass + root mass) and root to

shoot ratio (RSR) were calculated for the basal and apical parts separately and at the

whole clone level. Also, number of leaves was recorded at harvest for basal and apical

ramets, and calculated for the entire clone.

2.4.Statistical analysis

We used two-way ANOVA to examine the effects of stem connection and sand

burial on RGR, and two-way ANCOVA to test the effects of stem connection and sand

burial on total dry mass, shoot dry mass, root dry mass, number of leaves and RSR of the

apical part, the basal part and the whole clonal fragment (apical plus basal part) of C.

edulis. For ANCOVA, initial fresh mass of the whole clone was included as a covariate.

A posteriori Tukey tests were used for multiple comparisons. The transformation ln(x)

was applied to root mass of apical ramets and RSR of both basal and apical ramets to

meet the requirements of normality and homogeneity of variances. Mortality reduced the

number of replicates, as indicated by the error degree of freedom of the analyses.

Statistical tests were performed with IBM SPSS Statistics 23.0 (IBM Corp., Armonk, NY,

USA).


82

3. Results

3.1.Performance of apical ramets

RGR, shoot mass and total mass of the apical ramets were significantly greater

when they remained unburied than when they were buried in sand (Table 1A, Fig. 1A, D,

E). Connection to the basal ramets reduced RSR and root mass of apical ramets compared

to disconnection, and this pattern was maintained both for the burial and unburial

treatments (Table 1A, Fig. 1B, C). Connection had no effect on RGR, shoot mass or total

mass (Table 1A).

3.2.Performance of basal ramets

Neither connection to the apical ramets or sand burial significantly affected RGR,

shoot mass or total mass of the basal ramets (Table 1B, Fig. 2A, D, E). Connection to the

apical ramets reduced number of leaves of the basal ramets compared to disconnection

(Table 1B, Fig. 2F). Connection to the apical ramets significantly increased RSR and root

mass of the basal ramets when the apical ramets remained unburied, but had little impact

when the apical ramets were buried, as indicated by the significant interaction effect of

connection x burial (Table 1B, Fig. B, C).

3.3.Performance of the whole clonal fragment

Connection between the apical and the basal ramets significantly increased RGR

and decreased RSR of the whole clonal fragment (Table 1C, Fig. 3A, C). Sand burial of

the apical ramets significantly reduced RGR of the whole clonal fragment (Table 1C, Fig.

3A). Sand burial of the apical ramets reduced root mass of the clonal fragment when the

connection between the apical and basal ramet was maintained, but had no little impact

when the connection was severed, as indicated by the interaction effect of connection x

burial (Table 1C, Fig. 3C). Neither connection nor burial affected shoot mass, total mass

or number of leaves of the whole clonal fragment (Table 1C).

4. Discussion

Results supported our main hypothesis that physiological integration would report

a benefit for apical ramets in terms of growth. Also, as predicted, the benefit of integration

was especially important when apical ramets suffered burial stress. Thus, reduction of

growth due to sand burial was significantly greater in the disconnected treatment than

Chapter I

83

when apical ramets remained connected. This result suggests that transport of resources

from ramets growing in favorable patches can be a suitable strategy to the successful

colonization of coastal sand dunes. Sand burial is a common situation for plants inhabiting

coastal sand dunes (Maun 2009). Results showed that sand burial conducted to negative

values in plant growth. These negative values indicate that sand burial dramatically

reduced the light received by apical ramets, probably to a level below the light

compensation point, with respiration being greater than photosynthesis. In this situation,

plant carbon gain is negative, and consequently plant growth was also negative.

Physiological integration buffered the negative impact of sand burial, suggesting that

buried apical ramets were subsidized from basal ramets via photo-assimilates transport.

Many previous studies have studied the effects of physiological integration for ramets

growing under a wide variety of stressful conditions, as shade (Hartnett & Bazzaz 1983),

shade and nutrient deficiency (Slade and Hutchings 1987), water scarcity (Roiloa &

Retuerto 2005), salinity (Salzman & Parker 1985), heavy metals (Roiloa & Retuerto

2006, 2012), pathogens (D’Hertefeldt & van der Putten 1998), defoliation (Schmid et al.

1988; You 2014; Wang et al 2017) and also sand burial (Yu et al. 2002, 2004; Chen et al.

2010), generally reporting a benefit of integration for the sustained ramets, especially

when growing in unfavorable conditions. Also, previous experiments reported benefits of

physiological integration for C. edulis. Thus, Roiloa et al. (2010) found that physiological

integration favored the expansion of C. edulis clones in a coastal sand dune, especially

when confronted with native species. Similarly, Lechuga-Lago et al. (2016)

demonstrated in a greenhouse experiment that clonal integration facilitates the growth of

apical ramets of C. edulis under water stress conditions. The benefit of physiological

integration that our study reported in terms of plant growth was not observed in the total

mass at the end of the experiment. This apparent inconsistency could be explained by the

temporal lag between the negative carbon gain experienced by the plant, and the time

necessary to transfer this negative effect into a significant reduction in the final plant

biomass. In other words, it would be expected that a longer duration of the experiment

would conduct to a reduction in the final biomass of buried plants, especially in the

disconnected treatment.


84

Fig. 1 Relative growth rate of fresh mass (RGR, A), root to shoot ratio (RSR, B), root dry mass (C), shoot

dry mass (D), total dry mass (E), and number of leaves (F) (mean ± SE) of the apical ramets of Carpobrotus

edulis. Letters indicate significant differences between treatments according to Tukey test.

Chapter I

85

Fig. 2 Relative growth rate of fresh mass (RGR, A), root to shoot ratio(RSR, B), root dry mass (C), shoot

dry mass (D), total dry mass (E), and number of leaves (F) (mean ± SE) of the basal ramets of Carpobrotus

edulis. Letters indicate significant differences between treatments according to Tukey test.


86

Fig. 3 Relative growth rate of fresh mass (RGR, A), root to shoot ratio (RSR, B), root dry mass (C), shoot

dry mass (D), total dry mass (E), and number of leaves (F) (mean ± SE) of the clonal fragments (i.e. apical

plus basal ramets) of Carpobrotus edulis. Letters indicate significant differences between treatments

according to Tukey test.

Chapter I

87

Table 1. Effects of connection and sand burial on the growth and biomass allocation of apical ramets (A),

basal ramets (B), and whole clonal fragments (C) of Carpobrotus edulis. The initial fresh mass of the clonal

fragment was used as a covariate for all variables except relative growth rate of fresh mass (RGR). RSR –

root to shoot ratio. RGR – relative growth rate. F and p of two-way ANOVA/ANCOVA are given. For the

apical ramets, degree of freedom (DF) is 1, 13 for all variables except RGR for which the error DF is 12;

for the basal ramets, DF is 1, 25 for all variables except RGR for which the effort DF is 24; for the whole

clonal fragment, DF is 1, 26 for all variables except RGR for which the error DF is 25. Numbers are in bold

when p < 0.05.

Variable Covariate Connection Burial C x B

F p F p F p F p

(A) Apical ramets

RGR - - 2.63 0.127 22.97 <0.001 2.03 0.177

RSR 0.08 0.780 54.97 <0.001 0.06 0.809 0.01 0.969

Root mass 0.70 0.421 13.19 0.003 0.99 0.338 0.08 0.783

Shoot mass 9.00 0.010 1.65 0.221 6.49 0.024 0.52 0.484

Total mass 8.77 0.011 0.99 0.339 6.43 0.025 0.46 0.510

No. of leaves 4.74 0.048 0.69 0.423 0.77 0.395 0.01 0.970

(B) Basal ramets

RGR - - 0.01 0.917 1.76 0.196 0.02 0.892

RSR 0.17 0.688 3.08 0.092 6.69 0.016 9.76 0.004

Root mass 6.65 0.016 0.76 0.392 1.13 0.299 7.33 0.012

Shoot mass 6.47 0.018 1.71 0.202 2.40 0.134 0.94 0.341

Total mass 7.44 0.011 1.27 0.270 1.77 0.195 0.34 0.568

No. of leaves 2.16 0.154 4.82 0.038 1.59 0.219 0.81 0.377

(C) Clonal fragment

RGR - - 7.16 0.012 4.47 0.044 0.85 0.364

RSR 5.04 0.033 5.04 0.033 0.03 0.869 1.34 0.257

Root mass 0.01 0.908 0.01 0.908 0.63 0.435 4.55 0.043

Shoot mass 3.94 0.058 3.94 0.058 2.26 0.145 2.97 0.097

Total mass 4.14 0.052 3.39 0.077 2.16 0.154 3.21 0.085

No. of leaves 1.32 0.262 1.32 0.262 0.69 0.415 1.38 0.251


88

Interestingly, results showed that biomass partitioning was significantly affected

by physiological integration. Thus, basal ramets increased their proportional biomass

allocated to produce roots (RSR) when connected to unburied ramets. On the contrary,

basal ramets significantly increased the production of aboveground structures when their

linked apical ramets suffered sand burial. Plastic responses allow plants to adjust their

morphology and physiology to increase resources acquisition efficiency (Grime &

Mackey 2002; Valladares et al. 2007; Mommer et al. 2011). Physiological integration

allow clonal systems to act as cooperative systems by (i) showing capacity for resource

interchange between connected modules (Pitelka & Ashmun 1985; Jonsdottir & Watson

1997), and also by (ii) developing modular plasticity, this is, responding non-locally to

the conditions experienced by their linked ramets (de Kroon et al. 2005, 2009). Our results

showed that physiological integration induced a non-local compensatory plastic response

of basal ramets in response to the sand burial conditions experienced by apical ramets,

being consistent with the modular plasticity hypothesis proposed by de Kroon et al. (2005,

2009). This plastic non-local response can also be interpreted as a type of division of

labour (i.e. ramets specialization in resource acquisition within a clonal system, sensu

Alpert & Stuefer 1997; Hutchings & Wijesinghe 1997; Stuefer 1998). Thus, when apical

ramets remained unburied, physiological integration induced the production of roots in

basal ramets. In this situation, basal ramets specialized in acquisition of soil-based

resources while apical ramets significantly reduced the proportional biomass allocated to

roots and specialized in aboveground growth. This specialization could contribute to the

fast expansion of apical ramets, as the carbon saved from root production would be used

for the colonization of the sand dune surface. Previously, in a greenhouse experiment,

Roiloa et al. (2013) demonstrated the presence of developmentally-programmed division

of labour in C. edulis, with basal ramets specializing in uptake of soil-based resources

and apical ramets increasing their chlorophyll content and aboveground propagation. In

contrast, our study showed that physiological integration significantly reduced the

proportional root biomass of basal ramets when their apical ramets suffered from sand

burial. In other words, when apical ramets experienced a severe light reduction, connected

basal ramets changed the biomass-partitioning pattern, and significantly increased the

production of light-capturing structures. This was interpreted as a non-local

compensatory response, whereby basal ramets increased the light-capturing efficiency to

compensate the decline of light received by the apical buried ramets. This response

Chapter I

89

allowed in some way to buffer for the negative carbon gain experienced by apical ramets,

where the burial reduced the light received bellow the compensation point.

Summarizing, our results demonstrate that physiological integration buffers the

negative impact of sand burial suffered by apical ramets of the clonal invasive plant C.

edulis. In addition, results showed that physiological integration induced a non-local

plastic response in basal ramets that was dependent on the conditions experienced by their

apical ramets. Thus, results showed a developmentally programmed division of labor

when apical ramets remained unburied, with basal ramets specialized in acquisition of

soil-based resources while apical ramets specialized in aboveground expansion. On the

contrary, when apical ramets were sand buried, basal ramets changed their biomass

allocation pattern and increased the production of photosynthetic structures, alleviating

the light stress suffered by their connected apical ramets. Although recent experiments

have demonstrated the importance of physiological integration for the successful

expansion of the clonal invader C. edulis (Roiloa et al. 2010, 2013, 2014a,b, 2016; Portela

& Roiloa 2017), our study is the first to test the effect of physiological integration in

response to sand burial in C. edulis. Our results are pioneering in reveal that physiological

integration allows the invasive C. edulis to withstand sand burial when colonizing coastal

dunes, which can have important consequence to understand the invasive success of this

clonal species in coastal systems.

Acknowledgments

This work was supported by the Spanish Ministry of Economy and Competitiveness

(Grant CGL2013-44519-R to S. R. R.), co-financed by the European Regional

Development Fund (ERDF). This is a contribution from the Alien Species Network (Ref.

ED431D 2017/20 – Xunta de Galicia, Autonomous Government of Galicia).


90

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Slade, A. J., & Hutchings, M. J. (1987). An analysis of the costs and benefits of physiological integration

between ramets in the clonal perennial herb Glechoma hederacea. Oecologia 73, 425–31.

Stuefer, J. F. (1998). Two types of division of labour in clonal plants: benefits, costs and constraints.

Perspectives in Plant Ecology, Evolution and Systematics 1, 47-60.


92


in Potentilla anserina. Ecology Letters 2, 135–139.



Applied Vegetation Science 11, 45-54.

Valladares, F., Gianoli, E., & Gomez J. M. (2007). Ecological limits to plant phenotypic plasticity. New

Phytologist 176, 749–763

Wang, P., Li, H., Pang, X. Y., Wang, A., Dong, B. C., Lei, J. P., Yu, F. H., & Li, M. H. (2017). Clonal

integration increases tolerance of a phalanx clonal plant to defoliation. Science of the Total Environment

593–594, 236–241. doi: 10.1016/j.scitotenv.2017.03.172.


Aizoaceae). Contributions from the Bolus Herbarium 15, 76-107.

You, W., Dan, Y., Dong, X., Han, C., & Liu, C. (2014). The invasive plant Alternanthera philoxeroides

benefits from clonal integration in response to defoliation. Flora - Morphology Distribution Functional

Ecology of Plants 209(11), 666-673.

Yu, F. H., Chen, Y., & Dong, M. (2002). Clonal integration enhances survival and performance of

Potentilla anserina, suffering from partial sand burial on Ordos plateau, China. Evolutionary Ecology

15(4), 303-318.

Yu, F. H., Dong, M., & Krusi, B. (2004). Clonal integration helps Psammochloa villosa survive sand burial

in an inland dune. New Phytologist 162, 697-704.

Chapter II

Effects of clonal integration in the expansion of two alien Carpobrotus

species into a coastal dune system – a field experiment

Rubén Portela1, Sergio R. Roiloa1

1BioCost Group, Biology Department, Faculty of Science, Universidade da Coruña, A Coruña

15071, Spain.

Published as Portela, R., & Roiloa, S. R. (2017). Effects of clonal integration in the expansion

of two alien Carpobrotus species into a coastal dune system – a field experiment. Folia

Geobotanica, 52(3-4), 327-335. doi: 10.1007/s12224-016-9278-4

Chapter II

95

Abstract

What makes a plant to be a successful invader is one of the most interesting questions in

modern ecology. Comparative studies including congeners differing in invasiveness are

a straightforward approach to detect potential traits explaining invasions. In this

experiment we studied the importance of clonal integration and the capacity to buffer

fragmentation in the expansion of two stoloniferous invaders, Carpobrotus edulis,

considered more invasive, and Carpobrotus acinaciformis, considered less invasive. In

particular we aim to determine whether differences in these clonal traits may explain

differences in invasiveness between both species. We report evidence that clonal

integration favour the expansion of the two exotic clonal species into a sand dune system.

Benefit derived from clonal integration by itself does not explain differences in

invasiveness between these two exotic species. However, our results indicate that the

greater invasiveness of C. edulis could be explained by a higher capacity to buffer the

negative effect of fragmentation in comparison with C. acinaciformis. To elucidate the

real contribution of clonal traits in plant invasions, new comparative studies should be

conducted including more clonal species.

Keywords: biological invasions; biomass partitioning; Carpobrotus; clonal integration;

congeneric species; fragmentation; invasiveness; storage organs.


96

1. Introduction

Clonal growth is characterised by the production of genetically identical offspring

(named ramets) that may remain connected by stolon or rhizome internodes (Price &

Marshall 1999). Stolon or rhizome connections permit the transport of resources within

the clonal system, and ramets within a clone are thus physiologically integrated. This

capacity for clonal integration has been repeatedly documented and allows clones to

behave as cooperative systems, enabling ramets to colonize and survive in unfavourable

patches (e.g. Hartnett & Bazzaz 1983; Slade & Hutchings 1987; Alpert 1999; Saitoh et

al. 2002; Roiloa & Retuerto 2006). In addition, stolon and rhizome internodes can storage

resources. This function as reserve organs could be an important factor in the survival and

re-growth of clonal plants after an episode of fragmentation allowing the mobilization of

resources to buffer stress situations (Stuefer & Huber 1999; Suzuki & Stuefer 1999;

Goulas et al. 2001).

Clonal growth has been pointed out as a characteristic that could increase plant

invasiveness (Pyšek 1997; Liu et al. 2006; Wang et al. 2008; Song et al. 2013). This idea

is based in the rationality that many of the most successful invasive plant species show

clonal propagation. Invasive species modify the stability and functioning of local

communities, and displace native plants with the consequent loss of biodiversity

(Vitousek et al. 1996; Mack et al. 2000; Strayer 2012). Understanding the mechanisms

underlying the process of invasions is an interesting aim in ecological research (Alpert et

al. 2000; Levine et al. 2003; Blackburn et al. 2011). Recent studies have showed the

importance of clonal integration in the expansion clonal invaders (Liu et al. 2006;

Otfinowski & Kenkel 2008; Wang et al. 2008; Song et al. 2013; Roiloa et al.,2014a,b,

2016). A recent meta-analysis conducted by Song et al. (2013) showed the relationship

between clonal integration and invasiveness. Thus, the benefit of clonal integration for

recipient ramets growing in unfavourable conditions was larger in more invasive species

(Song et al. 2013). On the other hand, the role of stolons and rhizomes as storage organs

could play a crucial role in the colonization of new environments by invasive clonal

species, especially after a process of fragmentation (Dong et al., 2010, 2012; Konlechner

et al., 2016; Lin et al., 2012). As a result, clones can act as cooperative systems, buffering

the potential negative environmental conditions, and colonizing a wide variety of new

habitats that otherwise would be unexploitable by independent plants (e.g. Hartnett &

Bazzaz 1983; Salzman & Parker 1985; Slade & Hutchings 1987; Wijesinghe & Handel

Chapter II

97

1994; Jónsdóttir & Watson 1997; Yu et al. 2004; Saitoh et al. 2002; Roiloa & Retuerto

2006).

There are different comparative approaches that provide different insights into

potential determinants of invasiveness (van Kleunen et al. 2010a). Manipulative

experiments comparing congeners differing in invasiveness have been suggested as a

correct approach to detect key traits explaining the expansion of exotic species (Mack

1996; Nijs et al. 2004; van Kleunen et al. 2010a). Thus, the comparison between a less-

and more-invasive clonal species can reveal significant insights into the importance of

clonal traits in plant invasions. In this study we examined the effect of fragmentation in

the expansion of two clonal invaders, Carpobrotus edulis and Carpobrotus acinaciformis,

colonizing a coastal sand dune. C. edulis and C. acinaciformis are closely related species,

with a similar life history and growth form, and catalogued as invaders in coastal habitats

of South Europe. Importantly, C. edulis, due to its higher occurrence, is considered more

invasive than C. acinaciformis in the Mediterranean basin (Lambinon 1995; Suehs et al

2001). It seems that clonal invasive species may benefit from both clonal integration and

clonal storage organs. Thus, we aim to determine the importance of these two clonal traits

to explain the different invasiveness of the two clonal species C. edulis and C.

acinaciformis. If clonality contributes to invasiveness of Carpobrotus sp. we

hypothesized that the benefit derived from clonal integration and/or the capacity to buffer

fragmentation should be more significant in C. edulis, considered more invasive, than in

C. acinaciformis, considered less invasive.


2.1.Study species

Carpobrotus edulis (L.) N.E. Br and Carpobrotus acinaciformis (L.) L. Bolus,

commonly known as ice plants, are succulent stoloniferous plants belonging to the

Aizoaceae family and native to the Cape Region in South Africa (Wisura & Glen 1993).

Today, both species invade coastal systems of Mediterranean climate areas around the

world, with the consequent negative impact on diversity of the native flora (D’Antonio &

Mahall 1991; D’Antonio 1993; Traveset et al. 2008; Vilà et al. 2008). Both species have

an extensive plagiotropic monopodial system and show a radial growth with a structure

of nodes and internodes (Wisura & Glen 1993). New ramets can produce roots after direct


98

contact with the substrate. Clonal reproduction allows Carpobrotus sp. to form dense

mats by the production of apical ramets that remain physiologically integrated by stolon

connections (Wisura & Glen 1993). This capacity for clonal propagation allows

Carpobrotus sp. to spread horizontally and effectively colonize the surrounding area

(Roiloa et al. 2010).

According to Wisura & Glen (1993), C. edulis is the only taxa of the genus with

yellow flowers, while C. acinaciformis has magenta flowers. Leaf equilaterality has been

suggested as a useful characteristic for discriminating Carpobrotus taxa. In this sense, C.

edulis presents an equilateral leaf-cross section, while C. acinaciformis has isosceles leaf-

cross section (Gonçalves 1990; Suehs et al. 2004). In the Mediterranean basin, C. edulis,

due to its higher occurrence, is considered to be more invasive than C. acinaciformis

(Lambinon 1995; Suehs et al. 2001).


Thirty-two similar-sized unrooted ramet pairs of C. edulis and C. acinaciformis

were collected in a rocky coast area in A Coruña (NW Spain) (43°22'N, 08°24'W). Each

ramet pair was obtained by excising the third and fourth ramet from the apex of a maternal

clump. With this procedure we standardize the age, size and development stage of the

plants used in the experiment, allowing a more reliable comparison between treatments.

Both species were collected in the same area in order to avoid confounding effects derived

from different conditions at the origin area. However, with the objective of increase the

genetic diversity included in the study, each pair of ramets was obtained from a different

maternal clump. Selected clumps were at least 15 m apart from each other, and it is

assumed that each clump represents a different genotype. The collected plants were

transplanted into a coastal sand dune system in Sellese (Ares, A Coruña, NW Spain)

(43°25'N, 08°13'W) where the experimental treatments were executed. This coastal sand

dune system represents a typical habitat invaded by Carpobrotus sp. and where threatened

native species inhabit.

The experimental design comprised two crossed factors: ‘species’ (C. edulis, C.

acinaciformis) and ‘connection’ (connected, severed). In the ‘connection’ treatment,

ramet pairs were either left connected (clonal integration is allowed) or severed from each

other (clonal integration is prevented). Disconnection reflects the fact that disturbance

frequently breaks clonal fragments into smaller groups under natural conditions (Stuefer

Chapter II

99

& Huber 1999; Latzel & Klimešová 2009). No negative effects (as sudden death or

diseases) were observed due to the stolon severance. Therefore, we discard any

interference in our results derived from an initial trauma. Considering that in clonal plants

acropetal transport of soil-based resources (i.e. from the basal to the apical ramet)

generally exceeds basipetal transport (i.e. from the apical to the basal ramet) (Alpert &

Mooney 1986; Price & Hutchings 1992; Alpert 1996), each pair was subjected to a regime

of resource availability to induce clonal integration. Basal ramets were planted in 0.5l

plastic pots, which were embedded at the ground level of the dune, and filled with potting

compost (containing all main nutrients and trace elements required for optimal growth of

plants: N = 230; P2O5 = 180; K2O = 230; Mg = 150; S = 350, in mg/l) (high nutrient

conditions). Apical ramets were planted in 0.5l plastic pots, which were embedded at the

ground level of the dune, and filled with sand (low nutrient conditions). This regime of

resource availability mimics the natural conditions of Carpobrotus sp. colonizing coastal

sand dunes, where basal ramets usually create a dense layer of organic matter producing

a fertile soil, whereas developing apical ramets spread into a new area of sand with low

nutrient content (Novoa & González 2014). This variance in resource availability, with

basal ramets under more favorable conditions and developing ramets colonizing less

favorable patches, denotes the importance of clonal integration for the expansion of

Carpobrotus sp. in natural habitats.

Plant material was transplanted in an approximate area of 200 x 100 m within the

dune system. To avoid possible confounding effects of orientation, all studied apical

ramets were growing from the basal plant towards the ocean. C. edulis and C.

acinaciformis ramets under the connected and severed treatments were placed

interspersed. Each of the four experimental treatments (2 levels of connection x 2 levels

of species) was replicated 8 times. The experiment was maintained for 3 months, from 20

April until harvest on 23 July 2015, under the natural conditions prevailing in the dune

system.

2.3.Measurements

At the end of the experiment, basal and apical ramets were harvested individually,

divided into aboveground parts (leaves and stolons) and roots, dried at 80° for 72h and

weighed. Total biomass was calculated for each basal and apical ramet as the sum of

aboveground and root biomass. Root, aboveground and total biomass at whole clone level


100

(basal + apical ramets) was also calculated. Proportional biomass allocated to roots was

also determined as root biomass / total biomass (root mass ratio, RMR) for basal and

apical ramets, and at whole clone level.

2.4.Data analyses

Prior to analyses, data were checked for normality and homoscedasticity using

Kolmogórov-Smirnov and Levene test respectively. The Levene test showed that RMR

and root biomass data, and their transformations of apical ramets did not fit with the

assumptions of homoscedasticity. As a consequence non-parametric statistics were used

for these data. Differences in root, aboveground and total biomass, and RMR of basal

ramets and the whole clone were compared by two-way ANOVA, with ‘species’ and

‘connection’ as between-subject effects. Similarly, differences in aboveground and total

biomass of apical ramets were analysed by two-way ANOVA, with ‘species’ and

‘connection’ as main factors. We used Scheirer-Ray-Hare’s test, the non-parametric

equivalent of ANOVA, to examine variations between treatments in root biomass and

RMR of apical ramets. The number of replicates used in each treatment was reduced at

the end of the experiment, as indicated by the error degree of freedom of the analyses. A

total of 8 pairs of ramets (1 C. edulis and 4 C. acinacifomis in the connected treatment,

and 1 C. edulis and 2 C. acinaciformis in the severed treatment) were stolen during the

experiment. Other 3 apical ramets (2 C. edulis and 1 C. acinaciformis in the severed

treatment) did not produce any roots and died during the experiment. Significance levels

were set at P<0.05. Statistical tests were performed with SPSS 15.0 (SPSS, Chicago, IL,

USA).

3. Results

3.1.Basal ramets

Root biomass and the proportional biomass allocated to roots (RMR) of basal

ramets were significantly affected by the connection treatment (Table 1). Connection

significantly increased root biomass and RMR of basal ramets (Fig. 1). No significant

differences were detected for root biomass and RMR between species or for the

interaction between species and connection (Table 1). Aboveground and total biomass of

basal ramets were not significantly affected by the treatments (Table 1).

Chapter II

101

3.2.Apical ramets

The aboveground and total biomass of apical ramets of C. edulis were significantly

greater than the aboveground and total biomass obtained for apical ramets of C.

acinaciformis (Table 1, Fig. 1). Connection significantly increased the aboveground and

total biomass of apical ramets (Table 1, Fig. 1). In addition, aboveground and total

biomass of apical ramets were significantly affected by the interaction between species

and connection (Table 1). Connection significantly increased the aboveground and total

biomass of apical ramets, however this increase was especially important for ramets of

C. acinaciformis in comparison with ramets of C. edulis (Fig. 1). Root biomass and RMR

of apical ramets were not significantly affected by the treatments (Table 1).

3.3.Whole clone

Connection significantly increased the aboveground and total biomass at the

whole clone level (Table 1, Fig. 1). No significant effects of species or the interaction

between species and connection were detected for aboveground and total biomass of

whole clones (Table 1). Root biomass and RMR at whole clone level were not

significantly affected by the treatments (Table 1).

4. Discussion

The results of this study showed that clonal integration provides a benefit for

apical ramets colonizing the dune system, both in clones of C. edulis and in clones of C.

acinaciformis. Total biomass of apical ramets was significantly increased by connection,

reporting a benefit for the expansion of these invaders. Clonal integration is considered

one of the most striking traits associate with clonal reproduction. Transport of essential

resources from established to developing ramets, or from ramets in favorable patches to

connected ramets under more stressful conditions, has been extensively demonstrated as

a successful strategy in clonal plants (e.g. Hartnett & Bazzaz 1983; Salzman & Parker

1985; Slade & Hutchings 1987; Saitoh et al. 2002; Roiloa & Retuerto 2006; Roiloa et al.

2014c). Thus, clonal integration could provide to apical ramets the necessary resources

to outcompete neighbouring species, contributing to the success of the clonal growth habit

in terrestrial plant communities (Kliměs et al. 1997). Similarly, previous results with C.


102

edulis have showed a benefit of clonal integration in the expansion of this invader, both

in field (Roiloa et al. 2010) and in greenhouse (Roiloa et al. 2013, 2014a) experiments.

Table 1. Results of two-way analysis of variance (ANOVA) to examine the effects of connection and

species on root, aboveground and total biomass, and biomass allocated to roots (root mass ratio, RMR) of

basal ramets and whole clones. Effect of connection and species on aboveground and total biomass of apical

ramets was also tested by two-way ANOVA. Effect of connection and species on root biomass and RMR

of apical ramets was analysed by the non-parametric equivalent to two-way ANOVA Scheirer-Ray-Hare

test. Values of P < 0.05 are in bold. See Fig. 1 for data.

Interestingly, our results showed a significant increase of the root biomass by

connected basal ramets in comparison with severed basal ramets. However, aboveground

biomass of basal ramets was not significantly affected by connection. As consequence,

we obtained a significant increase of the proportional biomass allocated to roots (RMR)

in connected basal ramets, both in the clones of C. edulis and those of C. acinaciformis.

We interpreted this result as a plastic response of basal ramets in order to attend the

demand from their connected apical ramets. This result demonstrates that biomass

partitioning between above- and belowground structures is affected by clonal integration,

with a significant change in root biomass but not in aboveground biomass due to

connection. Plasticity in biomass partitioning allows plants to cope efficiently with

variation in resources availability (Grime & Mackey 2002; Valladares et al. 2007;

Mommer et al. 2011). Because abundant resources can be acquired more economically

than scarce resources, connected ramets in clonal systems usually allocate more energy

Root biomass Aboveground biomass Total biomass RMR

d.f. F P d.f. F P d.f. F P d.f. F P

Basal ramet

Species 1 0.090 0.767 1 0.409 0.530 1 0.269 0.610 1 0.249 0.624

Connection 1 10.953 0.003 1 0.408 0.530 1 1.213 0.284 1 14.634 0.001

Species x connection 1 0.272 0.608 1 1.395 0.251 1 1.282 0.271 1 0.129 0.724

Error 20 20 20 20

Apical ramet

Species 1 0.784 0.376 1 7.800 0.012 1 7.850 0.012 1 0.409 0.522

Connection 1 2.261 0.133 1 28.540 <0.001 1 18.115 0.001 1 3.065 0.080


Error 17 17 17 17

Whole clone

Species 1 0.266 0.612 1 1.332 0.264 1 1.341 0.263 1 0.188 0.670

Connection 1 0.928 0.349 1 6.821 0.018 1 6.719 0.019 1 1.276 0.274


Error 17 17 17 17

Chapter II

103

to the acquisition of locally abundant resources, and the subsequent translocation of

resources between the connected ramets brings an increase of the performance for the

whole clone (Friedman & Alpert 1991; Birch & Hutchings 1994; Stuefer et al. 1996;

Alpert & Stuefer 1997; Hutchings & Wijesinghe 1997). Similar responses to attend the

demand of developing ramets have been reported before, both at physiological (Roiloa &

Retuerto 2005, 2007) and at morphological (Roiloa & Hutchings 2012, 2013) level. Our

results suggest that clonal integration modified the plastic responses of basal ramets.

Thus, basal ramets increased their capacity to uptake soil-based resources in order to

support more efficiently their connected developing ramets. These results suggest that

clonal integration and the associate non-local changes in biomass partitioning can

increase the expansion of apical ramets in these invasive species.

Figure 1. Root, aboveground, and total biomass in g (mean + SE), and proportional biomass allocated to

roots (determined as the root mass ratio, RMR) (mean + SE) of connected (filled bars) and severed (empty

bars) of basal ramets, apical ramets and whole clones (basal + apical ramets) of C. edulis and C.

acinaciformis. See Table 1 for ANOVAs results.

Our results did not support our hypothesis that the benefits of clonal integration

are more pronounced in C. edulis than in C. acinaciformis. It seems logical to presume


104

that some plant traits could contribute more to plant invasiveness than others. In this

sense, characteristics associated to clonal propagation have been suggested as key

attributes explaining plant invasions (Pyšek 1997; Liu et al. 2006; Wang et al. 2008; Song

et al. 2013). This is idea is based on the rationality that many of the most successful plant

invaders show clonal growth. Capacity for clonal integration allows clonal plants to

compete successfully in a wide range of habitats (e.g. Hartnett & Bazzaz 1983; Alpert &

Stuefer 1997; Saitoh et al. 2002; Roiloa & Retuerto 2006; Roiloa et al. 2010), and

therefore could contribute to the expansion of clonal invaders (Song et al. 2013). As we

proposed, if clonal integration contributes to the expansion of invaders, it seems logical

to predict a higher benefit of clonal integration in the more invasive species in comparison

with the less problematic species. In this sense, comparisons between congeners differing

in the degree of invasiveness have been suggested as a suitable method to detect traits

underlying biological invasions (Mack 1996; Nijs et al. 2004). Previous studies have

found that invasive exotic species show higher values for traits related to performance

than non-invasive species, suggesting that these traits might be related with invasiveness

(van Kleunen et al. 2010b). However, contrary to our prediction, the benefit of clonal

integration in the apical ramets was significantly more accentuate in C. acinaciformis than

in C. edulis.

It is important to remark that the higher benefit of clonal integration detected in

C. acinaciformis in comparison with C. edulis, was due not to an increase of the biomass

in the connected treatment but to the reduction in the severed treatment. This is, C.

acinaciformis suffered more from disconnection than C. edulis. In other words, we can

infer that C. edulis is less affected by disconnection than C. acinaciformis. Our results

showed that disconnection significantly reduces the total biomass of apical ramets, and

this reduction was especially important in C. acinaciformis. Thus, C. edulis buffered

better the negative impact of disconnection than C. acinaciformis. We can interpret that

C. edulis is better coping with process of fragmentation, and this could favor a rapid

spread and could result in a more successful invader than C. acinaciformis. In this sense,

clonal plants are frequently affected by processes of fragmentation (Stuefer & Huber

1999; Latzel & Klimesŏvá 2009), and the capacity to survive and growth after

fragmentation has important implications for the colonization of new environments by

clonal plants, including invasive species (Dong et al. 2012; Konlechner et al. 2016).

Stolons of clonal plants act as reserve organs, and resources stored in the stolon can be

Chapter II

105

mobilized helping to buffer stress conditions, as motivated by fragmentation (Stuefer &

Huber 1999; Dong et al. 2012). In this sense, it has been showed that clonal structures

play an important role in the colonization of new environments, especially in coastal

habitats, as the occupied by the studied species (Maun, 2009). Future studies comparing

the capacity for resources storage and mobilization between C. edulis and C.

acinaciformis could contribute to explain the difference cost of disconnection reported in

our study.

In conclusion, here we report evidence that clonal integration increases the growth

of apical ramets both in C. edulis, considered more invasive, and C. acinaciformis,

considered less invasive. As a consequence, at least in this case, clonal integration by

itself does not explain differences in invasiveness between these two exotic species.

However, our results indicate that invasiveness of C. edulis could be explained by a higher

capacity to buffer the negative effect of fragmentation in comparison with C.

acinaciformis. In this study we used only one pair of species, and new common garden

experiments including more pairs of clonal species differing in invasiveness are

mandatory to allow a broader generalization of the results, and as a consequence to reveal

the real repercussion of clonal propagation in plant invasions.

Acknowledgments

Financial support for this study was provided by the Spanish Ministry of Economy and

Competitiveness and the European Regional Development’s Fund (ERDF) (grants Ref.

CGL2013-44519-R, awarded to S. R. R.). We are grateful to two anonymous referees

and to the editor Jitka Klimesova for their valuable comments on an earlier version of

this paper.


106

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Roiloa, S. R., & Hutchings, M. J. (2013). The effects of physiological integration on biomass partitioning in

plant modules: an experimental study with the stoloniferous herb Glechoma hederacea. Plant Ecology,

214(4), 521-530.

Roiloa, S. R., & Retuerto, R. (2005). Presence of developing ramets of Fragaria vesca increase

photochemical efficiency in parent ramets. International Journal of Plant Sciences, 166, 795-803.



Roiloa, S. R., & Retuerto, R. (2007). Responses of the clonal Fragaria vesca to microtopographic

heterogeneity under different water and light conditions. Environmental and Experimental Botany, 61(1),

1-9. doi: http://dx.doi.org/10.1016/j.envexpbot.2007.02.006






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418.

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Salzman, A. G., & Parker, M. A. (1985). Neighbors ameliorate local salinity stress for a rhizomatous plant in

a heterogeneous environment. Oecologia, 65(2), 273-277.


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Song, Y. B., Yu, F., Keser, L. H., Dawson, W., Fischer, M., Dong, M., & Van Kleunen, M. (2013). United we

stand, divided we fall: A metaanalysis of experiments on clonal integration and its relationship to


Strayer, D. L. (2012). Eight questions about invasions and ecosystem functioning. Ecology Letters, 15(10),

1199-1210. doi: 10.1111/j.1461-0248.2012.01817.x







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Traveset, A., Moragues, E., & Valladares, F. (2008). Spreading of the invasive Carpobrotus aff. acinaciformis

in Mediterranean ecosystems: the advantage of performing in different light environments. Applied

Vegetation Science, 11, 45-54.

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Phytologist, 176(4), 749-763.



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van Kleunen, M., Weber, E., & Fischer, M. (2010). A meta-analysis of trait differences between invasive and

non-invasive plant species. Ecology Letters, 13(2), 235-245. doi: 10.1111/j.1461-0248.2009.01418.x

Vilà, M., Siamantziouras, A. K. D., Brundu, G., Camarda, I., Lambdon, P., Médail, F., Moragues, E., Suehs,

C. M., Traveset, A., Troumbis, A. Y., Hulme, P. E. (2008). Widespread resistance of Mediterranean island

ecosystems to the establishment of three alien species. Diversity and Distributions, 14, 839-851.



Wang, N., Yu, F. H., Li, P. X., He, W. M., Liu, F. H., Liu, J. M., & Dong, M. (2008). Clonal integration affects

growth, photosynthetic efficiency and biomass allocation, but not the competitive ability, of the alien

invasive Alternanthera philoxeroides under severe stress. Annals of Botany, 101(5), 671-678.

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experiment with Potentilla simplex. Journal of Ecology, 82(3), 495-502. doi: 10.2307/2261258



Yu, F., Dong, M., & Krüsi, B. (2004). Clonal integration helps Psammochloa villosa survive sand burial in

an inland dune. New Phytologist, 162, 697-704.

Chapter III

Biomass partitioning in response to resources availability: a

comparison between native and invaded ranges in the clonal invader

Carpobrotus edulis

Rubén Portela1, Rodolfo Barreiro1, Sergio R. Roiloa1

1BioCost Group, Biology Department, Faculty of Science, Universidade da Coruña, A


Published as Portela, R., Barreiro, R., & Roiloa, S. R. (2019). Biomass partitioning in

response to resources availability: a comparison between native and invaded ranges in the

clonal invader Carpobrotus edulis. Plant Species Biology, 34(1), 11-18. doi:

10.1111/1442-1984.12228

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Abstract

Identifying the underlying mechanisms of plant invasiveness is a fast-moving research

topic in current ecology. Phenotypic plasticity has been pointed out as a trait that can

contribute to plant invasiveness. This experiment examines the presence of rapid adaptive

evolution favoring plastic biomass partitioning during the invasion process. With that

aim, we tested differences in patterns of biomass allocation between populations of

Carpobrotus edulis from South Africa (native area) and the Iberian Peninsula (invaded

area) growing under different nutrient, water and light availabilities in a common garden

experiment. Here we demonstrate that biomass partitioning in response to nutrient

availability in C. edulis differs between populations from native and invaded ranges,

indicating that this trait could be under selection during the invasion process. Thus,

nutrient shortage significantly increased the proportional production of roots in

populations from the invaded range, but not in populations from the native area. This

plastic root-foraging response may contribute to the optimization of nutrient uptake by

plants, and therefore could be considered as an adaptive strategy. Understanding the

ecological implications of rapid evolution for plastic biomass partitioning is important in

determining processes of plant adaptation to new environments, and contributes to

disentangling the mechanisms underlying plant invasiveness.

Keywords: biomass partitioning; Carpobrotus edulis; clonal growth; plant invasiveness;

rapid evolution.


114

1. Introduction

Biological invasions represent one of the most important threats for biodiversity

conservation at a global scale (Vitousek et al. 1996; Mack et al. 2000; Strayer 2012),

which is a fast moving research area in current ecology. Nowadays, a core question in the

study of biological invasions is to identify the traits favoring species invasiveness, for

which different plausible hypotheses have been developed (Richardson & Pyšek 2006;

Catford et al. 2009). Phenotypic plasticity has been pointed out as a trait that can

contribute to plant invasiveness (Richards et al. 2006; Davidson et al. 2011; Pichancourt

& Van Klinken 2012; Keser et al. 2014). Phenotypic plasticity has been defined as the

changes in the phenotypic expression of a genotype under different environmental

conditions. Plasticity, at both the morphological and physiological level, can be

considered as an important mechanism that allows plants to cope with new or changing

environments (Grime & Mackey 2002; Valladares et al. 2007; Mommer et al. 2011). A

widely described case of phenotypic plasticity is plant reaction to limiting resources,

mainly water, light or nutrients. In such situation, plant responses involve changes in

biomass partitioning and adjusting energy allocation to develop the structures responsible

for obtaining the most limiting resource, and therefore optimize plant growth, as stated

by the optimal partitioning theory (Thornley 1972; Bloom et al. 1985; Hilbert 1990;

Gleeson & Tilman 1992). Thus, plasticity in biomass partitioning allows plants to cope

successfully with heterogeneity in resource availability (Grime & Mackey 2002;

Valladares et al. 2007; Mommer et al. 2011). Similarly, rapid adaptive evolution of

introduced populations has also been pointed out as a mechanism that explains plant

invasiveness (Lee 2002; Maron et al. 2004; Sax et al. 2007; Colautti & Barrett 2013;

Colautti & Lau 2015). Thus, populations at the new range experience rapid evolution to

produce or intensify those traits reported to be an advantage in their new local

environment. In this scenario, it is realistic to predict rapid selection favoring genotypes

with high phenotypic plasticity during the invasion process (Lande 2015).

In this study we examine the patterns of biomass allocation in the clonal invader

Carpobrotus edulis in response to the availability of three essential resources: water, light

and nutrients. We analyzed the allocation patterns through relative biomass allocation,

which is the proportion of biomass devoted to producing above- and belowground

structures, estimated as the root mass/total mass ratio. In addition, in order to detect the

presence of adaptive selection of capacity for plastic biomass partitioning during a

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115

process of invasion, we tested differences in patterns of biomass allocation between

populations of C. edulis from the native and the invaded ranges. For this, individuals from

South Africa (native area) and the Iberian Peninsula (invaded area) grew in a common

garden experiment under different nutrient, water and light levels, and changes in biomass

partitioning were measured. Intraspecific comparisons between individuals from native

and invaded ranges, especially those conducted under common environmental conditions,

are considered a suitable approach in order to detect causes of plant invasiveness (van

Kleunen et al. 2010). Specifically, our hypotheses were as follows. (a) Proportional

allocation of biomass to roots will be affected by resource availability. Thus, low levels

of belowground resources (nutrients and water) will result in an increase of root biomass,

whereas low levels of aboveground resources (light) will produce a decrease in the

proportional root growth. In other words, plants will respond to low light conditions by

increasing the proportional production of leaves. Plasticity allows plants to adjust biomass

distribution between below- and aboveground structures to enhance resource uptake

(Hilbert 1990; Gleeson & Tilman 1992). Consequently, an increase in the proportional

production of roots is expected under low availability of soil-based resources, and an

increase of the proportional biomass allocated to leaves is expected in shade conditions

(Bloom et al. 1985). (b) We also predict that changes in biomass partitioning patterns in

response to resource availability will be more accentuated in populations from the

invaded range than in populations from the native range. This hypothesis is based on the

understanding that positive selection of favorable traits, as a capacity for biomass

partitioning, could lead to rapid adaptive evolution of genotypes during the process of

plant invasion (Lee 2002; Lande 2015), and therefore phenotypes with greater plastic

foraging abilities could be better represented at the introduced range.

2. Materials and methods

2.1.Study species

Carpobrotus edulis (L.) N.E. Br is a succulent clonal plant native to the Cape

Region (South Africa) and considered invasive in coastal systems of Australia, New

Zealand, southern Europe and the USA (D’Antonio & Mahall 1991; Traveset et al. 2008).

C. edulis has an extensive plagiotropic monopodial system and show a radial growth with

a structure of nodes (called ramets) that can remain physiologically integrated by


116

internode connections (Wisura & Glen 1993; Roiloa et al. 2010, 2014; Portela & Roiloa

2017). This type of vegetative growth allows C. edulis a very effective colonization of

the surrounding area (Roiloa et al. 2010). Frequently, C. edulis plants occupying coastal

ecosystems are exposed to disturbances that break the clonal system into fragments of

different sizes (Roiloa & Retuerto 2016, Roiloa et al. 2017). New ramets of C. edulis can

produce roots after direct contact with the substrate, and can survive even if disconnected

from the basal part of the clonal fragment. Previous studies have been conducted to

determine several aspects of C. edulis ecology, including plant–pollinator networks

(Bartomeus et al. 2008), plant–soil feedbacks (de la Peña et al. 2010) or hybridization

studies (Vilà & D’Antonio 1998; Suehs et al. 2004). Also, the effect of physiological

integration (Roiloa et al. 2010, 2013, 2014, 2016; Lechuga-Lago et al. 2016; Portela &

Roiloa 2017) and the role of storage organs (Roiloa and Retuerto 2016; Roiloa et al. 2017)

in the performance of C. edulis have been recently studied. However, plasticity in biomass

partitioning in response to resource availability has not been previously studied in C.

edulis.

2.2.Plant material

Plant material of C. edulis was collected in spatially separated populations: four

in the native range (Cape Region, South Africa), and four in the invaded range (Iberian

Peninsula) (Fig. 1). In order to obtain a wider genetic representation, 36 clumps separated

by at least 25m from each other were selected in each population. Four-member unrooted

clonal fragments were excised at the edge of each clump. Clonal fragments contained the

first four ramets from the apices. Plant material was collected in winter 2015 and

maintained in common garden conditions for 10 months before the experiment began. A

random bulk sample of these plants was used for this experiment.


In January 2016, 80 unrooted ramets with a similar size were selected from the

plant stock, and placed individually in 0.4L plastic pots. Selected plants comprised the

third apical ramet from each clonal fragment, thus ensuring that all the plant material used

in the experiment had the same developmental stage. The experimental design consisted

of two crossed factors, with region (native and invaded) and resources (control, water,

light and nutrients) as main factors. For the region factor, plants from native (South

Africa) and invaded (Iberian Peninsula) ranges were included in the experiment. For the

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117

resources factor, plants were subjected to control, water, light and nutrient treatments. In

the control treatment, plants were maintained at field capacity by watering when needed

(well-watered conditions), grew in a substrate consisting of a 1:1 mixture of potting

compost (containing all main nutrients and trace elements required for optimal growth of

plants: N = 230, P2O5 = 180, K2O = 230, Mg = 150, S = 350, in mg L−1) and sand (high

nutrient conditions), and received 100% of ambient light (un-shaded conditions). In the

water treatment, plants were subjected to the same light and nutrient conditions as the

control, but did not receive water at all during the experiment (low water conditions).

Plants in the light treatment were under the same water and nutrient conditions as those

experienced by plants in the control treatment, but ambient light was reduced to 10% with

a polypropylene shading screen (measured with a Light Meter LX-107, Lutron Electronic

Enterprise, Taipei, Taiwan) (low light conditions). The nutrient treatment consisted of the

same water and light conditions as used for the control, but with plants growing in sand

without an additional supplement of nutrients (low nutrients conditions).

Experimental treatments were replicated 10 times (n = 10). Plants from each

population sampled at the native and invaded areas were randomly assigned to each

resource treatment. Initial plant size was estimated by fresh mass. Preliminary analysis

showed that the initial plant sizes did not differ significantly between the treatments

(ANOVAs: F1,64 = 0.303, p = 0.584 for region; F3,64 = 0.189, p = 0.904 for resources; F3,64

= 0.722, p = 0.543 for region × resources). The experiment was carried out in a light- and

temperature controlled growth chamber at the Unit of Ecology of the University of A

Coruña, with a 12/12-h photoperiod and at 21ºC. This photoperiod is typically registered

during early springtime on the NW Iberian Peninsula coast. Treatments began on January

11, 2016 and all plant material was harvested after 35 days to avoid the onset of resource

limitation caused by the confinement of roots within the pots.

2.4.Measurements

At the end of the experiment, ramets were harvested individually, divided into

shoot parts (leaves and stems) and roots, dried at 70ºC for 72h and weighed. Total mass

was calculated for each ramet as the sum of the shoot and root mass. Proportional biomass

allocated to roots was determined as root mass/total mass (root mass ratio, RMR).


118

Figure 1. Site location (latitude and longitude) of Carpobrotus edulis populations sampled from the native

(Cape Region in South Africa) and the invaded (Portugal and Spain in Europe) regions. (SA, South Africa;

PT, Portugal; SP, Spain).

2.5.Statistical analyses

Prior to analyses, data were checked for normality and homoscedasticity using

Kolmogorov–Smirnov and Levene tests. All data met the requirements for parametric

analysis of variance (ANOVA) and no transformations were required. Differences in root

mass, shoot mass, total mass and proportional mass allocated to roots (RMR) of plants

after the experiment were compared by two-way ANOVA with “region” and “resources”

as between-subject effects. When results were significant, a posterior Tukey tests were

applied to detect differences between the four resource treatments, and a t-test used for

differences between the native and invaded range within each resource treatment. A total

of eight plants died during the experiment and were not included in the data analyses, as

indicated by the error degree of freedom of the ANOVA. Significance levels were set at

p < 0.05. Statistical tests were performed with IBM SPSS Statistics 23.0 (IBM Corp.,

Armonk, NY, USA).

Chapter III

119

3. Results

Root mass and the proportional biomass allocated to roots (RMR) were

significantly affected by resource availability (Table 1). Thus, both for native and the

invaded range populations, root mass and RMR were significantly reduced in the low

light treatment (Figure 2). Interestingly, the analyses showed a significant effect of the

interaction between the factors region and resource treatments on root mass and RMR

(Table 1). There was a significant increase in the root mass and RMR in response to the

reduction in nutrient availability (low nutrients conditions); however, this response was

only detected in the plants from the invaded range and not in plants from the native range

(Figure 2). The effect of treatments on total and shoot mass was not statistically

significant (Table 1).

4. Discussion

The results obtained support our first hypothesis that the partitioning of biomass

between above- and belowground parts would be affected by resource availability. Thus,

as we predicted, the proportional allocation of biomass to roots by plants, both from the

native and the invaded range, growing under low light conditions, was significantly lower

than in plants growing in the high light environment. In other words, as described by the

optimal partitioning theory, plants responded to light limitation by allocating more

biomass to produce leaves, the structures to acquire the most limiting resource (Thornley

1972; Bloom et al. 1985; Hilbert 1990; Gleeson & Tilman 1992). This plasticity in

biomass partitioning would enhance the plant’s capacity to buffer variations in resource

availability, and thus could facilitate fast adaptation to changing or new environments.

Changes in phenotypic expression in response to environmental conditions (i.e.

phenotypic plasticity) have been extensively studied before (Silvertown 1998; Sultan

2001; van Kleunen & Fischer 2005; Valladares et al. 2006), as this is an important trait

that explains the ability of plants to cope with variations in resource availability (Grime

& Mackey 2002; Valladares et al. 2007). However, contrary to our prediction, we did not

0observe a plastic response of the plants to water stress. Thus, for the populations from

both the native and invaded ranges, reduction of water availability did not lead to an

increase of the proportional biomass allocated to roots. The most plausible explanation

for this unexpected result is that water stress was not enough to trigger a plant response,


120

even when plants were subjected to water deprivation throughout the entire experiment.

C. edulis is a succulent plant well adapted to colonize harsh habitats such as coastal sand

dunes (Wisura & Glen 1993), where water stress is severe and represent one of the most

important limiting factors (Maun 2009). Future studies testing the effect of water

deprivation during a longer period than the tested in our experiment would determine

more precisely the plastic response of C. edulis to water scarcity.

Figure 2. Root (A), shoot (B) and total (C) biomass in g (mean ± SE), and proportional biomass allocated

to roots (D) (determined as the root mass ratio, RMR; mean ± SE) of plants growing in control, low water,

low light, and low nutrients treatments. Plants from the invaded range (Iberian Peninsula) (filled bars) and

from the native range (South Africa) (empty bars) are represented separately. Letters on the bars indicate

differences between treatments found by Tukey test. Stars indicate differences between invaded and native

range for each resources availability treatment. See Table 1 for ANOVA results.

Interestingly, as we predicted in our second hypothesis, the responses to resource

level were more accentuated in populations from the invaded range than in those from the

native range. This response was especially evident in the nutrient treatment. Thus, low

nutrient availability increased significantly the proportional production of roots in

comparison with the high nutrient conditions (control treatment). However, this response

to nutrient availability was only detected in the populations from the invaded range, and

not in the populations from the native area. This root-foraging response may contribute

to optimization of nutrient uptake by plants, and therefore could be considered as an

adaptive strategy. Comparisons between populations from native and invaded ranges in

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121

common garden experiments are considered as an accurate approach to detect the

presence of selection pressures during the invasion process, and provide significant

insights into the causes of plant invasiveness (Lee 2002; Hierro et al. 2005; van Kleunen

et al. 2010). Phenotypic plasticity allows plants to acquire resources efficiently, in

particular when growing under environmental heterogeneity (Grime & Mackey 2002;

Valladares et al. 2007; Mommer et al. 2011), and consequently could be considered as an

important trait allowing plants to cope successfully with new environments, especially if

these differ significantly from the original environments. It seems logical that phenotypic

plasticity can be positively selected during the process of invasion, allowing exotic

species to successfully colonize new environments (Lande 2015). Rapid evolutionary

adaptation has been described during invasion processes (Prentis et al. 2008; Whitney &

Gabler 2008; Colautti & Barrett 2013; Vandepitte et al. 2014; Colautti & Lau 2015; Roiloa

et al. 2016), and phenotypic plasticity has been suggested as a trait contributing to species

invasiveness (Parker et al. 2003; Richards et al. 2006)

In addition, plastic responses to the environment could be considered a key trait

for clonal invaders, as the low genetic diversity usually present in clonal plants could be

compensated for by plasticity, buffering the lack of adaptation based on sexual

reproduction. Many of the most problematic invasive plant species show clonal

propagation, and clonality has been suggested as an important attribute explaining plant

invasiveness (Pyšek 1997). In particular, traits associated to clonal growth, such as the

capacity for physiological integration (Liu et al. 2006; Wang et al. 2008; Song et al. 2013)

or the presence of storage structures (as stolons and rhizomes) (Dong et al. 2010, 2012;

Lin et al. 2012) have been studied as mechanisms favouring the expansion of clonal

invaders. Also, recent studies have demonstrated the benefits of these clonal traits for the

propagation of C. edulis (Roiloa et al. 2010, 2013, 2014; Lechuga-Lago et al. 2016;

Roiloa & Retuerto 2016; Portela & Roiloa 2017; Roiloa et al. 2017). However, our

experiment is the first testing if biomass partitioning in response to levels of resources

differs between populations from native and invaded ranges in clonal C. edulis.

Previously, Roiloa et al. (2016) reported a greater benefit of division of labor for C. edulis

at the nonnative range, indicating the presence of rapid adaptive evolution. Division of

labor is defined as a specialization to acquire locally abundant resources developed by

connected modules of clonal systems, which generally demonstrates an overall benefit at

Table 1. Results of two-way analysis of variance (ANOVA) to examine the effects of resources availability (control, water, light and nutrients) and region (native: South Africa,

and invaded: Iberian Peninsula) on root, shoot, total mass, and root mass ratio (RMR). See Fig. 2 for data.

Root mass Shoot mass Total mass RMR

Effect d.f. F P d.f. F P d.f. F P d.f. F P

Resources 3 7.434 <0.001 3 0.399 0.754 3 0.540 0.657 3 8.468 <0.001 Region 1 0.008 0.931 1 2.197 0.143 1 2.138 0.149 1 0.050 0.824

Resources x region 3 3.054 0.035 3 0.277 0.842 3 0.384 0.765 3 2.660 0.056

Error 64 64 64 64

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123

the whole clone level (Alpert & Stuefer 1997; Hutchings & Wijesinghe 1997). This

division of labour could be considered as a modular plasticity (sensu de Kroon et al. 2005,

2009) where, in opposition to the optimal partitioning theory (Bloom et al. 1985; Gleeson

& Tilman 1992), there is an increase of the biomass allocation to structures that acquire

the proportionally most abundant resource.

Here, we demonstrate that individual ramets of C. edulis (i.e. not physiologically

integrated) increased their capacity for plasticity in biomass partitioning at the introduced

range, suggesting that it is not only clonal traits could explain the successful expansion

of our target species. This result is especially interesting as C. edulis inhabits coastal

habitats where it is frequently subjected to natural disturbances that fragment the clonal

system into portions of different sizes (Roiloa & Retuerto 2016; Roiloa et al. 2017).

Fragmented clonal plants can be transported long distances, which is an important

mechanism for colonization of new environments in coastal areas (Harris and Davy

1986), and could have significant effects for plant invasions (Trakhtenbrot et al. 2005).

In this situation, individual ramets of C. edulis cannot obtain the widely described benefit

of physiological integration (e.g. Hartnett & Bazzaz 1983; Salzman & Parker 1985; Slade

& Hutchings 1987; Alpert 1999; Saitoh et al. 2002; Roiloa & Retuerto 2006) and therefore

would rely greatly on individual plasticity to cope with the new habitat. In addition,

coastal dune ecosystems occupied by C. edulis frequently present a gradient of nutrients,

increasing their availability from the shoreline to inland (Rajaniemi and Allison, 2009).

Thus, changes in patterns of biomass allocation can also contribute to rapid adaptation of

C. edulis to local conditions, with a heterogeneous distribution of nutrients along a

gradient. In summary, plasticity in biomass allocation can represent a benefit, in addition

to clonal attributes, favoring invasiveness of C. edulis. A recent study conducted with the

invasive clonal herb Alternanthera philoxeroides reported high levels of phenotypic

plasticity, in populations from both the native range (Argentina) and introduced range

(USA and China) , pointing out that phenotypic plasticity is a common trait for the success

of this clonal invader (Geng et al. 2016).

In spite of the greater plasticity in biomass partitioning in response to nutrient

availability detected in populations from the introduced range, the results did not reveal

differences in total biomass between C. edulis populations; that is, the predictable benefit

derived from phenotypic plasticity was not transferred into plant growth. Although we

did not detect differences in total biomass in the short term, the greater plasticity exhibited


124

by populations from the invaded range could indicate important benefits in the long term

for the expansion of C. edulis. The net carbon gain of the plants is affected by patterns of

carbon allocation, and the increase in biomass allocated to photosynthetic structures

(shoot biomass) will increase the rate of carbon uptake, and as consequence plant growth.

The increase in carbon gain accumulated during a long period could bring a significant

increase in total biomass for C. edulis in the invaded region. This reveals the importance

of timescales for experimental design and interpretation of results when studying the

ecological effects of biomass partitioning. Similarly, the significant increase in biomass

allocated for root production under low nutrient conditions would allow C. edulis to

efficiently buffer soil-based resources against scarcity at the introduced range, favoring

invasiveness of this clonal species.

In conclusion, in our study we demonstrate that biomass partitioning in response

to nutrient availability in C. edulis differs between populations from native and invaded

ranges, indicating that this trait could be under selection during the invasion process.

Understanding the ecological implications of rapid evolution for the capacity for biomass

partitioning is important for determining processes of plant adaptation to new

environments, and contributes to the disentangling of the mechanisms underlying plant

invasiveness. However, for a more accurate generalization about the importance of plastic

biomass partitioning in explaining plant invasiveness, future research including multiple

species pairs should be conducted. Also, long-term experiments should be conducted in

order to accurately determine how the levels of resources could affect biomass allocation

patterns. Because changes in root to shoot ratios can affect plant carbon gain in the long

term, short-term responses cannot be directly extrapolated and time should be considered

as an important factor.

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Acknowledgments

We thank L. Álvarez from the Spanish Ministry of Agriculture, Food, and the

Environment for assistance with the authorization procedure for the plant material

importation from South Africa. Financial support for this study was provided by the

Spanish Ministry of Economy and Competitiveness (project Ref. CGL2013-44519-R, co-

financed by the European Regional Development Fund (ERDF), awarded to S. R. R.).

This is a contribution from the Alien Species Network (Ref. ED431D 2017/20 – Xunta

de Galicia, Autonomous Government of Galicia).


126

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Chapter IV

Importance of plasticity in response to soil nutrient content and

competitive ability in explaining invasiveness of the clonal

Carpobrotus edulis: a trans-continental study

Rubén Portela1, Rodolfo Barreiro1, Peter Alpert2, Cheng-Yuan Xu3,

Bruce L. Webber4,5,6, Sergio R. Roiloa1

1BioCost Group, Department of Biology, Faculty of Science, Universidade da Coruña, A


2Biology Department, University of Massachusetts, Amherst, MA 01003, USA.

3School of Health, Medical and Applied Sciences, Central Queensland University,

Bundaberg, QLD 4670, Australia.

4CSIRO Land and Water, 147 Underwood Avenue, Floreat, Western Australia 6016,

Australia.

5School of Biological Sciences, The University of Western Australia, 35 Stirling

Highway, Crawley, Western Australia 6009, Australia.

6Western Australian Biodiversity Science Institute, 133 St Georges Terrace, Perth,

Western Australia 6000, Australia.

Chapter IV

133

Abstract

A pivotal objective in the study of biological invasions is to determine which traits favor

the spread and dominance of invasive species. The main aim of this study is to determine

the importance of morphological plasticity in response to soil nutrient availability and

competitive ability to explain the invasive success of the clonal Carpobrotus edulis. To

reach this objective two different comparative approaches were used: (i) comparison

between invasive and non-invasive exotic Carpobrotus congeners, and (ii) comparisons

between C. edulis populations from native (South Africa) and introduced ranges (Iberia,

California and Australia). Results suggest the presence of rapid adaptive evolution during

the process of invasion of C. edulis. Thus, populations from the introduced range showed

significantly greater plant growth in response to soil nutrient addition than populations

from the native range. However, detected differences in plant growth were not transferred

into a higher capacity for competition in non-native range populations. Results did also

found differences between the invasive C. edulis and their less invasive congeners (C.

chilensis and C. acinaciformis), with mixed results in relation to our hypotheses. Studying

phenotypic plasticity and competitive ability seems key for disentangle the underlying

mechanisms of C. edulis invasiveness, and should be taken into account for devising

future management policies for this invasive species.

Keywords: Adaptive selection; biomass partitioning; Carpobrotus; plant invasions;

phenotypic plasticity; competitive ability.


134

1. Introduction

Biological invasions are considered to be one of the main threats to biodiversity

at a global scale (Vitousek et al. 1996; Mack et al. 2000; Strayer 2012), being a very

active research topic in modern ecology (Richardson & Pyšek 2008). A pivotal objective

in the study of biological invasions is to determine which traits favor the spread and

dominance of invasive species (Richardson & Pyšek, 2006; Thuiller, et al. 2006; van

Kleunen, et al. 2011; Ordonez 2014). Determinants of plant invasiveness are expected to

be complex, as different factors can interact facilitating or hindering the plant success

through an invasion process (Levine et al. 2003; Pyšek & Richardson 2007). After

introduction at the non-native range, exotic plants have to cope with new environmental

conditions and compete with native species. In this sense, both phenotypic plasticity and

competitive ability have been pointed as traits that could play an important role during

plant invasions (Blossey & Nötzold 1995; Davidson, et al. 2011; Pichancourt & van

Klinken 2012; Keser et al. 2014; Schultheis & MacGuigan 2018).

Phenotypic plasticity is the capacity of a single genotype to produce different

phenotypes according to the characteristics of the environment in which it develops

(Bradshaw 1965), allowing plants to successfully conform the conditions encountered in

a new or changing habitat (Sultan 2000; Grime & Mackey, 2002). For example, plastic

changes in biomass partitioning allow plants to cope with spatial and temporal variation

in essential compounds availability, as water, nutrients and light, increasing the resources

acquisition effectiveness (Gleeson & Tilman 1992; Mommer et al. 2011; Valladares et al.

2007). In this sense, phenotypic plasticity is expected to contribute to the successful

colonization of new environments by plants, and previous studies have pointed it as a key

trait to explain plant invasiveness (Parker et al 2003; Richards et al. 2006; Lande 2015).

On the other hand, it is logical to presume that plants with high capacity for plastic

adaptation to new or changing conditions would gain a competitive advantage for the

acquirement of limiting resources over those plants less plastic. Interspecific competition

can be defined as the ability of a species to acquire limiting resources, reducing its

availability to other species. This capacity to efficiently acquire resources has been

pointed out as an important determinant to explain the establishment and expansion of

invasive species (Burke & Grime 1996; Gioria & Osborne 2014; Levine et al. 2003;

Matzek 2012).

Chapter IV

135

There are different comparative approaches for assessing determinants of

invasiveness, including comparison between invasive and non-invasive exotic species

and comparisons between populations from native and introduced ranges (van Kleunen

et al. 2010). Arguably, the most straightforward approach to detect traits explaining plant

invasiveness is the comparison between invasive and non-invasive exotic species (Mack

1996; Nijs et al. 2004; van Kleunen et al. 2010), the "target-area approach" (Pyšek et al.

2004), by establishing a relationship between the presence (or intensity) of a specific trait

and the invasive success. This approach would explain why some species become

invasive while others remain as exotic non-invasive. On the other hand, comparisons

between populations from the native and non-native range of an invasive species are

considered a correct approach in order to detect positive selection of key traits during the

invasion process, reporting important information about the drivers of plant invasiveness

(Hierro et al. 2005; Lee 2002; van Kleunen et al. 2010). Thus, rapid adaptive evolution

of exotic naturalized species, by which favorable traits are selected, has been pointed out

a plausible explanation of plant invasiveness (Bossdorf et al. 2005; Maron et al. 2004;

Lavergne & Molofsky 2007; Colautti & Lau 2015).

The main aim of our study is to determine the importance of morphological

plasticity in response to soil nutrient availability and competitive ability to explain the

invasive success of the clonal Carpobrotus edulis. To reach this objective, we have

designed an experiment with two different approaches: (A) comparison between four

congeners of Carpobrotus differing in invasiveness (C. edulis, C. acinaciformis, C.

chilensis and C. virescens) (target-area approach), and (B) comparison between

populations of C. edulis from native and non-native range around the world (South Africa,

Iberia, California, and South Australia) (inter-range approach). Common garden

experiments were conducted with plants exposed to different soil conditions (high and

low nutrients availability) and interspecific competition levels (no competition and

completion). Both plasticity and competitive ability are expected to be greater in the

invasive C. edulis than in their less invasive congeners. Also, both plasticity and

competitive ability are expected to be greater in populations of C. edulis from the invaded

range than in the population from the native range. Specifically, we hypothesized that (1)

plant growth and plasticity in response to nutrients availability would be greater in the

invasive C. edulis from California than in the exotic non-invasive C. chilensis from

California. (2) Plant growth and plasticity in response to nutrients availability would be


136

greater in the invasive C. edulis from NW Iberia than in the less extended C. acinaciformis

from NW Iberia. (3) When growing together, competitive ability would be greater in the

invasive C. edulis would than in the co-occurring exotic non-invasive congener C.

chilenesis. (4) When growing together, competitive ability would be greater in the

invasive C. edulis would than in the co-occurring less extended invasive congener C.

acinaciformis. (5) Competitive ability of the invasive C. edulis over the native Ammophila

arenaria would be greater than the competitive ability of the invasive less extended C.

acinaciformis over the native A. arenaria. (6) Plant growth and plasticity in response to

nutrients availability of C. edulis would be greater at the invaded range (NW Iberia, SW

Australia, and Northern California) than at the native range (South Africa). (7)

Competitive ability of the C. edulis from the invaded range (SW Australia) over the native

congener C. virescens would be greater than the competitive ability of the C. edulis from

the native range (South Africa) over the native congener C. virescens. (8) Competitive

ability of C. edulis from the invaded range (NW Iberia) over the native A. arenaria would

be greater than the competitive ability of C. edulis from the native range (South Africa)

over the native A. arenaria (schematic representation for the different approaches, A-B,

and specific hypotheses, 1-8, is shown in Fig. 1).


2.1.Studied species

Carpobrotus edulis (L.) N. E. Br. and Carpobrotus acinaciformis (L.) L. Bolus,

commonly named ice plants, are succulent clonal plants belonging to the Aizoaceae

family, natives to the Cape Region in South Africa (Wisura & Glen 1993). According to

Wisura and Glen (1993), C. edulis is the only species of the genus with yellow flowers,

while C. acinaciformis has magenta flowers. However, since C. edulis flowers may vary

on its color with aging (from yellow to pink), leaf equilaterality has been suggested as a

useful trait for discriminating these Carpobrotus taxa. Thus, C. edulis presents an

equilateral leaf-cross section, while C. acinaciformis has an isosceles leaf cross-section

(Gonçalves 1990; Suehs et al. 2004; Campoy et al. 2018). C. edulis has been catalogued

as invasive species of coastal systems of Mediterranean climate regions around the world,

including California, South Europe, South Australia, and Chile, causing a negative impact

on native flora diversity (D'Antonio & Mahall 1991; Traveset et al. 2008; Campoy et

Fig. 1. Schematic representation of the experimental design: target-area (A) and inter-range (B) approaches are showed. Representation of the specific hypotheses tested (1-8)

is included within each approach and tested trait (phenotypic plasticity and competitive ability). See text for details.


138

al. 2018). Similarly, C. acinaciformis invade coastal habitats in south Europe, but C.

edulis has been suggested to be more invasive, due to its higher occurrence (Lambinon

1995; Suehs et al. 2001). Carpobrotus virescens (Haw.) Schwantes, commonly known as

coastal pig-face, is a species of the genus Carpobrotus native from Western Australia and

no invasive elsewhere. It grows on limestone cliffs or sand dunes at coastal ecosystems.

Flowers have pale pink petals ranging between 20-36mm long and a yellow center.

Carpobrotus chilensis (Molina) N. E. Br., commonly named sea fig, is present on the

California coast co-occurring with populations of C. edulis. The origin of C. chilensis is

unknown, but probably it is also native from Southern Africa (Vivrette 2012). It has been

present in California since at least the 1600s, but showing no negative effects on native

species and is not considered as invasive (Bicknell & Mackey 1998; Vivrette 2012). C.

chilensis is smaller than C. edulis, and show rose-magenta flowers. Carpobrotus spp.

develop an extensive monopodial system and present radial growth with a structure of

nodes and internodes. Clonal reproduction allows Carpobrotus spp. to form dense mats

by the production of apical ramets that remain physiologically integrated by stem

connections, allowing the plants to spread horizontally and colonize the surrounding area.

Ammophila arenaria (L.) Link, commonly named marram grass, is a perennial

grass of the Poaceae family. Its native range is the coastline of Europe and North Africa,

growing on mobile or semi-stable sand dunes (Purer 1942). A. arenaria is very effective

at stabilizing dunes, being extensively used for that purpose worldwide. It was introduced

on the 1800s on the Pacific coast of North America, where it displaces native species

(Slobodchikoff & Doyen 1977). The reproduction of this species is mainly clonal, through

rhizomes, since the seedling mortality is very high due to desiccation, burial or sand

erosion (Huiskes 1977). From the rhizome emerges a vigorous and extensive root system

that fixes both plant and sand, and acts as water storage (Chergui et al. 2017).

2.2.Collection and propagation

Plant material for the experiment included five species: four congeners of the genus

Carpobrotus (C. edulis, C. acinaciformis, C. chilensis and C. virescens) and the clonal A.

arenaria. These species were collected in four regions around the world: NW Iberia

(Europe), northern California (USA), SW Australia, and the Cape Region (South Africa).

Thus, C. edulis was collected in 4 populations at the native range (South Africa), and 11

populations at the invaded range (4 in NW Iberia, 4 in California, and 3 in SW Australia).

C. acinaciformis was collected at 4 populations in NW Iberia (non-native range), the

Fig. 2. Map showing the collection sites around the world for the plant material used in the experiment. See Table S1 for latitude and longitude information and species sampled

at each site location.


140

exotic non-invasive C. chilensis was collected in 4 populations at California (non-native

range), C. virescens material was collected in 3 population in their native range (SW

Austrlia), and A. arenaria was collected in 1 population from NW Iberia (native range)

(see Fig. 2 and Table S1 for detailed information of the collecting sites). Sampling

protocol was similar for all the species and sites. At each site plant individuals were

sampled at least 2m apart from each other to increase genetic representation. In the case

of Carpobrotus spp., clonal fragments contained at least 2 unrooted apical ramets were

collected. For A. arenaria individual plants containing at least 3cm of rhizome were

collected. Plant material was collected in winter 2015 (South Africa and Iberia

samplings), January 2016 (California sampling) and November 2016 (Australia

sampling). All the material collected at the field was transported to an experimental

garden at the University of A Coruña (Spain) and grown in common conditions (trays

filled with sand from coastal sand dunes and watered regularly) until the experiment

began, in order to reduce maternal environment effects.


In March 2017 a total of 280 healthy plants from the different collection site were

selected from the plant stock for use in the experiments, and placed individually in 2 L

plastic pots. For Carpobrotus spp. the plants used in the experiment corresponded with

the first ramet from the apices, and were selected for size uniformity. None of the selected

plants had roots at the start of the experiment. For testing differences between invasive

and exotic less-invasive Carpobrotus in phenotypic plasticity (target-area approach in

plasticity) the experimental designs consisted of two crossed factors with ‘species’

(invasive, exotic less-invasive) and ‘nutrients’ (high, low) as main factors. The ‘species’

factor included the invasive C. edulis from California (non-native range) and the exotic

non-invasive C. chilensis from California (non-native range), and the invasive C. edulis

from NW Iberia (non-native range) and the less invasive C. acinaciformis from NW Iberia

(non-native range). In the ‘nutrients’ factor plants were either subjected to a regime of

low nutrients (pots filled with sand dune) or grew under high nutrients conditions (pots

filled with sand dune plus 8g of slow release granular fertilizer, Osmocote Bloom, ICL

Specialty Fertilizers Iberia; NO3 = 212; NH4 = 268; P2O3 = 280; K2O = 720; Fe = 14; Mn

= 2; Cu = 1.8; Mo = 0.8; Zn = 0.4 in mg·L−1). Plants from the each of the populations

sampled in the field were represented and randomly assigned to the experimental

treatments. Each treatment was replicated 10 times (n = 10) (see Fig. 1 for schematic

Chapter IV

141

representation of the experimental design). The experiment was carried out at the same

experimental garden of the University of A Coruña (Spain) where plant material was

propagated. Plants were watered regularly during the experiment to avoid water stress.

Treatments began on 20 March and continued for 110 days, until July 2017.

2.4.Measurements

Treatments were ended after 110 days, on July 2017, when plants started to show

crowding of roots at the bottom of pots. Each ramet of Carpobrotus plus its new roots,

stems, and offspring ramets was weighed so that change in fresh mass during treatment

could be measured. Each ramet was then separated into shoots (leaves plus stems) and

roots, dried at 70oC for 72h, and weighed. Root to shoot ratio (RSR) was calculated as

dry mass of roots divided by dry mass of shoots. In order to obtain an additional insight

into the effect of nutrient treatments on plant performance, we also calculated the increase

of biomass due to nutrient addition (Δ dry mass = mean average total dry mass in high

nutrients conditions – mean average total dry mass in low nutrients conditions). For this

calculation, the sum of SE of both nutrient treatments was used as variation.

Response to competition was calculated using a relative interaction index (RII,

Armas et al. 2014), (Bw - Bo) / (Bw + Bo), where Bw is the total final dry mass of the target

plant growing with another plant, and Bo is the total final dry mass of the target plant

growing alone. This index is symmetrically distributed around zero and ranges from -1 to

1; negative values indicate competition and positive values indicate facilitation. In order

to obtain variance values for the statistical analyses, replicates were obtained by randomly

matching plants growing alone and under competition.

2.5.Statistical analyses

Data were analyzed using one- and two-way ANOVAs with species of

Carpobrotus, soil nutrient availability (low or high), and region of collection as fixed

effects depending upon the analysis. Differences were considered statistically significant

at P < 0.05. Initial fresh mass was used as co-factor in ANOVAs, except for Δ fresh mass,

Δ dry mass and RII. Preliminary analyses were made with the Kolmogorov-Smirnov and

Levene tests, the Ln(x) transformation was used when required to meet the normality

assumptions. This transformations are indicated in the ANOVA results on the figures. All

the obtained RII values were significantly different from 0, according to Student's t-test

(P < 0.05). When an effect with more than two levels was significant (P < 0.05), a Tukey


142

test was used to detect differences among individual means. A total of 14 plants died

during the experiment and were not included in the data analyses, as indicated by the error

degree of freedom of the ANOVA. Analyses were conducted with the software IBM

SPSS Statistics, version 23 (IBM Corp., Armonk, New York, USA).

3. Results

3.1.Target-area approach: phenotypic plasticity

Nutrients significantly increased the Δ fresh mass, total mass, root mass, shoots

mass, and RSR both in the invasive C. edulis, in the exotic non-invasive C. chilensis and

in the invasive C. acinaciformis (Fig. 3). Total mass and shoot mass were significantly

greater in the invasive C. edulis than in the exotic non-invasive C. chilensis from

Californian populations (Fig. 3d). On the contrary, Δ fresh mass, total mass, and shoots

mass were significantly greater in C. acinaciformis than in C. edulis from Iberian

populations (Fig. 3f,g,i). Interestingly, the interaction effect of ‘species’ and ‘nutrient’

factors was significant for Δ fresh mass. This is, while C. edulis and C. acinaciformis

showed no differences while growing under low nutrient conditions, the Δ fresh mass was

significantly greater in C. acinaciformis than in C. edulis in the high nutrient treatment

(Fig. 3f). Benefits from nutrient addition, estimated as Δ dry mass, significantly differed

between the invasive C. edulis and the exotic non-invasive C. chilensis in Californian

populations (Fig. 4a), and it was also significantly greater for C. acinaciformis than for

C. edulis in Iberian populations (Fig. 4b).

3.2.Target-area approach: competitive ability

The relative interaction index (RII) was compared between the invasive C. edulis

and the exotic non-invasive C. chilensis from Californian populations, and between C.

edulis and C. acinaciformis from Iberian populations. The RII values obtained for C.

chilensis were significantly higher than those obtained for C. edulis from California (Fig.

5a). The effect of ‘species’ on RII was not significant either for C. edulis in competition

with C. acinaciformis or between C. edulis and C. acinaciformis when competing with

the native A. arenaria (Fig. 5b,c).

Chapter IV

143

Fig. 3. Response (mean + SE) to soil nutrient availability in Carpobrotus edulis and C. chilensis

(widespread non-invasive introduced species) from California (CA); and in C. edulis (widespread invasive

introduced species) and C. acinaciformis (less widespread invasive introduced species) from Iberia (IB): a,

f -- net change in fresh mass; b, g -- final total dry mass; c, h, -- final dry root mass; d, i -- final dry shoot

mass; e, j -- final dry root / shoot mass (RSR). Tables in graphs show results of ANOVAs (F, P) for effects

of initial fresh mass (I, covariable), species (S), nutrients (N), and species x nutrients (X); d.f. was 1 for

each factor and residual error was 33 for CA and 35 for IB. Significant results (P < 0.05) are in bold. Data

transformation, if any, is indicated on top of each data set.


144

3.3.Inter-range approach: phenotypic plasticity

Nutrients significantly increased the Δ fresh mass, total mass, root mass, shoots

mass, and RSR of C. edulis both in the native (South Africa) and in the non-native range

(Iberia, California and Australia) (Fig. 6). Similarly, ‘region’ factor significantly affected

all the studied variables except RSR. Thus, Δ fresh mass, total mass, root mass, and shoots

mass were greater in populations from the non-native range (Iberia, California and

Australia) in comparison with populations of C. edulis from the native range (South

Africa) (Fig. 6). Additionally, the benefit of nutrient addition, estimated as Δ dry mass,

was significantly greater in populations of C. edulis from the non-native range (Iberia,

California and Australia) than in populations form the native range (South Africa) (Fig

7).

Fig. 4. Δ dry mass (mean + SE) of the invasive C. edulis from California (CA) and the exotic non-invasive

C. chilensis from California (CA) (a); and the invasive C. edulis from Iberia (IB) and the exotic less invasive

C. acinaciformis from Iberia (IB) (b). Tables in graphs show results of ANOVAs (F, P) for the effect of

species factor; significant results (P < 0.05) are in bold.

3.4.Inter-range approach: competitive ability

The relative interaction index (RII) was compared between the invasive C. edulis

from the native range (South Africa) and C. edulis from the non-native range (Australia)

in competition with the Australian native C. virescens (Fig. 8a). Similarly, RII was

compared between native population of C. edulis (South Africa) and populations of C.

edulis from non-native range (Iberia) when competing the native A. arenaria from Iberia

(native range) (Fig. 8b). For both comparisons the effect of ‘species’ on RII was not

statistically significant (Fig. 8).

Chapter IV

145

Fig. 5. Competitive response (relative interaction index, RII; mean + SE) of (a) introduced Californian

(CA) Carpobrotus edulis (invasive) and chilensis (non-invasive) to each other; (b) introduced Iberian (IB)

C. edulis (widespread invasive) and C. acinaciformis (less widespread invasive) to each other; and (c)

introduced Iberian C. edulis and C. acinaciformis to the native, co-occurring dominant grass Ammophila

arenaria. Tables in graphs show results of ANOVAs (F, P) for the effect of species factor; significant

results (P < 0.05) are in bold. Data transformation, if any, is indicated on top of each data set.

4. Discussion

Elucidating the mechanisms behind plant invasions is a core question in modern

ecology (Richardson & Pyšek 2008). Here, we aimed to determine how key traits for plant

performance, as phenotypic plasticity and competitive ability, could be explaining the

invasive success of the clonal C. edulis. With this aim, an experimental design with a

double approach was employed: comparison between invasive and exotic non-invasive

congeners of Carpobrotus (target-area approach), and comparison between population of

C. edulis from the native and non-native range (inter-range approach). Both approaches

have been described as a suitable methodology to reveal the underlying mechanisms for

plant invasiveness (van Kleunen et al. 2010). The comparison between introduced species

differing in invasiveness seems to be a straightforward method to detect traits involved

in plant invasiveness, and using congeners could be specially suitable as inter-specific

differences are somewhat reduced. The expected outcome derived from this approach is

that those traits favoring invasiveness should be more evident in invasive than in exotic

non-invasive species (van Kleunen et al. 2010). In this study, we compared the effect of

soil nutrient availability on biomass partitioning and plant growth between populations

of the invasive C. edulis and the exotic non-invasive C. chilensis from California.

According to our prediction, C. edulis showed a greater benefit from nutrient addition, in

terms of increase in its total biomass, than C. chilensis. This denotes that efficiency in

acquisition of soil-based resources could be an important trait explaining differences in


146

Fig. 6. Response (mean + SE) to soil nutrient availability in Carpobrotus edulis from its native range in

South Africa and three regions in its introduced range: a -- net change in fresh mass; b -- final total dry

mass; c -- final dry root mass; d -- final dry shoot mass; e -- final dry root / shoot mass (RSR). Tables in

graphs show results of ANOVAs (F, P) for effects of initial fresh mass (I, covariable), region (R), nutrients

(N), and region x nutrients (X); d.f. was 3 for region, 1 for nutrients and residual error was 57. Significant

results (P < 0.05) are in bold. Letters above bars indicate differences between means (Tukey test, P < 0.05)

within the high nutrient treatment (uppercase) and within the low nutrient treatment (lowercase). Data

transformation, if any, is indicated on top of each data set.

Chapter IV

147

invasiveness between these species. Similarly, we tested differences between the Iberian

populations of C. edulis and the congener C. acinaciformis growing under two levels of

nutrients availability. Contrary to our predictions, the benefit of nutrient addition was

significantly greater in C. acinaciformis than in C. edulis, considered as more invasive

due to its higher occurrence in the Mediterranean basin (Lambinon 1995; Suehs et al.

2001). However, our results showed that growth was significantly more accentuate in C.

acinaciformis than in C. edulis under high nutrient conditions. A plausible explanation

for this unexpected result is that the higher occurrence of C. edulis in comparison with C.

acinaciformis could be due to different introduction histories, and not differences in the

degree of invasiveness. In the Iberian Peninsula, Carpobrotus spp. were introduced at the

beginning of the twentieth century for soil stabilization and as an ornamental plant in

gardens (Gonçalves 1990; Campoy et al. 2018). However, C. edulis has been more

intensely planted than C. acinaciformis and as consequence its occurrence is higher

(Gonçalves 1990). In this sense, there is a lag period between the introduction of an exotic

species in the new environment and its impact (Larkin 2012). This is, C. acinaciformis

might have been introduced less extensively than C. edulis (Gonçalves 1990), and this

could be the reason why it is considered less problematic (Suehs et al. 2001). From our

results we can infer that target-area approach also contains limitations, and using

congeners with similar introduction history is desirable to avoid confounding

interpretations. Interestingly, the greater capacity of C. acinaciformis to gain benefits

from soil nutrient enrichment that we have detected in our study could be indicative of a

high potential for expansion in this species that should be considered in order to predict

the scene of future invasions.

When testing the effects of soil nutrient availability by comparing plant growth

between C. edulis populations from the native (South Africa) and the non-native range

(Iberia, California and Australia), results showed a significant benefit of nutrients

addition, especially for those populations from the non-native range. Common garden

experiments comparing plant traits between populations from native and non-native range

are considered a suitable methodological approach to detect the presence of selection

pressures during the invasion process, and consequently to disentangle the causes of plant

invasiveness (Hierro et al. 2005; van Kleunen et al. 2010). Our results suggest that

nutrient use-efficiency has been positively selected in C. edulis populations at non-native

range, indicating a probable event of rapid evolution during the invasion.


148

Previous studies have reported

adaptive selection of traits in

populations of C. edulis at the invaded

range. Roiloa et al. (2016) found that the

benefit of physiological integration (i.e.

capacity for resources sharing between

connected modules in clonal plants) was

significantly higher in apical ramets of

C. edulis from Iberian populations

(invaded range) than in populations

from South Africa (native range),

suggesting that this clonal trait may

have been subjected to evolutionary

adaptation in the non-native range. Also, Portela et al. (2019) demonstrated that biomass

partitioning in response to nutrient availability differed between native and invaded range

populations of C. edulis, indicating that this trait could be under selection and favor

invasiveness. Plant reaction to resources availability is a well-known example of

phenotypic plasticity. Thus, as stated by the theory of optimal biomass allocation, plants

can modify their biomass partitioning patterns in order to favor the production of the

structures responsible for acquire the most limiting resource (Bloom et al. 1985; Gleeson

& Tilman 1992). In contrast with this prediction, our results did not reflect an increase in

the proportional biomass allocated to roots in the low nutrient treatment. This response

was similar for all the species and regions studied, denoting that plasticity in biomass

partitioning, at least in this case, cannot explain the invasiveness of C. edulis, in contrast

with the findings from Portela et al. (2019). This is, although our results found a greater

plant growth of C. edulis from the invaded range than in plants from the native range, due

to nutrient addition, this result was not apparently motivated by changes in biomass

partitioning, and we should explore whether another physiological mechanism could

explain the higher nutrient use-efficiency detected in C. edulis at the invaded range.

Plastic responses of plants to changing environmental conditions and level of resources

can be suggested to play a major role to explain plant invasiveness (Parker et al. 2003;

Richards et al. 2006; Lande 2015). In particular, changes in patterns of biomass

partitioning are expected to increase nutrient use-efficiency, favoring colonization of new

or changing environments. Plants respond to soil nutrient scarcity by increasing the

Fig. 7. Δ dry mass (mean + SE) of C. edulis plants from

the native (SA: South Africa) and non-native range (IB:

Iberia, CA: California, AU: Australia). Values are mean

+ SE. Letters on the bars indicate differences between

regions found by the Tukey test. Table in graph shows

results of ANOVA (F, P) for the effect of region factor;

significant results (P < 0.05) are in bold.

Chapter IV

149

proportional biomass allocated to produce roots; on the contrary, typical response to

nutrient rich soil is a decrease in the proportional production of roots (Bloom et al. 1985).

Thus, biomass allocation to root production would buffer soil nutrient scarcity, and

proportional reduction of root biomass under high nutrient condition would increase net

carbon gain, and consequently plant growth.

Fig. 8. Competitive response (relative interaction index, RII; mean + SE) of native South African (SA) and

introduced Australian (AU) or Iberian (IB) Carpobrotus edulis to species native to its introduced range: a)

C. virescens; b) Ammophila arenaria. Tables in graphs show results of ANOVAs (F, P) for the effect of

species factor.

Competition for resources is considered one of the most important ecological

interactions for plant performance, and particularly affecting the invasive potential of

exotic species (Burke & Grime 1996; Gioria & Osborne 2014). In this sense, the evolution

of increased competitive ability hypothesis (EICA), one of the most important hypotheses

formulated to explain biological invasions, state that release or reduction of natural

enemies at the introduced range will favor selection of individuals that allocate less

energy to defense and more to growth, enhancing its competitive ability in new

environment (Blossey & Nötzold 1995; Callaway & Ridenour 2004). Our results did not

found differences in competitive ability, determined by the relative interaction index

(RII), between C. edulis from the native (South Africa) and the introduced range (Iberia

and Australia) when competing with native plants (A. arenaria in Iberia, and C. virescens

in Australia). This result suggests that competitive ability has not been selected during

the process of invasion in the case C. edulis, at least for the studied populations. Previous

studies have found contradictory results with the EICA hypothesis (Vilà et al. 2003;

Bossdorf et al. 2004; Willis et al. 2010; Felker-Quinn et al. 2013). A proposed explanation

is that, if competition in the new habitat is lower, there will be no positive selection of


150

competitive ability in the introduced species, since this would not lead to fitness

improvement (Bossdorf et al. 2004). Significant differences in competitive ability were

neither found when we compared C. edulis and C. acinaciformis competing between them

or with a native species (A. arenaria). Both species co-occur at the same coastal habitats

in Iberia, and consequently can establish competitive interaction between them and with

native species inhabiting these areas. However, our study reveals no differences in

competitive ability between both exotic species, and probably indicates that suggested

distinctions in invasive potential between both species (Lambinon 1995; Suehs et al.

2001) could be due to different invasions histories instead to differences in their

invasiveness. Results also reported that the invasive C. edulis was not more competitive

than the co-occurring exotic non-invasive C. chilensis. On the contrary, while competition

reduced plant growth in C. edulis, results showed that when both species grew together

C. chilensis experienced an increase in growth. This is, performance of C. chilensis was

facilitated by the presence of C. edulis. Previous studies have showed that C. edulis

intensively modify soil conditions, creating a layer of organic material (Novoa &

González 2014) that could be favoring the growth of C. chilenesis. However, whether this

is the cause for this is an unexpected result is unresolved, and future studies should be

conducted to elucidate the reasons.

Concluding remarks

This study suggests the presence of adaptive evolution during the process of

invasion of the clonal C. edulis. Results show that populations from the introduced range

(Iberia, California, and Australia) significantly increased their growth in response to soil

nutrient addition, in comparison with populations from the native range (South Africa).

This benefit in plant growth was not motivated by plastic changes in biomass partitioning,

and other physiological mechanisms should be implicated to explain the differences in

nutrient use-efficiency between populations of C. edulis from the native and invaded

range. However, these differences in plant growth were not related with a higher capacity

for competition in non-native range populations. This study also compared plasticity in

biomass allocation, plant growth, and competitive ability between different congeneric

Carpobrotus species differing in invasiveness. It was found a greater benefit from nutrient

addition, in terms of increase in its total biomass, in C. edulis from California than in the

less invasive congener C. chilensis, suggesting that plastic response to soil nutrient

Chapter IV

151

content could explain potential differences in plant invasiveness between these species.

On the other hand, contrary to our hypothesis, it was also found a greater benefit from

nutrient addition in C. acinaciformis than in C. edulis. The relevance of competitive

ability in coastal habitats invaded by C. edulis needs to be further studied to elucidate the

role this trait plays in biological invasions of this species.

Acknowledgments

Authors are thankful to A. Novoa, R. Bermúdez-Villanueva, P. Yeoh and K. Batchelor

for assistance with the collection of plants, L. Álvarez from the Spanish Ministry of

Agriculture, Food, and the Environment for assistance in authorizing the import of plants

to Spain, and J. Sones and the University of California Natural Reserve System for access

to the Bodega Marine Reserve. Australian samples were collected under licences

SW018396 and CE005442 issued by the Western Australian Department of Parks and

Wildlife. This work was supported by the Spanish Ministry of Economy and

Competitiveness (Grant CGL2013-44519-R awarded to S. R. R.), co-financed by the

European Regional Development Fund (ERDF), and by a CSIRO Julius Career award (to

B. L. W.). This is a contribution from the Alien Species Network (Ref. ED431D 2017/20

– Xunta de Galicia, Autonomous Government of Galicia).


152

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156

Supplementary material

Table S1. Sampling sites for the different species used in the experiment. World region, site location, and

coordinates are included. For each species and region the invasive status is also included. See Fig. 2 for

map representation.

Region Species Status Location Coordinates

Northern

California

C. edulis Invasive introduced

Bodega Bay 38o 19' 18'' N, 123o 04' 16'' W

Abbotts Lagoon, North 38o 07' 42'' N, 122o 57'14'' W

Abbotts Lagoon, South 38o 06' 26'' N, 122o 57' 32'' W

Pescadero 37o 16' 23'' N, 122o 24' 34'' W

C. chilensis Non-invasive

introduced

Bodega Bay 38o 19' 18'' N, 123o 04' 16'' W

Abbotts Lagoon, North 38o 07' 42'' N, 122o 57'14'' W

Abbotts Lagoon, South 38o 06' 26'' N, 122o 57' 32'' W

Pescadero 37o 16' 23'' N, 122o 24' 34'' W

North-West

Iberian Peninsula


Grove 42o 28' 18'' N, 8o 51' 25'' W

Caminha 41o 51' 11'' N, 8o 51' 57'' W

Castelo do Neiva 41o 37' 02'' N, 8o 48' 40'' W

Quiaios 40o 13' 31'' N, 8o 53' 20'' W

C. acinaciformis Invasive introduced

Seselle 43o 25' 45'' N, 8o 13' 35'' W

A Coruña 43o 22' 53'' N, 8o 24' 37'' W

A Lanzada 42o 25’ 57'' N, 8o 52’ 25'' W

Limens 42o 15’ 04'' N, 8o 48’ 11'' W

A. arenaria Native Baldaio 43o 17' 58'' N, 8o 40' 10'' W

South Africa C. edulis Native

Kleinmond 34o 20' 22'' S, 19o 02' 07'' E

Hawston 34o 23' 28'' S, 19o 07' 35'' E

Fish Hoek 34o 07' 58'' S, 18o 26' 05'' E

Cape of Good Hope 34o 20' 26'' S, 18o 27' 34'' E

South-West

Australia


Henderson Cliffs 32o 10' 20'' S, 115o 46' 19'' E

Piara Waters 32o 8' 10'' S, 115o 55' 55'' E

Star Swamp 31o 51' 29'' S, 115o 45' 29'' E

C. virescens Native

Woodman’s Point 32o 08' 05'' S, 115o 44' 37'' E

Campbell Barracks 31o 57' 18'' S, 115o 45' 18'' E

Peasolm Beach 31o 54' 29'' S, 115o 45' 18'' E

Chapter IV – Supplementary material

157

Fig. S1. Diagram of the plant material used in the experiment. Populations are Iberian Peninsula (IB), South

Africa (SA), California (CA) and Australia (AU).


158

Fig. S2. Initial fresh biomass (mean + SE) of different species, and populations of C. edulis. Letters above

bars indicate differences between means (Tukey test, P < 0.05).

Section II

Alternanthera philoxeroides

Also starring Agasicles hygrophila


159

1. Biology and native distribution

Alternanthera philoxeroides (Mart.) Griseb (Phylum Magnoliophyta, Class

Magnoliopsida, Order Caryophyllales, Family Amaranthaceae, Subfamily

Gomphrenoideae), commonly known as alligator weed, is an amphibious, perennial herb

native to the Parana River basin in South America (Julien 1995). Native range includes

Argentina, Brazil and Paraguay (Anderson et al. 2016). It is a perennial plant, with

opposite and sessile leaves 2-7cm long and 1-2cm wide. It has white flowers 8-10mm in

diameter, arranged in hemispherical pseudo-spikes (Fig. 1). A. Philoxeroides has two

distinct ecotypes: when the plant grows in aquatic environments it has hollow stems that

give it buoyancy, while in terrestrial environments its stems lack that hollow (Lu & Ding

2010). When growing over water, the plant forms floating bushes and remains anchored

to the substrate by roots (Zuo et al. 2012).

Figure 1. On the left, detail of a flower of A. philoxeroides from the population of Fisterra (Galicia, NW

Spain). On the right, the plant invading a pond in the Wuhan Botanical Garden (Hubei Province, China).

The plant reaches 1m in height in terrestrial environments and 60cm in aquatic

environments. In the case of fragmentation of the floating bushes, the plant is dragged

downstream and establishes when making contact with the shore. A. philoxeroides

Section II

160

develops an extensive root system, which can have up to ten times more biomass than the

aerial structures (Schooler et al. 2008). The plant has the ability to resprout from its roots

in case of destruction of the aerial structures, due to mechanical removal, heavy herbivory

damage or death by freezing (see Fig. 5) (van Oosterhout 2007). In aquatic environments,

the plant grows forming an extensive monospecific tapestry, blocking irrigation channels

and covering waterways, transforming them into marshy environments (Sainty et al.

1998). A single bush can reach a length of 15m and become so robust they can support a

man. It has tolerance to salinity, supporting 10% seawater in stagnant water or up to 30%

in water flows.

1. Invaded range

Temperature is the main factor that limits the geographical distribution of A.

philoxeroides (Julien, Skarratt, & Maywald 1995). It is a tropical plant, its optimum

growth is at 30⁰C, while at low temperatures the vegetative development slows down.

There is no growth below 7⁰C or sprouts below 5⁰C (Shen et al. 2005). Frost destroys

leaves and stems, but roots are able to survive. There are naturalized populations of A.

philoxeroides in many countries, most of which have triggered biological invasions with

severe environmental consequences (Fig. 2). The first mention of the plant in the US dates

from 1897 in the south of the country, from where it has expanded to more than a dozen

states (Zeigler 1967). It was introduced in New Zealand in 1906 (Roberts & Sutherland

1989) and in Australia in the 1940s (Julien & Bourne 1988). In Asia it is naturalized in

Indonesia, Thailand, Sri Lanka (Anderson et al. 2016), India (Pramod et al. 2008), Japan

(Kusumoto et al. 2011) and China (Wang et al. 2005). The first mention of A.

philoxeroides in Europe occurred in France in 1971, on the Garonne River (Fig. 3)

(Dupont 1984). In Italy the species was established in the Arno River and later in the

Tevere River (Iamonico & Sánchez-Del Pino 2016). In Spain there is a naturalized

population of small size in Fisterra, Galicia, where it has been present for at least a decade

(Romero & Amigo 2015). The origin of this population are nearby abandoned

greenhouses. Although the plants were completely uprooted in the past years, they have

regrown and the size of the population is increasing.

Figure 2. Distribution map of A. philoxeroides worldwide. Map obtained from GBIF in May, 2019.

Figure 3. Distribution map of A. philoxeroides in Europe. Map obtained from GBIF in May, 2019.


163

2. Invasiveness

The main factor that makes A. philoxeroides an aggressive invader is its rapid

growth. In its native range, A. philoxeroides is capable of reproducing both sexually and

asexually (Julien et al. 1995; Sainty et al. 1998). It has been previously stated that

reproduction of this species is exclusively asexual in those areas where the plant has been

introduced (van Oosterhout 2007). A. philoxeroides shows an extremely low genetic

variability in China, a country where the plant has proliferated successfully during the

last decades, so it is suspected that clonal growth is the likely explanation for the success

of this invasive plant (Wang et al., 2005; Ye et al. 2003). However, contamination by

seeds of A. philoxeroides on soil sourced from China has been reported, so it seems that

the plant is still able to produce viable seeds in the invaded range (Anderson et al. 2016).

The vegetative propagation of the species occurs through stems or taproots. When

developing in waterways, fragmentation of the floating bushes is beneficial for the plant,

allowing the rapid colonization of the downstream watercourse. This favors the rapid

expansion over a wide area once the species has been established. The capacity of A.

philoxeroides to regenerate from its taproots is remarkable (Fig. 5) (van Oosterhout

2007). Fragments of a few centimeters are able to sprout in moist soil. This is one of the

reasons why the eradication of A. philoxeroides is difficult to achieve through physical

control, since small root fragments allow the plant to resprout once it has been eliminated.

In marshy areas, it is almost impossible eliminate all plant fragments from the substrate

during its removal.

One of the reasons for the widespread expansion of A. philoxeroides is its

similarity to Alternanthera sessilis (L.) R.Br. A. sessilis, commonly known as sessile

joyweed, is an aquatic plant, similar in appearance to A. philoxeroides. The difference

between both species is that A. sessilis lacks petioles in its flowers, which grow directly

over the stems (hence the name of the species). It is a widely distributed species in tropical

regions of Asia and Oceania, also introduced in other countries but not considered

invasive (Fig. 4). The plant has culinary utility in Southeast Asia (particularly in Sri

Lanka), besides being used in traditional medicine, as an ornamental plant and also in

aquariums (Grubben & Denton 2004; Hossain et al. 2014; Walter et al. 2014). In Australia

and New Zealand, awareness campaigns on both species have been conducted to avoid

the inadvertent propagation of A. philoxeroides (Sainty et al. 1998).

Section II

164

Figure 5. A. philoxeroides has a great capacity for regeneration from its taproots. Dry taproot fragments

(left) and plants sprouting a week later (right). Photographs taken at the Universidade Federal São Carlos

(São Carlos, Brazil).

3. Ecological impact

A. philoxeroides is considered one of the worst aquatic weeds worldwide due of

its quick growth and persistence (Wang et al. 2008). In China, the plant invades

abandoned farmlands, waterways and rice fields. When growing in crop fields, A.

philoxeroides reduces yields for various crops, including rice, potato, wheat and corn

(Mehmood et al. 2017; Yi 1992). In several countries, it successfully competes with

native species and displaces them, decreasing the biodiversity and species richness

(Chatterjee & Dewanji 2014; Guo & Wang 2009). It is capable of displacing native

species, leading to monospecific communities where other exotic species subsequently

proliferate. Its capacity to convert watercourses into swamps, and to block irrigation

channels causing floods, creates optimal environments for mosquito breeding, which is

problematic in countries of South Asia where malaria and similar diseases are present

(Sainty et al. 1998). This has serious consequences for human health and also for

livestock. When growing in aquatic environments, biomass decomposition of A.

philoxeroides alters the nutrient cycle of the water, which benefits the entry of other exotic

species into the ecosystem (Bassett et al. 2010). In addition, it also alters the natural

distribution of aquatic invertebrate communities (Bassett et al. 2012). Dense populations


165

growing in water may decrease the content of dissolved oxygen in it (Quimby & Kay

1976).

5. Legal status

In the US, A. philoxeroides has varying classifications at a federal or state level;

it is classified as noxious weed in Alabama, Arizona, Arkansas, California, Florida, South

Carolina and Texas (http://plants.usda.gov/core/profile?symbol=ALPH). In Australia, A.

philoxeroides has been declared noxious weed in all states; government departments are

compelled to control and/or eradicate the species. In New Zealand, A. philoxeroides is

listed as an unwanted organism under the Biosecurity Act (1993) and it is included on the

National Pest Plant Accord List. This bans the sale, propagation and distribution of the

plant throughout New Zealand. In Europe, it is included in the list of invasive exotic plants

according to the EPPO (European and Mediterranean Organization for the Protection of

Plants)(Anderson et al. 2015). In Spain, A. philoxeroides is included in the Spanish

Catalog of Invasive Alien Species. It is classified as an invasive species in Royal Decree

630/2013, 2nd August.

6. Management

There are different methods to achieve the control and eradication of A.

philoxeroides, either physical, chemical or biological (Sainty et al. 1998). Physical

control is useful in early stages of invasion. It involves not only eliminating the aerial part

of the plants, but also their roots, to avoid outbreaks in later years. If the elimination of

the substrate is not possible, for example in marshy lands, it will be necessary to monitor

the population for several years to prevent its reappearance. Precautions should also be

taken when cleaning the machinery used, to avoid further spread of the invasion. The

downside of physical control is that it requires considerable effort, between 4 and 10 hours

of work per square meter (Clements et al. 2014). Fire is not a recommended option, since

roots of the plant survive and it causes a high impact on the ecosystem. Chemical control

is a good option to eliminate large areas occupied by the plant, which would otherwise

require a considerable elimination effort. A. philoxeroides is resistant to several

herbicides. Field tests have shown the effectiveness of glyphosate to fight the plant when

it grows in floating mats, although this has a negative impact on the native flora (Sainty

http://plants.usda.gov/core/profile?symbol=ALPH

Section II

166

et al. 1998). However, in terrestrial environments glyphosate does not completely

eliminate the underground parts of the plant. Dichlobenil or metsulfuron have proven to

be useful in this type of situation (Bowmer et al. 1989; Clements et al. 2014).

Biological control has proven to be the most efficient and effective way to control

populations of A. philoxeroides in advanced stages of invasion. It requires a minimum

cost and is a sustained treatment over time, but has certain limitations (Sainty et al. 1998).

The first program for the biocontrol of A. philoxeroides was carried out in the 1960s in

the US. Several states in the south of the country, particularly Florida, were extensively

affected by the plant, what entailed considerable economic losses. Predators from the

native range of the plant were identified, food preference tests were subsequently carried

out and three candidates were selected to be used as biological control agents: the flea

beetle Agasicles hygrophila Selman & Vogt, the moth Vogtia malloi Pastrana and the

thrips Amynothrips andersoni O'Neill (Buckingham 1996). Of these three, A. hygrophila

proved to cause massive damage to the plant, from which both larvae and adults feed, and

has been released in several countries, with some successfully control results (Lin, et al.

1984; Ma et al. 2003; Sainty et al. 1998).

The life cycle of A. hygrophila is closely linked to A. philoxeroides. The female

adults lay the eggs on the underside of the leaves, from which the larvae emerge. To

perform the metamorphosis, the third-stage larva makes a hole in a hollow stem of A.

philoxeroides, inside which the pupa forms (Maddox 1968). Once the imago emerges, it

feeds on both the leaves and the outside part of the stems of the plant. The adults have a

remarkable mobility, since they perform jumps of up to half a meter and are able to fly

several kilometers in search of food. Therefore, once introduced into a region, they

rapidly expand over successive generations. This insect feeds exclusively on A.

philoxeroides, also requiring the leaves for egg laying and the cavity of the stems to

perform the metamorphosis (Maddox 1968; Wang et al. 1988). The latter is critical, since

it prevents the life cycle of the insects from being completed in those populations of A.

philoxeroides with terrestrial ecotype, as they lack any cavity in their stems (Ma et al.

2003; Pan et al. 2011). However, both adult insects and larvae are able to feed on plants

regardless of their ecotype (Fig. 6).


167

Figure 6. Larvae (left) and adult (right) of A. hygrophila feeding on the leaves of A. philoxeroides.

Photographs taken at the Beijing Forestry University (Beijing, China).

A. hygrophila has been used as biocontrol of A. philoxeroides in many countries.

The main factor that limits the success of this control agent is temperature (Stewart 1996;

Stewart et al. 1999). Both species are native to tropical regions of South America, so they

develop optimally at relatively high temperatures. However, low temperatures kill insects

and prevent eggs from hatching. Furthermore, while frosts destroy the aerial part of the

plants, they can resprout from their roots the following spring. The easiest solution to this

problem is the reintroduction of the insect over several years (Buckingham et al. 1983).

In the US, the insect settled in the southern states of the country, but as the populations

moved northwards they encountered colder winters, which hindered their permanent

settlement. Nevertheless, in those areas where the insect was established, it has decimated

the populations of A. philoxeroides in such an extent that, even if the plant is not

completely eradicated, it no longer poses an economic or environmental threat

(Buckingham 1996).

Section II

168

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6664.2012.00443.x

Chapter V

Effects of physiological integration on defense strategies against

herbivory by the clonal plant Alternanthera philoxeroides

Rubén Portela1,2, Bi-Cheng Dong2 , Fei-Hai Yu2,3,4, Rodolfo Barreiro1,

Sergio R. Roiloa1

1BioCost Group, Biology Department, Universidade da Coruña, A Coruña 15071,

Spain.

2School of Nature Conservation, Beijing Forestry University, Beijing 100083, China.

3Institute of Wetland Ecology and Clone Ecology, Taizhou University, Taizhou 318000,

China.

4Zhejiang Provincial Key Laboratory of Plant Evolutionary Ecology and Conservation,

Taizhou University, Taizhou 318000, China.

Published as Portela, R., Dong, B. C., Yu, F. H., Barreiro, R., & Roiloa, S. R. (2019).

Effects of physiological integration on defense strategies against herbivory by the clonal

plant Alternanthera philoxeroides. Journal of Plant Ecology, 12(4), 662-672. doi:

10.1093/jpe/rtz004

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175

Abstract

The plant-herbivore relationship is one of the most fundamental interactions in nature.

Plants are sessile organisms, and consequently rely on particular strategies to avoid or

reduce the negative impact of herbivory. Here we aimed to determine the defense

strategies against insect herbivores in the creeping invasive plant Alternanthera

philoxeroides. We tested the defense response of A. philoxeroides to herbivory by a leaf-

feeding specialist insect (Agasicles hygrophila) and a polyphagous sap-feeding insect

(Planococcus minor). We also tested the mechanisms triggering defense responses of A.

philoxeroides by including treatments of artificial leaf removal and jasmonic acid

application. Furthermore, we examined the effect of physiological integration on these

defense strategies. The combination of artificial leaf removal and jasmonic acid

application produced a similar effect to that of leaf-feeding by the real herbivore.

Physiological integration influenced the defense strategies of A. philoxeroides against

herbivores, and increased biomass allocation to aboveground parts in the apical ramets

damaged by real herbivores. Our study highlights the importance of physiological

integration and modular plasticity for understanding the consequences of herbivory in

clonal plants.

Keywords: Agasicles hygrophila; Alternanthera philoxeroides; alligator weed; clonal

integration; herbivory; Planococcus minor.

Section II – Alternanthera philoxeroides

176

1. Introduction

Clonal growth in plants consists on the propagation through vegetatively produced

modules, named ramets, which can remain physically connected via stolons, rhizomes or

horizontally growing roots (Pitelka & Ashmun 1985; Klimes et al. 1997; Price & Marshall

1999; Xu et al. 2012). One of the most interesting traits associated with clonal plants is

the capacity for (clonal) physiological integration that allows the exchange of resources

and signals between connected ramets of clonal plants (Pitelka & Ashmun 1985).

Physiological integration is especially important for developing ramets or ramets located

in resource-poor patches, which are supported by resource translocation from established

ramets or ramets in resource-rich patches (Hartnett & Bazzaz 1983; Slade & Hutchings

1987; Jónsdóttir & Watson 1997; Alpert 1999; Saitoh et al. 2002; Roiloa & Retuerto

2006). Effects of physiological integration have been previously studied in a variety of

situations with heterogeneous distribution of environmental factors, including light,

nutrients, water, salinity, heavy metals, sand burial, wind erosion, pathogens and

defoliation (Salzman & Parker 1985; Schmid et al. 1988; D’Hertefeldt & van der Putten

1998; Yu et al. 2004, 2008; Roiloa & Retuerto 2012; You et al. 2014; Wang et al. 2017a,

b). Benefits of clonal integration commonly outweigh its potential costs, allowing clonal

plants to overcome stressful conditions and colonize a wide array of environments (Roiloa

& Retuerto 2012; Song et al. 2013; You et al. 2014; Roiloa et al. 2016; Wang et al. 2017a,

b).

The plant-herbivore relationship is one of the most fundamental interactions in

nature (Begon et al. 2014). Plants are sessile organisms, and consequently rely on

particular strategies to avoid or reduce herbivore attacks and/or impacts (Schoonhoven et

al. 2005). To study the impact of plant-herbivore interactions, one should also consider

the underlying mechanism triggering defense responses in plants. Induced defense

mechanisms allow plants injured by herbivores to activate defense responses, whereas

such defense responses will not be activated in the absence of herbivory (Karban &

Baldwin 1989; Agrawal 2000). Therefore, induced defense has been described as a

strategy to maximize the benefits and minimize the costs of defenses in environments

where herbivory is variable (Bråthen et al. 2004). Simulated leaf removal is a commonly

used approach to study consequences of herbivory. However, plant responses to

herbivores are complex and can also be activated by other mechanisms (Baldwin 1990).

For instance, the presence of the salivary fluid secretion of herbivores may activate

Chapter V

177

defense responses of plants (Felton & Eichenseer 1999), but leaf removal by itself may

not (Baldwin 1990; Felton & Eichenseer 1999). Also, the exogenous application of

jasmonic acid or its derivate, methyl jasmonate, may trigger the same defensive responses

as those caused by real herbivores (Thaler et al. 1996; Baldwin 1998). Thus, the use of

leaf removal and jasmonic acid may help separate mechanical and chemical effects when

studying induced defense to herbivory (Baldwin 1996; Agrawal et al. 1999), and the

combined application of both treatments may represent a scenario more similar to real

herbivory (van Kleunen et al. 2004).

While many studies have examined the effects of clonal integration on responses

of clonal plants to both abiotic and biotic factors (Roiloa & Retuerto 2012; Song et al.

2013; Roiloa et al. 2016; Wang et al. 2017b), including those to simulated herbivory

(Schmid et al. 1988; You et al. 2014; Wang et al. 2017a), little is known about its effects

on the responses of clonal plants to real herbivory (but see Gómez et al. 2007, 2008).

Schmid et al. (1988) reported a benefit of clonal integration in defoliated ramets in a

simulated herbivory experiment. Similarly, Wang et al. (2017a) found that clonal

integration increased the tolerance to heavy defoliation in a rhizomatous plant. The role

of clonal integration in generating induced defense to herbivores has also been tested

(Gómez & Stuefer 2006; Gómez et al. 2007, 2008). These studies provide evidence that

induced systemic resistance (induced defense developed at a non-local scale) is mediated

by clonal integration, which can mitigate the damage suffered by defoliated ramets

(Gómez & Stuefer 2006; Gómez et al. 2007, 2008). However, no previous study has

assessed the mechanisms triggering the defense responses against herbivory and how

these responses are affected by clonal integration.

We conducted a greenhouse experiment to test the effect of physiological

integration on the defense responses to herbivory in the clonal plant Alternanthera

philoxeroides. We compared the difference in growth and biochemical responses of A.

philoxeroides attacked by a leaf-feeding specialist insect (Agasicles hygrophila) and a

polyphagous sap-feeding insect (Planococcus minor). In addition, it was examined

whether the induced response to herbivory was triggered simply by mechanical damage

(leaf removal) or by a combination of mechanical damage and signaling hormones

(jasmonic acid). Finally, we tested the role of clonal integration in the defense strategies

of A. philoxeroides against herbivores. Specifically, we hypothesized (1) that the leaf-

feeding specialist insect will provoke a more negative impact on plant growth than the


178

generalist sap-feeding herbivore, (2) that the combination of leaf removal and jasmonic

acid application will be necessary to activate the defensive response, and (3) that the

negative impact of herbivory will be alleviated by physiological integration.

2. Materials and methods

2.1.Study species

Alternanthera philoxeroides (Mart.) Griseb. (Amaranthaceae), commonly known

as alligator weed, is an amphibious, perennial herb native to the Parana River basin in

South America (Julien et al. 1995). The geographical distribution of the species out of its

native range includes the south of North America, Australia, New Zealand, Indonesia,

Thailand and China (Julien et al. 1995; Sainty et al. 1998). A. philoxeroides is highly

invasive and considered one of the worst weeds in the world (Wang et al. 2008; Yu et al.

2009). This species can grow both on land and in water, and form floating mats in water

courses. A. philoxeroides in China shows extremely low genetic diversity and does not

produce viable seeds (Lu et al. 2013). Instead, it reproduces clonally by stem and/or root

fragments (Julien et al. 1995; Sainty et al. 1998; Yu et al. 2009; Dong et al. 2010). It can

displace native species, block irrigation systems, and increase the risk of floods (Julien &

Bourne 1988). Over time, water bodies become swamps covered by the plant (Sainty et

al. 1998).

Agasicles hygrophila Selman and Vogt (Coleoptera: Chrysomelidae), commonly

known as alligator weed leaf beetle, is a species native to South America (Maddox 1968).

The larvae and adult of A. hygrophila both feed on leaves and stems of A. philoxeroides,

and eventually cause it death. The species has been introduced to North America, China

and New Zealand as an agent to control A. philoxeroides, with excellent results in North

America but not always successful in other regions (Burgin et al. 2010; Lu & Ding 2011).

Planococcus minor Maskell (Hemiptera: Pseudococcidae), is a polyphagous pest

native to Asia (Cox 1989). Its host range is wide, including more than 250 species in 80

families (Venette & Davis 2004). P. minor has a short life cycle, a high reproductive rate,

and a polyphagous nature (Francis et al. 2012). The species is considered a pest in India

and China, and a potential risk in USA (Venette & Davis 2004).

Chapter V

179


All the plant material used in this experiment was collected in Zhejiang Province,

China. A total of 150 clonal fragments, each consisting of six connected ramets and a

stem apex, were cut off from stock plants. The two oldest ramets (hereafter called “basal

ramets”) of clonal fragments were placed in a container with water to promote root

formation. The other four youngest ramets (hereafter called “apical ramets”) remained

outside the water. After 15 days, seven groups of plants, each containing 12 clonal

fragments of similar size, were selected. Each fragment was placed in two adjacent pots,

with the two basal ramets in one pot and the remaining four apical ramets in the adjacent

pot. All the pots with apical ramets were enclosed by insect cages (25 cm long × 25 cm

wide × 50 cm high).

The experiment consisted of six levels of herbivory crossed with two levels of

clonal integration (with or without), making a total of 12 treatments (Fig. 1). There were

seven replicates for each treatment. To minimize differences due to possible

environmental patchiness in the greenhouse we used a block design with seven blocks.

Each block contained one replicate of each treatment and one group of the 12 plants

(clonal fragments) of A. philoxeroides. The 12 plants within each group were randomly

assigned to the 12 treatments in one block. The six herbivory treatments were a control

(no real or simulated herbivory), two treatments of real herbivory (herbivory by the

specialist herbivore A. hygrophila and herbivory by the generalist P. minor) and three

treatments of simulated herbivory (jasmonic acid application, artificial leaf removal and

both jasmonic acid application and artificial leaf removal). For the treatments without

clonal integration, the apical and basal parts of a clonal fragment were disconnected by

cutting off the stem internode connecting them, and for the treatments with clonal

integration, they were left connected.

We released four adults of A. hygrophila into each cage in the specialist herbivory

treatment, and put 15-25 adults of P. minor on the stems of the apical ramets in each cage

in the generalist herbivory treatment. In the leaf removal treatment, we artificially

removed 50% of the leaves that were over 2 cm long in the apical ramets. This treatment

setup has been previously described as an accurate mean of simulated herbivory damage

of A. philoxeroides by A. hygrophila (Schooler et al. 2006). In the jasmonic acid

application treatment, we sprayed the apical ramets with a 100 µmol L-1 jasmonic acid

solution (J2500-100MG; Sigma-Aldrich, St. Louis, Missouri, USA) with a 1:1 mixture of


180

95% ethanol and distilled water as the solvent. The target ramets in the jasmonic acid

application treatment were covered by a plastic bag until the solution dried out to prevent

it from affecting neighbor plants due to its volatility. In the treatment with both jasmonic

acid application and leaf removal, we removed 50% of the leaves of the apical ramets

which were over 2 cm long, and then sprayed them with the jasmonic acid solution. Leaf

removal and jasmonic acid treatments were applied once a week.

Figure 1. Schematic representation of the experimental treatments consisting of two crossed factors with

herbivory (control, leaf-feeding Agasicles hygrophila, sap-feeding Planococcus minor, jasmonic acid, leaf

removal, jasmonic acid + leaf removal) and clonal integration (connected, disconnected) as main factors.

The experiment was carried out in a greenhouse belonging to Forest Science

Company, Ltd., of Beijing Forestry University in Beijing, China. Plants were grown

under a natural photoperiod and watered regularly to avoid water stress. During the

experiment, the air temperature in the greenhouse was 27.5 ± 0.3°C, and the relative

humidity 81.3 ± 1.5% (mean ± SE), measured daily with Hobo Temp/RH loggers (HOBO

UX100-003; Onset Computer Co., Bourne, MA, USA). Treatments began on July 10,

2016 and continued for five weeks.

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2.3.Measurements

At harvest, number of ramets of the apical part and the basal part were counted

separately. The apical part and basal part were each divided into leaves, stems and roots,

dried at 75ºC for 48 h, and then weighed. We also measured concentrations of total

phenolics and condensed tannins of leaf samples. The method for determining total

phenolics using Folin Ciocalteu reagent was adapted from Mcdonald et al. (2001); the

method for measuring tannins was the acid butanol assay described by Gessner & Steiner

(2005). A total of 50 mg of dried, ground leaf material was needed for each measurement.

Due to lack of sufficient leaf material, only five replicates were used for these

measurements.

2.4.Data analysis

We used three-way ANOVA to test effects of block, clonal integration, herbivory

and clonal integration × herbivory on total mass, leaf mass, stem mass, root mass, number

of ramets and root to shoot ratio of the apical part, the basal part and the clonal fragment

(apical plus basal part). We used two-way ANOVA to examine effects of clonal

integration, herbivory and their interaction on total phenolics and tannins. All data were

checked for normality and homoscedasticity. Where necessary, data were transformed to

meet the requirements of ANOVA. As a consequence, the square root transformation was

applied to the following variables: leaf mass, root mass and root to shoot ratio of the

clonal fragment, stem mass, root to shoot ratio and tannins of the apical part, and number

of ramets, leaf mass, root mass, stem mass and root to shoot ratio of the basal part. When

the effects were significant, we applied a posterior Tukey test to detect differences among

the six herbivory treatments, and t-test for differences between the connected and

disconnected treatments within each herbivory treatment.

Mortality during the experiment affected one apical part from the disconnected

treatment with herbivory by A. hygrophila. Two of the basal parts did not develop, i.e.

one from the connected, control treatment and the other from the disconnected treatment

with herbivory by A. hygrophila. Dead or undeveloped plants were not included in the

analyses. Significance levels were set at p = 0.05 after Bonferroni correction. All analyses

were conducted with IBM SPSS, version 23 (IBM Corp., Armonk, NY).


182

3. Results

3.1.Performance of basal ramets

Connection with the apical ramets significantly increased both root mass and root

to shoot ratio of the basal ramets when the apical ramets were subjected to herbivory by

A. hygrophila, and also increased root to shoot ratio of the basal ramets when the apical

ramets were subjected to jasmonic acid application plus leaf removal (Fig. 2e, f, Table 1).

Connection significantly reduced the tannin concentration in the basal ramets when the

apical ramets were treated with jasmonic acid (Fig. 5, Table 2). However, connection had

no significant effect on other growth estimates or on the concentration of phenolics of the

basal ramets (Tables 1 and 2, Figs. 2 and 5). In addition, herbivory of the apical ramets

had no significant effect on the growth or physiology of the basal ramets (Tables 1 and 2,

Figs. 2 and 3).

3.2.Performance of apical ramets

Herbivory significantly affected all growth measures and root to shoot ratio of the

apical ramets (Table 1). Compared to the control, herbivory by A. hygrophila, leaf

removal, and jasmonic acid application plus leaf removal generally reduced biomass of

the apical ramets (Fig. 3a-e). Connection with the basal ramets significantly increased

root to shoot ratio of the apical ramets, especially when the apical ramets were attacked

by A. hygrophila and P. minor (Fig. 3f). Herbivory of the apical ramets significantly

affected the tannin concentration of the apical ramets (Table 2). The apical ramets with

jasmonic acid application showed the lowest tannin concentration (Fig. 5d).

3.3.Performance of clonal fragments

Herbivory of the apical ramets significantly affected leaf, stem, root and total mass

of the whole clonal fragment (apical plus basal ramets; Table 1). Compared to the control,

biomass of the clonal fragment reduced when the apical ramets were subjected to

herbivory by A. hygrophila, leaf removal, and jasmonic acid application plus leaf removal

(Fig. 4a, c-e). Root to shoot ratio of the clonal fragment was also affected by herbivory

(Table 1), showing the highest value when the apical ramets were attacked by the

specialist herbivore A. hygrophila (Fig. 4f).

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4. Discussion

We compared the growth and physiological responses of the invasive plant A.

philoxeroides between two types of aboveground herbivory, by the leaf-feeding specialist

A. hygrophila and the polyphagous sap-feeding herbivore P. minor. As we predicted, the

negative impact on apical ramets was stronger when A. philoxeroides was consumed by

A. hygrophila than by P. minor. This is because the two insect species caused different

kinds of damage to the host plant. Both adults and larvae of A. hygrophila devoured the

leaves and external part of the stems, resulting in lower total, leaf and stem mass

compared to the control. Damage was smaller in plants attacked by P. minor whose

feeding system sucks the phloem. The reduction of photosynthetic structures caused by

the leaf-feeding insect will reduce the net carbon gain of A. philoxeroides, as its

photosynthetic capacity will be reduced but respiration in other tissues will be maintained.

On the other hand, the sap-feeding insect is not expected to damage photosynthetic

structures so that plant growth will be less affected.

Our second objective was to determine the mechanism triggering defense

responses of A. philoxeroides to herbivory. For morphometric variables of apical ramets,

the effects of leaf removal and the combined application of jasmonic acid and leaf

removal were similar to the effects caused by the real attack of the specialist herbivore A.

hygrophila. These results are reasonable because A. hygrophila is a leaf-feeding herbivore

and mechanical leaf removal has been described as an accurate simulation of its herbivory

damage in A. philoxeroides (Schooler et al. 2006). Interestingly, we detected a non-local

effect of herbivory on basal ramets. In a compensatory response, basal ramets connected

to apical ramets under A. hygrophila herbivory and the combined application of jasmonic

acid and leaf removal significantly increased root biomass of basal ramets. The absence

of a similar compensatory response under the attack by the sap-feeding generalist P.

minor suggests that leaf damage is necessary to trigger this non-local response. Our

results also indicate that a combination of mechanical leaf removal and jasmonic acid

application is required to induce this compensatory response, whereas leaf removal alone

may not have an appreciable effect. Such results are in line with previous studies reporting

that the combined use of leaf removal and jasmonic acid may represent a more realistic

scenario to mimic real herbivory (van Kleunen et al. 2004).


184

Figure 2. Biomass (a, c-e), number of ramets (b) and root to shoot ratio (f) of the basal ramets of

Alternanthera philoxeroides. Treatment codes: C - control, Ag – herbivory by the specialist Agasicles

hygrophila, Pl – herbivory by the generalist Planococcus minor, JA- jasmonic acid application, LR - leaf

removal, and JA+LR - jasmonic acid application plus leaf removal. Values are mean + SE. Stars indicate

significant differences between the connection and disconnection treatment within each herbivory

treatment. See Table 1 for ANOVA results.

Chapter V

185

Figure 3. Biomass (a, c-e), number of ramets (b) and root to shoot ratio (f) of the apical ramets of



removal, and JA+LR - jasmonic acid application plus leaf removal. Values are mean + SE. Letters on the

bars indicate differences between the herbivory treatments. Stars indicate significant differences between

the connection and disconnection treatment within each herbivory treatment. See Table 1 for ANOVA

results.


186

Figure 4. Biomass (a, c-e), number of ramets (b) and root to shoot ratio (f) of the complete clonal system

of Alternanthera philoxeroides. Treatment codes: C - control, Ag – herbivory by the specialist Agasicles



bars indicate differences between the herbivory treatments. See Table 1 for ANOVA results.

Chapter V

187

Table 1. ANOVAs results for effects of block, herbivory treatments, connection (clonal integration), and

herbivory x connection (H x C) on growth and root to shoot ratio of the basal part, the apical part and the

whole clonal fragment of Alternanthera philoxeroides. F and p values are given. Values for which p < 0.05

(after Bonferroni correction) are in bold. See Figs. 2-4 for data.

Basal part Apical part Clonal fragment

d.f. F p d.f. F p d.f. F p

Total mass

Block 6 21.05 0.066 6 5.09 0.027 6 13.31 0.015

Herbivory 5 0.87 0.511 5 10.84 <0.001 5 10.12 <0.001

Connection 1 1.91 0.216 1 1.27 0.303 1 3.02 0.133

H x C 5 0.98 0.446 5 0.88 0.505 5 1.61 0.190

Residuals 64 66 63

No. of ramets

Block 6 2.64 0.125 6 5.96 0.023 6 3.56 0.097

Herbivory 5 0.51 0.769 5 3.31 0.017 5 1.80 0.143

Connection 1 <0.01 0.981 1 1.93 0.214 1 0.29 0.609

H x C 5 0.95 0.466 5 0.96 0.457 5 1.92 0.123

Residuals 64 66 63

Leaf mass

Block 6 14.48 0.025 6 4.4 0.055 6 9.72 0.098

Herbivory 5 1.06 0.403 5 10.58 <0.001 5 9.11 <0.001

Connection 1 0.95 0.367 1 2.57 0.160 1 2.63 0.156

H x C 5 1.36 0.270 5 0.60 0.700 5 1.86 0.133

Residuals 64 66 63

Stem mass

Block 6 17.27 0.013 6 4.15 0.043 6 9.27 0.016

Herbivory 5 0.78 0.569 5 9.18 <0.001 5 9.91 <0.001

Connection 1 4.12 0.089 1 1.58 0.255 1 2.71 0.151

H x C 5 0.73 0.606 5 1.02 0.424 5 1.33 0.280

Residuals 64 66 63

Root mass

Block 6 6.12 0.025 6 34.39 0.399 6 26.02 0.132

Herbivory 5 2.17 0.085 5 4.45 0.004 5 3.30 0.017

Connection 1 9.20 0.023 1 5.47 0.058 1 0.49 0.511

H x C 5 1.39 0.259 5 1.86 0.131 5 2.33 0.069

Residuals 64 66 63

Root to shoot ratio

Block 6 1.00 0.469 6 4.50 0.131 6 8.92 0.171

Herbivory 5 1.13 0.367 5 3.14 0.021 5 3.45 0.014

Connection 1 11.88 0.014 1 21.29 0.004 1 3.41 0.114

H x C 5 2.33 0.069 5 1.92 0.121 5 1.58 0.198

Residuals 64 66 63


188

Several studies have showed that the exogenous application of jasmonic acid or

its derivate provokes the same defensive responses in different plant species as those

caused by predators (Thaler et al. 1996; Baldwin 1998). Several members of the

jasmonate family have been identified as intermediates in the long-range defensive

signaling (Truman et al. 2007; Huot et al. 2014). Moreover, a number of studies have

detected the role of jasmonic acid in defensive responses, including the generation of

chemical compounds (Thaler et al. 1996; Baldwin 1998; Cipollini & Sipe 2001), the

alteration of carbon allocation (Babst et al. 2005, Henkes et al. 2008), and long-distance

defensive signaling (Truman et al. 2007; Huot et al. 2014).

Fig. 5. Concentrations of total phenolics (a, b) and condensed tannins (c, d) in basal and apical ramets of




bars indicate differences between the herbivory treatments. Stars indicate significant differences between

the connection and disconnection treatment within each herbivory treatment. See Table 2 for ANOVA

results.

Chapter V

189

Contents of phenolic compounds (including tannins) in leaves are often used as a

measure of non-specific, quantitative and carbon-based anti-herbivory defense of plants

(Asquith & Butler 1986; Forkner et al. 2004). These compounds can lead to the inhibition

of enzymes (e.g. in the digestive tract of insects) or the formation of the insoluble complex

with plant proteins, reducing their dietary value (Asquith & Butler 1986). Several studies

have reported a correlation between the tannin or total phenolic content and the negative

effect on feeding herbivores (Karowe 1989; Forkner et al. 2004). However, we did not

find significant differences in the contents of phenolics and tannins between the control

and the real or simulated herbivory treatments. A plausible explanation for this

observation is that A. philoxeroides may have used constitutive defense that are developed

permanently (Wittstock & Gershenzon 2002), regardless of the presence or the type of

herbivory. As a result, induced chemical defense was not activated during our experiment.

It appears that A. philoxeroides based its defensive strategy against herbivory on

compensatory growth responses, rather than investing resources in the synthesis of

defensive chemicals. By reducing the investment in costly chemical defensive

compounds, A. philoxeroides could spend more resources on growth, thus obtaining a

competitive advantage over other species.

Table 2. ANOVA results for effects of herbivory, connection (clonal integration) and herbivory x

connection (H x C) on the total phenolics and condensed tannins in basal and apical ramets of Alternanthera

philoxeroides. F and p values are given. Values for which p < 0.05 (after Bonferroni correction) are in bold.

See Fig. 5 for data.

Basal part Apical part

d.f. F p d.f. F p

Phenolics

Herbivory 5 1.87 0.118 5 1.47 0.217

Connection 1 0.66 0.420 1 1.51 0.225

H x C 5 1.42 0.233 5 1.25 0.300

Residuals 48 48

Tannins

Herbivory 5 0.69 0.636 5 3.74 0.006

Connection 1 1.01 0.320 1 0.04 0.836

H x C 5 3.24 0.013 5 0.80 0.556

Residuals 48 48


190

Our third objective was to test whether physiological integration can help buffer

the negative impact of herbivory. Interestingly, physiological integration significantly

reduced root to shoot ratio and thus increased proportional biomass allocation to

aboveground structures of apical ramets under real herbivory by both insects. Increasing

biomass allocation to aboveground parts could potentially increase the relative ability of

A. philoxeroides to expand laterally. Therefore, the plastic response mediated by

physiological integration could be considered a resistance strategy against herbivores, and

might contribute to the successful expansion of A. philoxeroides. However, clonal

integration was found to increase biomass allocation to shoots of defoliated apical ramets

of A. philoxeroides (You et al. 2014). There are at least two plausible explanations for

such a discrepancy. First, our response was detected under real herbivory, while You et

al (2014) only simulated herbivory by leaf removal. Second, You et al. (2014) applied the

leaf-removal treatment only at the beginning of the experiment, while our insect herbivory

was continuous throughout the experiment. These discrepancies call attention to the

importance of considering the type (real vs. simulated) and duration (continuous vs.

occasional) of the treatments when testing herbivory effects.

Physiological integration also affected biomass allocation of basal ramets, and

basal ramets significantly increased biomass allocation to roots when they were

connected to apical ramets consumed by the leaf-feeding herbivore A. hygrophila. This

result could be interpreted as a non-local compensatory response of basal ramets to meet

the demand of apical ramets under the stressful conditions created by herbivory. The

plastic response detected in basal ramets in response to local conditions experienced by

apical ramets agrees with the modular concept of phenotypic plasticity in plants, which

proposes that physiological integration can modify local responses of modules (de Kroon

et al. 2009). It is important to note that the non-local compensatory response of basal

ramets mediated by physiological integration was only detected when apical ramets were

attacked by the leaf-feeding herbivore. As discussed above, the leaf-feeding insect A.

hygrophila imposed the strongest negative effects on apical ramets and, therefore, the

compensatory response was only activated under this attack.

Previous studies have reported that physiological integration can induce systemic

resistance within a clone (Gómez & Stuefer 2006; Gómez et al. 2007, 2008).

Consequently, un-attacked ramets can also activate defense in responses to the damage

suffered by other members of the clone, reducing the potential negative impact of

Chapter V

191

herbivory on the whole clone. However, we did not detect induced systemic resistance in

basal ramets that remained connected to apical ramets suffering from real or simulated

herbivory. The alerting signal is probably transmitted by phloem following a source-sink

gradient of resources (Gómez & Stuefer 2006). In our experiment, the predominant

gradient of resources and thus the direction of resource translocation were from un-

attacked basal ramets to apical ramets suffering from herbivory or leaf removal. As there

was no predominant gradient of resources from apical to basal ramets, the alerting signal

could not be transported from apical to basal ramets so that induced systemic resistance

could not occur (Gómez & Stuefer 2006).

We conclude that physiological integration influences the defense strategies of A.

philoxeroides against herbivores. Physiological integration increased the allocation to

aboveground parts in apical ramets attacked by herbivores, resulting in a more extensive

lateral growth. Our results highlight the importance of physiological integration and

modular plasticity for the interpretation of the effect of herbivory in clonal plants. In

addition, the combination of leaf removal and jasmonic acid application plays a similar

role in triggering the compensatory response of A. philoxeroides to that of herbivory by

the real leaf-feeding insect. Differences between leaf-feeding and sap-feeding herbivores

and between real and simulated herbivory should be taken into account to disentangle the

defensive response of clonal plants.

Acknowledgments

This research was supported by the National Key Research and Development Program of

China (2016YFC12011000) and NSFC (31570413, 31500331) to F.-H. Y. and B.-C. D.

S. R. R., R. B. and R. P. acknowledge funding from the Spanish Ministry of Economy

and Competitiveness (project Ref. CGL2013-44519-R, cofinanced by the European

Regional Development Fund, ERDF, granted to S. R. R.). This is a contribution from the

Alien Species Network (Ref. ED431D 2017/20 – Xunta de Galicia, Autonomous

Government of Galicia).


192

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Xu, L., Yu, F. H., van Drunen, E., Schieving, F., Dong, M., & Anten, N. P. R. (2012) Trampling, defoliation

and physiological integration affect growth, morphological and mechanical properties of a root-suckering

clonal tree. Annals of Botany, 109, 1001–8.

You, W., Dan, Y., Dong, X., Han, C., & Liu, C. (2014). The invasive plant Alternanthera philoxeroides

benefits from clonal integration in response to defoliation. Flora 209(11), 666-673.

Yu, F. H., Dong, M., & Krusi, B. (2004) Clonal integration helps Psammochloa villosa survive sand burial

in an inland dune. New Phytologist, 162, 697-704.

Yu, F. H., Ning, W., Alpert, P., He, W. M., & Ming, D. (2009). Physiological integration in an introduced,



Yu, F. H., Wang, N., He, W. M., Chu, Y., & Dong, M. (2008) Adaptation of rhizome connections in drylands:

Increasing tolerance of clones to wind erosion. Annals of Botany, 102, 571-577.

Chapter VI

Trans-generational effects in the clonal invader


Rubén Portela1,5, Bi-Cheng Dong2, Fei-Hai Yu3,4, Rodolfo Barreiro1,

Sergio R. Roiloa1, Dalva M. Silva Matos5

1BioCost Group, Biology Department, Universidade da Coruña, A Coruña 15071, Spain.

2School of Nature Conservation, Beijing Forestry University, Beijing 100083, China.

3Institute of Wetland Ecology and Clone Ecology, Taizhou University, Taizhou 318000,

China

4Zhejiang Provincial Key Laboratory of Plant Evolutionary Ecology and Conservation,

Taizhou University, Taizhou 318000, China.

5Lab. Ecologia e Conservação, Departamento de Hidrobiologia, Universidade Federal de

São Carlos, São Carlos 13565-905, Brazil

Published as Portela, R., Dong, B. C., Yu, F. H., Barreiro, R., Roiloa, S. R., & Silva

Matos, D. M. (2019). Trans-generational effects in the clonal invader Alternanthera

philoxeroides. Journal of Plant Ecology. doi: 10.1093/jpe/rtz043

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Abstract

Recent studies have revealed heritable phenotypic plasticity through vegetative

generations. In this sense, changes in gene regulation induced by the environment, such

as DNA methylation (i.e. epigenetic changes), can result in reversible plastic responses

being transferred to the offspring generations. This trans-generational plasticity is

expected to be especially relevant in clonal plants, since reduction of sexual reproduction

can decrease the potential for adaptation through genetic variation. Many of the most

aggressive plant invaders are clonal, and clonality has been suggested as a key trait to

explain plant invasiveness. Here we aim to determine whether trans-generational effects

occur in the clonal invader Alternanthera philoxeroides, and whether such effects differ

between populations from native and non-native ranges. In a common garden experiment,

parent plants of A. philoxeroides from populations collected in Brazil (native range) and

Iberian Peninsula (non-native range) were grown in high and low soil nutrient conditions,

and offspring plants were transplanted to control conditions with high nutrients. To test

the potential role of DNA methylation on trans-generational plasticity, half of the parent

plants were treated with the demethylating agent 5-azacytidine. Trans-generational

effects were observed both in populations from the native and the non-native ranges.

Interestingly, trans-generational effects occurred on growth variables (number of ramets,

stem mass, root mass and total mass) in the population from the native range, but on

biomass partitioning in the population from the non-native range. Trans-generational

effects of the population from the native range may be explained by a ‘silver-spoon’

effect, whereas those of the population from the non-native range could be explained by

epigenetic transmission due to DNA methylation. Our study highlights the importance of

trans-generational effects on the growth of a clonal plant, which could help to understand

the mechanisms underlying its invasive success.

Keywords: 5-azacytidine; Alternanthera philoxeroides; alligator weed; clonal growth;

DNA methylation; epigenetic variation; plant invasions.


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1. Introduction

Invasive species represent a significant threat to global biodiversity, decreasing

the abundance and richness of native species (Gaertner et al. 2009; McGeoch et al. 2010).

Discovering the underlying mechanisms of biological invasions is a fast moving research

topic in modern ecology. One core question is to determine why some species have the

potential to become invasive when naturalized outside their native range (Alpert et al.

2000). A widely used approach is to study those life-history traits of the invasive species

that allow or facilitate this invasion process (Hamilton et al. 2005; Kolar & Lodge 2001).

Phenotypic plasticity is the ability of an organism to respond to different

environments by expressing its genotype according to the influence of the environment

in which it develops (Bradshaw 1965; Sultan 2000). Plasticity allows a genotype to have

a wider tolerance to environmental conditions, and therefore a higher fitness across

multiple habitats (Caño et al. 2008; Ghalambor et al. 2007; van Kleunen & Fischer 2005).

This feature is considered relevant in biological invasions, since those species that are

more plastic have greater facility to successfully adapt themselves to new environments

(Hulme 2008; Smith 2009; Zenni et al. 2014). A widely described case of phenotypic

plasticity is the changes that occur in response to scarcity of an essential resource for the

development of the plant, mainly water, light or nutrients. In such a situation, plants may

respond by increasing the allocation of biomass to the structures responsible for obtaining

the most limiting resource for growth (Chapin et al. 1987; Ryser & Eek 2000). For

instance, plasticity in shoot to root ratio allows the plant to maintain an optimal

development in conditions of scarcity (Aikio & Markkola 2002; Grossman & Rice 2012).

Interestingly, the environmental conditions experienced by the parental generation

may influence the offspring phenotype due to trans-generational effects (Galloway 2010;

Galloway & Etterson 2007; Latzel & Klimešová 2010; Sultan et al. 2009). For example,

resource levels experienced by the parental generation could regulate the phenotype

expressed by the offspring generation. Under this premise, it seems logical to anticipate

that offspring generations could gain a benefit when occupying an area with similar

conditions to those experienced by their parents (Dong et al. 2017, 2018). In this sense,

trans-generational effects could favor the successful establishment of offspring

generations, and therefore be considered adaptive (Galloway & Etterson 2007; Latzel et

al. 2014). This trans-generational plasticity is expected to be especially relevant in clonal

plants, for which reduction of sexual reproduction can decrease the potential for

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adaptation through genetic variation, and it could explain the success of clonal species in

many communities (Latzel & Klimešová 2010). Many studies have been conducted to

explore trans-generational plasticity in plants (Dong et al. 2017, 2018; González et al.

2016, 2017; Latzel et al. 2014; Latzel et al. 2010; Sultan et al. 2009). To date, however,

less effort has been dedicated to determining the importance of trans-generational

plasticity in plant invasiveness (Caño et al. 2016; Richards et al. 2012).

A number of studies point out that the influence of environmental conditions on

phenotypes could be mediated by heritable epigenetic variation (Hallgrímsson & Hall

2011; Verhoeven & Preite 2014), through which trans-generational plasticity can occur

(Boyko et al. 2010). Heritable epigenetic variation does not involve alterations of the

DNA sequence, but involves regulatory mechanisms of gene expression, including the

repression or silencing of particular genes (Wolffe & Matzke 1999). This could provide

an alternative way for a species to adapt to a new environment, i.e. as opposed to selection

for a different genotype (Bossdorf et al. 2008; Pérez et al. 2006). Epigenetic changes also

have the potential advantage of being easily reversible (Bender 2004). In organisms with

asexual reproduction, epigenetic changes allow an increase of the phenotypic variability

of the populations based on the environment in which they develop, without being

necessary an alteration in the genome sequence (Gao et al. 2010, Wang et al. 2019).

Changes in gene expression regulated by epigenetic mechanisms are also called trans-

generational memory mechanisms, due to the effects they have on the offspring (González

et al. 2016; Heard & Martienssen 2014). A widely described epigenetic regulatory

mechanism is DNA methylation (Bender 2004). Some epigenetic changes are heritable,

so they can be transmitted to offspring (Heard & Martienssen 2014; Martienssen & Colot

2001) and used as a strategy to cope with adverse environmental conditions (Asensi-

Fabado et al. 2016; Dong et al. 2017, 2018; Probst & Mittelsten 2015; Secco et al. 2017).

Trans-generational plasticity could be considered a favorable trait, allowing

invasive plants, especially those that are clonal, to overcome low genetic variation and to

adapt successfully to the new environment in which they develop. In this study we

assessed trans-generational plasticity in the clonal invader Alternanthera philoxeroides to

soil nutrients. Parental plants were grown under high and low nutrient conditions, whereas

offspring plants were grown in a common environment with high nutrients. Also, to test

the potential role of DNA methylation on trans-generational plasticity, half of the parent

plants were treated with the demethylating agent, 5-azacytidine. We hypothesized that


202

environmental conditions experienced by parental generations would affect the growth

and biomass allocation exhibited by offspring generation (i.e. trans-generational effect).

Specifically, we predict (i) that offspring ramets whose parents grew in high nutrients

conditions will show greater growth than ramets whose parents grew in low nutrients

conditions; and (ii) that offspring ramets whose parents grew in low nutrients conditions

will increase the proportional allocation of biomass to roots in comparison with offspring

of parents in high nutrient conditions.


2.1.Study species

Alternanthera philoxeroides (Mart.) Griseb. (Amaranthaceae), commonly known

as alligator weed, is a creeping perennial herb native to the Parana River region in South

America (Julien et al. 1995). It is an amphibious plant that can grow both on land and as

floating mats in water courses (Lu & Ding 2010). Sexual reproduction is unusual in its

introduced ranges (Buckingham 1996; Lu et al. 2013), but clonal reproduction from stem

buds is common in both aquatic and terrestrial habitats in the invaded areas (Dong et al.

2010, 2019; Julien et al. 1995; Sainty et al. 1998; Yu et al. 2009). In the introduced range

A. philoxeroides displaces native species, leading to monospecific communities, with the

consequent loss of biodiversity (Julien & Bourne 1988; Ma & Wang 2005). It can

completely cover watercourses and block irrigation systems in riparian crop fields,

causing floods (Sainty et al. 1998; Stewart 1996; Wang et al. 2008). The geographical

distribution of the species out of their native range includes the south of North America

(Buckingham 1996), Italy, France (Anderson et al. 2016), Australia, New Zealand (Julien

et al. 1995), Indonesia, Thailand and China (Sainty et al. 1998). A. philoxeroides is

considered one of the worst invasive weeds worldwide because of its rapid growth,

resistance to control methods, and ecological and economic impacts. A naturalized

population of A. philoxeroides has been recently described at the Iberian Peninsula

(Fisterra, NW Spain) (Romero & Amigo 2015) (see Fig. 1).

2.2.Plant material

Plant material used in the experiment was collected from three different locations:

one at the introduced area (Iberian Peninsula, in Fisterra: 42°56'10"N, 9°16'13" W), and

two at the native range (Brazil, in São Carlos: 21°59'1'' S, 47°52'43'' W, and Piracicaba:

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22°47'14'' S, 47°38'46'' W). From each location a single genotype was collected and

propagated. For the three locations plant material was collected in similar terrestrial

habitats. By collecting a single genotype we prevent from genetic differences within

populations to interfere with the epigenetic effects on character transmission. Thus, in the

absence of genetic variation, differences in gene expression should be solely caused by

an epigenetic effect (Bossdorf et al. 2008). In this sense, clonal plants can be considered

a suitable model for epigenetic studies. Plants were collected between June and August

2017 and vegetatively propagated in a greenhouse of the Federal University of São Carlos

(São Carlos, Brazil).


In October 2017, clonal fragments comprising the first, second and third ramets

from the apices were selected from the propagated plant material of the native (São Carlos

and Piracicaba) and non-native (Fisterra) populations. We selected 12 clonal fragments

from each population, totalizing a total of 36 clonal fragments (12 fragments x 3

populations). These ramets, which represent the parental generation, were randomly

assigned to factorial nutrient and de-methylation treatments, ensuring that all the

populations were equally represented in each treatment combination (see Fig. 1). For the

nutrient treatment half of the ramets grew in potting compost (high nutrients), and the

other half in washed sand (low nutrients). Potting compost used for the high nutrient

treatment contained all main nutrients and trace elements, and can be considered as fertile

soil providing optimal growth conditions. For the de-methylation treatment, half of the

ramets were treated with a de-methylating agent, 5-azacytidine (Sigma-Aldrich Brasil

Ltda. São Paulo, Brazil), which is a substance analogous to cytosine that inhibits DNA

methylation in eukaryotes (Čihák 1974). 5-azacytidine has been previously used to test

the role of epigenetic variation in phenotypic plasticity and trans-generational adaptation

to stress (Bossdorf et al. 2010; González et al. 2016). The de-methylation treatment was

applied in mid-November 2017, with six applications carried out over two weeks. Plants

were sprayed with a 50 mol/L solution of 5-azacytidine. The solution was applied early

in the morning, when stomata are expected to be open, to ensure absorption by leaves.

The 5-azacytidine solution was prepared with a mixture of acetic acid and distilled water

at a volume ratio of 1:1. Plants that did not receive the de-methylating treatment were

sprayed with the same solution without the de-methylating agent, in order to homogenize

any side effects of acetic acid on the plants. Each nutrient and de-methylation treatment


204

was replicated three times (n = 3). Ramets of the parental generation grew for two months

(October and November 2017).

Fig. 1. Scheme of the experimental design, including the treatment applied on the parental ramets and the

schedule of the experiment.

Afterwards, in December 2017, from each ramet of the parental generation three

axillary stems were obtained to create the offspring generation (n = 9), totalling 108

offspring fragments, 36 from each original population. The offspring generation grew

under high nutrients and not subjected to de-methylation (see Fig. 1). The transplant of

the offspring generation was performed in the first week of December 2017, and the plants

were maintained for 5 months, until harvest in April 2018. The initial size of the offspring

generation was determined as the number of leaves for each clonal fragment. Preliminary

analysis showed that the initial size of the offspring generation from the studied

populations differed significantly (ANOVA: F2,104 = 12.2, P < 0.001), with plants from

São Carlos having a higher number of leaves (7.1 ± 2.9, mean ± SE) than plants from

Piracicaba (5.4 ± 2.1) and Fisterra (4.5 ± 2.0). Parental and offspring ramets were placed

in 2.8 L plastic pots, providing enough space to avoid root confining during the

experiment. The experiment was carried out in the same greenhouse at the Federal

University of São Carlos (São Carlos, Brazil) where plant pre-cultivation was carried out,

under a natural day / night light cycle and ambient temperature. Plants were watered

regularly to avoid water stress.

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2.4.Measurements

At the end of the experiment each clonal fragment of the offspring generation was

separated into leaves, stems and roots, dried at 60 ºC for 48h and weighed. For each clonal

fragment we calculated the total mass (leaf mass + stem mass + root mass) and the

proportional biomass allocated to roots (root mass ratio: RMR = root mass / total mass).

Also, the number of individual ramets in each clonal fragment of the offspring generation

was recorded.

Statistical analysis

Analyses were run for the offspring generation. Data were analysed by a three-

way ANCOVA with nutrients (high and low nutrients), de-methylation (de-methylated or

not) and population (Fisterra, São Carlos, Piracicaba) as main factors. The initial plant

size of the offspring generation, estimated as the number of leaves per clonal fragment,

was used as a covariate to take into account variation among populations. Normality and

homoscedasticity were checked using the Kolmogorov–Smirnov and Levene tests. All

variables were square-root transformed to fulfil requirements for ANOVA. When results

were significant, we applied Tukey tests to detect differences between populations, and

between the treatments within each population. Mortality reduced the number of

replicates used in the different analyses, as indicated by the error degrees of freedom. All

analyses were conducted with the software IBM SPSS Statistics, version 23 (IBM Corp.

Armonk, NY).

3. Results

Nutrient conditions experienced by the parental ramets significantly affected

number of ramets, stem mass, root mass and total mass of the offspring generation (Table

1). Thus, when parental ramets grew in high nutrients, their offspring showed

significantly greater number of ramets, stem mass, root mass and total mass than offspring

whose parents grew under low nutrient conditions (Fig. 2a, c-e). De-methylation

treatment significantly reduced leaf mass, independent of the population and of the

nutrient conditions experienced by the parental generation (Table 1, Fig. 2b). The

interaction between nutrients and de-methylation significantly affected root mass and

RMR (Table 1). De-methylation increased root mass and RMR in the high nutrient

treatment, but reduced them in the low nutrient treatment (Fig. 2d,f).


206

Population origin significantly affected all the variables except root mass (Table

1). Thus, plants from Fisterra population (non-native range) showed smaller number of

ramets, leaf mass, stem mass, and total mass than plants from São Carlos and Piracicaba

populations (native range) (Fig. 2a–c,e). On the contrary, RMR was significantly greater

in Fisterra population (non-native range) than in São Carlos and Piracicaba populations

(native range) (Table 1, Fig. 2f). Interestingly, the effect of nutrient status experienced by

the parental generation was dependent on the population origin. The interaction between

nutrients and population significantly affected the number of ramets, stem mass, total

mass and RMR (Table1). When parent ramets grew in low nutrient conditions, we

observed a significant reduction in number of ramets, stem mass and total mass of

offspring ramets from São Carlos and Piracicaba populations (native range), but not in

those offspring from the Fisterra population (non-native range) (Fig. 2a,c,e). On the

contrary, the effect of nutrients on RMR was detected in the Fisterra population (non-

native range) but not in the São Carlos and Piracicaba populations (native range). Thus,

when parent ramets grew in high nutrient conditions, their offspring significantly reduced

RMR, and this effect was only detected in the Fisterra population (non-native range) (Fig.

2f). Remarkably, the Tukey test detected that low nutrients in the parental environment

significantly increased RMR of offspring from the Fisterra population (non-native range),

but this increase was only detected in plants not de-methylated (Fig. 2f).

The interaction between de-methylation and population significantly affected leaf

mass and total mass (Table 1). De-methylated plants showed a reduction in leaf mass, and

this effect was only detected in the São Carlos and Piracicaba populations (native range)

(Fig. 2b). On the other hand, plants from the São Carlos population (native range) reduced

total biomass due to de-methylation, but this trend was not found in either the Piracicaba

population (native range) or the Fisterra population (non-native range) (Fig. 2e). The

interaction between nutrients, de-methylation and population did not significantly affect

any of the variables (Table 1).

4. Discussion

Many of the worst invasive plants around the world are clonal (Yu et al. 2009).

This could a priori represent a paradox, as lack of genetic diversity associated with clonal

reproduction would reduce the capacity of clonal plants for adaptation to the new

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environment. This situation would be even more accentuated due to genetic drift,

presumably associated to the process of invasion, which would also reduce genetic

diversity of invasive populations (Keller & Taylor 2008). In spite of this apparent

limitation, clonality has been considered as a trait that could explain plant invasiveness

(Pyšek 1997). In the past, many studies have been conducted to explicitly test whether

traits associated with clonal growth favor the expansion of invasive plants (e.g. Roiloa et

al. 2016; Song et al. 2013; Wang et al. 2017; Chen et al. 2019). Furthermore, it has been

suggested that epigenetic regulations in gene expression would allow the establishment

of invaders in the short term (Pérez et al. 2006). However, the role of trans-generational

effects in the invasiveness of clonal species has been generally overlooked (but see Dong

et al. 2017, 2018). Our results suggest that trans-generational effects occurred in the

clonal invader A. philoxeroides.

As predicted, environmental conditions experienced by parental generation

affected the growth of the offspring generation, with ramets whose parents grew under

high nutrient conditions showing greater number of ramets, stem and total mass than

ramets whose parents grew in low nutrient conditions. This trans-generational effect on

number of ramets, stem mass and total mass was only detected in populations from Brazil

(native range), but not in the population from the Iberian Peninsula (non-native range). In

addition, this trans-generational effect was not mediated by DNA methylation, indicating

that other mechanisms were involved in this induced effect. One plausible mechanism to

explain this result is that parental plants growing under high nutrient conditions would

reduce the allocation of energy to produce roots, as stated by the optimal partitioning

theory (Gleeson & Tilman 1992; Hilbert 1990). Therefore, plants would invest more

resources in the production of above-ground structures, including the production of more

or greater offspring in the case of clonal plants. In addition, the acropetal transport of

resources from parent ramets established in favorable conditions will provide the

offspring generation with more resources than the resources received from parents under

less favorable conditions. Thus, offspring ramets that were subsided by parents growing

in high nutrient conditions are expected to gain greater early advantage than those ramets

supported by parents in low nutrient conditions. This initial advantage could be

maintained even after disconnection, contributing to the greater total biomass shown by

offspring of parent plants grown in high nutrient conditions. This transport of resources

between connected units of clonal systems is mediated by physiological integration, and


208

many previous studies have reported benefits for offspring ramets, especially when

subsidized by parents growing in favorable conditions (e.g. Hartnett & Bazzaz 1983;

Roiloa & Retuerto 2006; Saitoh et al. 2002; Slade & Hutchings 1987; Stuefer 1994). Our

results fit with a ‘silver-spoon’ effect, where individuals developing in favorable

conditions will gain performance advantages in the long term (Grafen 1988). Thus, it is

expected that the environment experienced by the parental generation can contribute

substantially to the phenotype of the offspring generation, and a parent plant growing in

good conditions will provide more resources to their descendants (Roach & Wulff 1987).

However, as mentioned above, this ‘silver-spoon’ effect was not detected in the

population from the Iberian Peninsula (non-native range). In addition, the population

from the non-native range, independent of the nutrient conditions experienced by their

parental generation, performed significantly worse than populations from the native

populations. A plausible explanation for these results is that much greater biomass was

allocated to roots (RMR) in the non-native population in comparison with the native

populations. By increasing RMR non-native plants probably reduced their net carbon

gain, as the ratio between photosynthetic and non-photosynthetic structures was reduced.

This would increase proportional respiration in relation with photosynthesis, and

consequently reduce plant growth. The significant increase in RMR detected in non-

native populations could be explained as an effect of physiological integration. Offspring

ramets were disconnected from the parental plants, and consequently potential resource

transport stopped. Under this situation, offspring ramets from the non-native population

responded by increasing the proportional production of roots to compensate for the lack

of subside received from the parental ramets. Previous experiments have demonstrated

that disconnection (physiological integration impeded) conducted to a significant increase

of the biomass allocated to produce roots in populations of the invasive Carpobortus

edulis at their non-native range (Roiloa et al. 2013, 2019). However, to truly confirm or

discard this conjecture, our study should have included a connection/disconnection

treatment to determine whether there is an effect of physiological integration in biomass

allocation patterns, and whether this effect differs between native and non-native

populations.

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Fig. 2. Number of ramets (a), leaf mass (b), stem mass (c), root mass (d), total mass (e) and root mass ratio

(RMR; f) of the offspring generation derived from the three populations of Alternanthera philoxeroides:

Fisterra (Iberian Peninsula, non-native range), São Carlos (Brazil, native range), and Piracicaba (Brazil,

native range). Values are mean + SE. See Table 1 for ANCOVA results. Letters indicate differences

between (upper case) and within (lower case) populations according to Tukey test (P < 0.05).

Table 1. Results of three-way ANCOVAs for the effects of nutrient level, de-methylation and population of origin on number of ramets, leaf mass, stem mass, root mass, total

mass and root mass ratio of the offspring generation. Initial size, estimated by the number of leaves, was used as a covariate in the model. Significant effects (P < 0.05) are

shown in bold. See Fig. 2 for data.

Nº of ramets Leaf mass Stem mass Root mass Total mass RMR

Effect df F P F P F P F P F P F P

Initial size 1 1.05 0.308 1.37 0.245 0.17 0.684 0.29 0.595 0.20 0.654 3.95 0.050

Nutrients (N) 1 17.74 <0.001 1.57 0.213 16.57 <0.001 6.19 0.015 11.33 0.001 0.75 0.390

De-methylation (D) 1 2.94 0.095 0.60 0.015 0.02 0.769 0.13 0.358 0.11 0.588 <0.01 0.417

Population (P) 2 81.31 <0.001 89.72 <0.001 101.4 <0.001 1.62 0.205 70.37 <0.001 154.38 <0.001

N × D 1 0.92 0.34 0.80 0.374 0.66 0.421 11.61 <0.001 3.39 0.069 19.83 <0.001

N × P 2 10.00 <0.001 2.64 0.078 6.08 0.003 2.83 0.065 4.54 0.013 3.78 0.027

D × P 2 1.45 0.24 6.97 0.002 2.70 0.073 1.73 0.184 3.91 0.024 1.11 0.335

N × D × P 2 0.71 0.493 1.61 0.206 1.05 0.354 0.37 0.691 0.72 0.491 1.87 0.161

Error 81

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Our results also showed a trans-generational effect in biomass partitioning,

estimated as proportional biomass allocated to roots (RMR). Interestingly, this trans-

generational effect on RMR was only detected in the population from the Iberian

Peninsula (non-native range) but not in those from Brazil (native range). Furthermore,

this trans-generational effect was probably mediated by DNA methylation, as the changes

in biomass partitioning induced by the parental effect were not present in plants whose

parental plant was subjected to de-methylation. Several studies have reported the presence

of epigenetic mechanism triggered by DNA methylation to explain the occurrence of

trans-generational effects in plants (Bender 2004). Previous works with A. philoxeroides

in China showed a positive correlation between phenotypic variability of plants and

methylations in their DNA (Gao et al. 2010), although the genome of the populations was

virtually identical due to the absence of sexual reproduction in the areas that the plant

invades (Wang et al. 2005; Ye et al. 2003). However, although our results suggests the

presence of an epigenetic control over the trans-generational effect detected in biomass

partitioning, examination of the methylation patterns in the parental and offspring

generations should be conducted to elucidate the presence of a epigenetic effect

modulated by DNA methylation.

Trans-generational effects on a plastic trait, such as biomass partitioning, could

have important adaptive implications. Plasticity in biomass allocation allows plants to

acquire resources efficiently, boosting plants to successfully colonize new environments

(Mommer et al. 2011; Valladares et al. 2007). A common plastic response developed by

plants is the increase of the biomass assigned to produce the structures responsible for

acquiring the limiting resources, as stated by the optimal partitioning theory (Bloom et al.

1985; Hilbert 1990; Thornley 1972). Our results showed that offspring of parent plants

grown in low nutrient conditions maintained high values of RMR (high proportion of

biomass allocated to roots) even when offspring ramets grew under high nutrient

conditions, denoting the presence of a trans-generational effect on biomass partitioning.

Trans-generational plasticity could be considered adaptive when offspring generation

gains a benefit of being early informed about their future local environment (Engqvist &

Reinhold 2016). That is, when the parental environment resembles the conditions

experienced by their offspring, trans-generational plasticity could represent an adaptive

advantage, since offspring can anticipate their plastic response. However, when parental

and offspring environments differ, benefits from trans-generational plasticity are not


212

expected, and may even result in a penalty for offspring. Phenotypic plasticity has been

suggested as an important trait contributing to plant invasiveness (Parker et al. 2003;

Richards et al. 2006), and it is plausible to anticipate that phenotypic plasticity can be

positively selected during the process of invasion, favouring the spread of exotic species

(Lande 2015). However, the importance of adaptive trans-generational plasticity in plant

invasions has received less attention. In comparison with natural selection, trans-

generational plasticity would lead to a faster adaptation to local conditions, which would

allow the rapid spread into the invaded environment. This could be particularly important

for clonal invaders, in which lack of genetic diversity could reduce the action of natural

selection, and the presence of trans-generational effects may play a key role in explaining

the invasiveness of some clonal species. Although it seems logical to predict that trans-

generational effects can potentially contribute to the invasiveness of some clonal species,

this correlation does not necessarily always occur. Exploring in advance which traits are

favouring the invasiveness of A. philoxeroides is mandatory to elucidate later whether

trans-generational effects are operating over this trait, and consequently contribute to

expansion success.

Our study demonstrates the existence of trans-generational effects on the invasive

clonal plant A. philoxeroides. Trans-generational effects were observed both in

populations from the native and the non-native ranges. Interestingly, trans-generational

effects were detected on growth variables (number of ramets, stem mass, root mass and

total mass) in the populations from the native range, and on biomass partitioning in the

population from the non-native range. In addition, trans-generational effects in the

population from the native range seem to be explained by a ‘silver-spoon’ effect.

However, the effects observed in the population from the non-native range could be

explained by epigenetic transmission due to DNA methylation. Our results demonstrate

that trans-generational effects occurred in the clonal invasive A. philoxeroides, and

suggest that the mechanisms underlying these effects could differ between native and

non-native populations. Future studies including better representation of populations from

the native and the non-native range, as well as more environments for the parental and

offspring generations, must be performed to obtain a more realistic picture of the

importance of trans-generational effects in the invasiveness of A. philoxeroides.

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213

Acknowledgments

We thank Driélli De Carvalho Vergne, Mariane Patrezi Zanatta and Raquel Stucchi

Boschi for greenhouse assistance, Lilian A. Arantes de Mattos, Marcus A. Duarte and

Larissa M. da Silva Pinto for laboratory assistance. During the experiment R. P. was

supported by a mobility grant from the University of A Coruña (Inditex-UDC 2017

program). This is a contribution from the Alien Species Network (Ref. ED431D 2017/20

– Xunta de Galicia, Autonomous Government of Galicia). D. M. S. M. thanks the

Brazilian Conselho Nacional de Desenvolvimento Científico e Tecnológico/CNPq

(307839/2014-1) for her Research Fellowship.


214

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Chapter VII

A dynamic model-based framework to test the effectiveness of

biocontrol targeting a new plant invader - the case of

Alternanthera philoxeroides in the Iberian Peninsula

Rubén Portela1,2, Joana R. Vicente2,3, Sergio R. Roiloa1, João A. Cabral3

1BioCost Group, Department of Biology, Faculty of Science, Universidade da Coruña, A


2InBIO - Rede de Investigação em Biodiversidade e Biologia Evolutiva/CIBIO - Centro

de Investigação em Biodiversidade e Recursos Genéticos, Universidade do Porto, 4485-

601 Vairão, Portugal.

3Laboratory of Applied Ecology, Centre for the Research and Technology of Agro-

Environment and Biological Sciences, University of Trás-os-Montes and Alto Douro,

Vila Real, Portugal.

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Abstract

Biological invasions are one of the major threats to biodiversity at the global scale,

causing numerous environmental impacts and having high direct and indirect costs

associated with their management, control and eradication. In this work, we modelled the

use of the specialist insect Agasicles hygrophila for the biocontrol of the invasive plant

species Alternanthera philoxeroides in Fisterra (Spain), where a single population has

been recently described. To assess the effectiveness of A. hygrophila as a biocontrol agent

in the region, a population dynamic model was developed in order to include the life-

cycle of both species, as well as the interaction among them. The results of the simulations

indicate that the control of this new invasive plant is possible, as long as several releases

of the insect are made along time. The proposed model can support the control or even

the eradication of the population of A. philoxeroides with a minimal impact on the

environment, whereas optimizing the associated costs. Additionally, the proposed

framework also represents a versatile dynamic tool, adjustable to different local

management specificities (objectives and parameters) and capable of responding under

different contexts. Hence, this approach can be used to guide monitoring efforts of new

invasive species, to improve the applicability of early management measures as

biocontrol, and to support decision-making by testing several alternative management

scenarios.

Keywords: biological invasions; Alternanthera philoxeroides; alligator weed; Agasicles

hygrophila; biocontrol; dynamic modelling; risk management.


222

1. Introduction

Human activities, whether voluntarily or not, allow certain species to overcome

the geographical barriers that limit their native distribution (Mack et al. 2000; Westphal

et al. 2008). Some of these species, called exotic or non-native, become naturalized and

can adapt to the new environments, produce self-sustaining populations and become

invasive, causing serious environmental impacts and high economic costs (Levine et al.

2003; Richardson & Pyšek 2006). Globally, the economic costs related to biological

invasions are extremely high, either due to the direct impact of the species in different

economic sectors related to several ecosystem services and benefits, such as forestry,

agriculture or fishing (Huber et al. 2002; Paini et al. 2016; Pejchar & Mooney 2009;

Villamagna & Murphy 2010) or indirect costs related to control and eradication

procedures (Pimentel et al. 2005). In this scope, the European legislation concerning

invasive species (Regulation 1143/2014) highlights the importance of prevention and

early response to biological invasions.

Methods for the control of invasive plant species can be classified in three main

categories: physical, chemical, or biological (Deng et al. 2009; Lavergne & Molofsky

2006). Each category has associated advantages and disadvantages, and some of the

methods are often used together to optimize the cost-effectiveness of the species control

(Lavergne & Molofsky 2006; Stern et al. 1959). Physical methods consist on the removal

of aerial plant’s parts and roots, usually requiring high levels of management effort, and

the total elimination of the plants is often not possible (e.g. if the species has the ability

to sprout from rhizomes; Deng et al. 2009; Seiger & Merchant 1997). Chemical control

methods are grounded on the use of herbicides, therefore some compounds can negatively

affect the non-target biodiversity bellow and above ground, being their use carefully

limited to avoid ecological contaminations (Kudsk & Streibig 2003; Madsen 2000; Sainty

et al. 1998). Both physical and chemical control methods present high implementation

costs, including the cost of labor, machinery operation and cost of herbicides (Sainty et

al. 1998; Sharma et al. 2005).

The third group of control methods, the so-called biological control, is grounded

on the use of living organisms to control invasive species (DeBach & Rosen 1991). This

category of control was developed based on the enemy release hypothesis, stating that

outside their native range most of the exotic plant species are free from specialist insect

predators and pathogens, therefore spending less resources in defensive mechanisms and

Chapter VII

223

more resources to grow and reproduce (Keane & Crawley 2002). According to this theory,

the use of the natural predators is very effective, since the defensive mechanisms of the

invasive plant against its specialist predator are weakened in the invaded environments

(Joshi & Vrieling 2005; Müller-Schärer et al. 2004). The main disadvantage of biological

control is that some biotic or abiotic factors can hinder the survival of the biocontrol

species in the environment where the invasive species is located (Buckingham 1996;

Julien et al. 1995). Biological control methods do not always eradicate the invasive

species, but allow populations to be contained at restricted density levels, of which the

species no longer poses an economical or environmental risk (DeBach & Rosen 1991;

Stern et al. 1959).

The alligator weed, Alternanthera philoxeroides (Mart.) Griseb, is an amphibious

perennial herb native from the Parana basin in South America (Julien et al. 1995). This

plant species has two different ecotypes: one exhibiting narrow stems when growing on

dry terrain, and another one showing hollow stems when developing in aquatic

environments, allowing individuals to float while remain anchored to the substrate

through the roots or floating in free mats (Zuo et al. 2012). It has been reported that the

underground biomass of the plant can be up to 10 times higher than its aerial biomass

(Schooler et al. 2008). A. philoxeroides is considered an invasive plant species in several

countries worldwide, including the US, Australia, New Zealand, China, India, Italy and

France (Anderson et al. 2016). A naturalized population has recently been described in

the northwest of the Iberian Peninsula (Romero & Amigo 2015), in Spain (Fig. 1a). A.

philoxeroides can cause numerous economic and environmental impacts by successfully

outcompeting with native plants, forming a dense monospecific mat on the surface of the

water (Pan et al. 2013). The magnitude of its expansion and possible environmental

impacts in the European Union are considered high according to EPPO (European and

Mediterranean Plant Protection Organization; Anderson et al., 2015). The Spanish

Catalogue of Exotic Invasive Species (CEEEI 2013) describes A. philoxeroides as one of

the worst aquatic pests worldwide. No naturalized populations of this species had been

described in the Iberian Peninsula when the Catalogue was updated, back in 2013, and

therefore there are no control or management plans for the species. However, other

authors have described the potential risk of this plant in the case it expands to the Iberian

Peninsula (Andreu & Vilá 2010).


224

The first example of weed biocontrol happened in India in the 18th century, with

the release of the cochineal insect Dactylopius ceylonicus (Green) (Hemiptera:

Dactylopiidae) in plantations of the alien cactus Opuntia monacantha (Wildenow)

Haworth (Cactaceae) (Zimmermann Moran et al. 2009). It was a serendipity, since the

intention was to obtain cochineal-dye (i.e. carminic acid) from the insect, but a biocontrol

success nonethless, exported to other countries thereafter. The first attempt to select a

biocontrol agent occurred in the early 20th century, for the eradication of the invasive

plant Lantana camara L. (Verbenaceae) in Hawaii (van Wilgen et al. 2013). This

represented a milestone in the biocontrol practices, since it emphasized the importance of

looking for host-specific predators that would not affect other species or interrupt already

established biocontrol programs. A notable exception to the use of specialized predators

is the moth Cactoblastis cactorum (Berg) (Pyralidae), introduced in several countries to

confront various exotic cacti (Habeck et al. 1998).

The insect Agasicles hygrophila Selman & Vogt (Coleoptera: Chrysomelidae) is

a specialist predator of A. philoxeroides in the native habitat (Vogt et al. 1979). A.

hygrophila has been extensively used as biocontrol agent of A. philoxeroides in several

countries, with successful control results in some particular conditions (Buckingham

1996; Lin et al. 1984; Ma et al. 2003b; Sainty et al. 1998). Two factors limiting the

distribution, and, therefore, the effectiveness of the biocontrol, are the average

temperatures and the ecotype of the plant (Julien et al. 1995). The plant has a greater cold

tolerance than the biocontrol agent. In fact, although frosts can destroy its aerial parts, the

plant has the ability to re-sprout from its roots after winter, however the low temperatures

will kill the biocontrol adults and eggs (Stewart 1996; van Oosterhout 2007). Moreover,

the terrestrial ecotype of A. philoxeroides lacks the hollow stems necessary for A.

hygrophila to complete its life cycle (Ma et al. 2003a; Pan et al. 2011), so the insect has

been reported to cause low damage levels to the terrestrial ecotype of the plant.

Ecological models (e.g. species distribution models, agent based models, dynamic

models) have a long history of applications in ecology and management (Elith &

Leathwick 2009; Kearney & Porter 2009; Vicente et al. 2019). However, only dynamic

models are useful tools to support decision-making regarding environmental management

and species conservation. They have been extensively used to support the design and

evaluation of biological control strategies, as they allow to incorporate the dynamic

processes behind both the invasive and the biocontrol species (Godfray & Waage 1991;

Chapter VII

225

González et al. 2017; Higgins et al. 1997; Liu et al. 2004; Withers et al. 2004). The

selection of key system components represents a fundamental step in the creation of

ecological models. Temperature is recognized as a crucial factor for the development of

the life cycle of insects (the most widely used biocontrol agents), determining not only

the survival of the individuals in a given environment, but also their reproductive success

and, thus, the entire population viability (Garay et al. 2015; Stewart 1996; Stewart et al.

1999). Therefore, through the development of dynamic models including temperature as

a key system component it is possible to accurately simulate the impact of the control

agent on the invasive species, as well as the most effective time of the year for the

application of the biological control and/or the number of necessary introduction attempts

to reduce the invasive population to low, desirable density levels (Garay et al. 2015; Shea

& Possingham 2000).

In this work we aim to (i) develop a system-dynamic model by integrating the life

cycles of the invasive plant Alternanthera philoxeroides and its native predator Agasicles

hygrophila, as well as their interactions; (ii) simulate the prey-predator relationships in

order to determine the effectiveness of using biological control to limit the expansion of

A. philoxeroides in the study-region; and (iii) optimize a cost-effective biocontrol of A.

philoxeroides by testing alternative management scenarios. The proposed framework

represents a contribution to demonstrate the applicability and replicability of this

approach for the improvement of biocontrol decision-making, especially in the case of

project-based risk assessments, but also as part of wider, strategic biodiversity

conservation programmes.


2.1.Study area

The study area is located in the northwest of the Iberian Peninsula, near Cape

Fisterra (Fig. 1). The location of individuals of A. philoxeroides (42° 56' 10'' N - 9° 16'

13'' W, 65 masl) was firstly described by Romero and Amigo (2015). The study area is

characterized by having a warm-summer Mediterranean climate (Cunha et al. 2011).

Winters are cold and humid, with a maximum temperature of 15.7±2.5⁰C and a minimum

temperature of 3.8±2.1⁰C in February, while summers are hot and dry, with a maximum

temperature of 28.6±3.3⁰C and a minimum temperature of 13.1±0.9⁰C in August.


226

Climatological data was obtained from a weather station of the Spanish Meteorological

Agency (AEMET), located at coordinates 42° 55' 29'' N - 9° 17' 29'' W. The distance

between the population and the weather station is less than one kilometer. The

climatological data used in this work belong to two intervals: January 1961 - November

1969 and January 1994 - April 2018. The total precipitation throughout the year ranges

from 800 to 1400 mm. The surrounding land is mainly used for forestry and agriculture,

with maize and cauliflower crops on plots adjacent to the population of A. philoxeroides.

At a short distance there is a commercial nursery of exotic plants, abandoned for years,

which probably was the origin of the introduction of A. philoxeroides in the region

(Romero & Amigo 2015).

Figure 1. Location of the study area where the population of A. philoxeroides was described (A) (white

dashed lines), in Fisterra, Galicia (Spain), considering the context of the Iberian Peninsula (B).

2.2.The modelling framework

The proposed modelling framework (Fig. 2) combines the vegetative growth of

the invasive plant A. philoxeroides and the population dynamics of its biocontrol agent,

the beetle A. hygrophila. The life cycles of both species are interactively simulated in

order to assess the cumulative efficacy of the biocontrol, estimated in terms of plant cover

and biocontrol viability over time. The simulation period was 10 years with the day as

unit of time, considered appropriate to capture the biocontrol population dynamics and

Chapter VII

227

the plant growth seasonal patterns, both affected by daily temperatures. For the model

development the software STELLA 9.0.3 was used.

The model allows to simulate the growth of the plant both in the absence or

presence of the biocontrol. For demonstrative purposes, on the basis of available scientific

data and field work evidences the following assumptions needed to be considered: i) the

population of A. philoxeroides in Fisterra has no predators in the study area (no significant

damage has been observed in the leaves), ii) the only way of reproduction of the plant is

clonal growth (confirmed by field observations during the present study), iii) A.

hygrophila is unable to obtain food from sources other than A. philoxeroides, iv) the

biocontrol requires the leaves of the plant to lay its eggs, and v) the biocontrol will not

have significant predation once released in the area.

2.2.1. Plant sub-model

The main factors considered in the sub-model of A. philoxeroides were: the

coverage of the plant, the coverage of defoliated plants (in the case of plants having

suffered damages by biocontrol herbivory) and the extension of the suitable area where

the roots of the plant could remain after the destruction of its aerial parts (either by

herbivory on stems or due to low temperatures; Fig. 2). The coverage of the plant was

calculated considering the area where the plant currently occurs and the cover percentage

of the species in the study area. In the study area the plant species occupies three strips of

47x2m, 43x2m and 8x1m, respectively, totaling 188 m2. The initial coverage of the

species in the study area was 20%, as estimated in April of 2018. Clonal growth has been

reported as the only form of reproduction of the species outside its native area (Julien

1995). The model allows the plant to clonally expand until occupying the total available

suitable area, considering that the growth rate is constrained at lower temperatures. In

case of being damaged by the biocontrol, the model parameters allows defoliated plants

to regenerate their leaves. Even if the aerial structures of the plant (stems and leaves) are

completely destroyed, the species retains the capacity to resprout from the roots, and this

process is also included in the modelled processes. (Stewart 1996; van Oosterhout 2007).

2.2.2. Biocontrol sub-model

The life cycle of A. hygrophila is multivoltine and its pupal phase occurs inside

the hollow stems of A. philoxeroides (Maddox 1968). Both the larva and the adult insect

feed on the leaves and the outer part of the stems of the plant, feeding exclusively on this


228

species (Lu et al., 2015). In order to reproduce complete breeding cycles of the biocontrol,

six stages of the species’ life cycle were considered: egg, first instar, second instar, third

instar, pupa, adult males and adult females (Maddox 1968; see Fig. 2). The processes and

factors intrinsically associated to the A. hygrophila population dynamics were based on

existing parameters and equations from several exhaustive studies regarding the species

general demographic and phenological attributes (e.g. Guo et al. 2014; Julien et al. 1995).

The most limiting factor in the life cycle of the biocontrol agent is temperature (Stewart

1996; Stewart et al. 1999), since it determines the number of eggs laid per female, the

duration of each stage of the life cycle and the mortality rates at each stage. In the model

simulations the daily temperature value was stochastically generated among realistic

limits defined by the normal values of maximum and minimum temperatures for the study

area. Since temperature conditions at the beginning of the year were not compatible with

a viable introduction of the biocontrol (as shown in Table 1 from preliminary test model

simulations), the timing of the introduction into the system should consider the days

required for laying eggs, hatching and subsequent maturation of the larvae through the

different stages until becoming adults. The sex ratio between male and female adults was

assumed as 1:1, since in natural circumstances an excess of individuals of one sex is

compensated with the generation of adults of the other sex, maintaining a balanced ratio

(Guo et al. 2014).

2.2.3. Plant-biocontrol interactions

Since resource availability limits the number of viable insects, we assumed

density-dependent mechanisms to model the species population dynamics regarding the

availability of food and leaves for laying eggs. In fact, as the biocontrol is a specialist that

feeds exclusively on A. philoxeroides, food requirements must be fully covered by plant

availability. Food requirements are included in the model based on the daily individual

consumption rates of leaf surface of the plant. In addition, both larvae and adult insects

have the ability to feed on the outside of the stems of the plant if no leaves are available.

This process causes the detachment of the stems and therefore the complete destruction

of the aerial parts of the plant. Both the foliar surface and the surface of the stems were

considered as food sources in the model, resulting in a response of death of the plants due

to damage to the stems if the amount of leaves is insufficient to cover the food needs of

the biocontrol. Moreover, leaves are necessary for the biocontrol egg laying. In spite of

not having exact data of the amount of eggs that a leaf can hold, we calculated the amount

Chapter VII

229

of eggs in each batch, which varies according to the temperature, and we estimate that

one leaf can hold up to three egg batches (Stewart 1996; Stewart et al. 1999). Therefore,

egg laying is hindered in the model if the amount of available leaves is fewer than those

needed.

Figure 2. Conceptual diagram of the dynamic model built for simulate the effectiveness of biocontrol on

the invasive species A. philoxeroides. The model consists of two different dynamic sub-models: the

population dynamics of A. hygrophila through its different life stages (in white) and the invasive plant

vegetative growth (in grey). Both sub-models interact by the herbivory pressure of the different biocontrol

life stages and the respective carrying capacity (expressed in availability of leaves and stems), as indicated

with grey arrows. All these interactions are influenced by the prevailing temperature conditions.

2.2.4. Management scenarios and biocontrol optimization

For demonstration purposes, the management scenarios considered for A.

philoxeroides biocontrol were based in two main complementary factors associated with

the A. hygrophila biocontrol efficacy: i) the number of adult insects released in each

introduction, and ii) the number of introductions made throughout the simulation period.

Each of these factors was classified in three categories: low, medium or high. Therefore,

nine possible scenarios were tested in our model combining the total number of insects

released: low (100 individuals), medium (200 individuals) or high (500 individuals); and


230

the number of introductions: one (single), every two years (biannual) or every year

(annual). The timing selected as the most effective period for biocontrol releases was the

beginning of April (Table 1).

Considering the stochastic seasonal influence of the daily temperatures, for each

of the 9 scenarios a total of 50 independent simulations were carried out, and the

biocontrol efficacy, expressed in terms of A. philoxeroides cover trends, was analysed

along time. For each scenario, the percentage of success (considered as such when the

cover area of A. philoxeroides was reduced or the population eradicated) was calculated

in order to classify the biocontrol global performance as bad (0-33% success), medium

(34-66% success) or good (67-100% success).

3. Results

3.1. Plant development stages and cover

In each simulation, A. philoxeroides’ cover extension potentially increases with a

higher growth rate in the warmest months, mostly between June and September. Aerial

structures always survive from one year to another, since there are no frequent frosts in

the study area. In the absence of biocontrol, the population potentially expands throughout

the simulation period, reaching an extension of 230m2 after 10 years (estimate obtained

from 50 independent simulations). This is a 600% increase from the initial extension

considered, i.e., about 38m2 of plant cover (Fig. 3A). Among the different considered

scenarios, in which different biocontrol intensities have been tested, the representative

results from each simulation, in terms of plant cover trends, are very heterogeneous,

including unchanged (Fig. 3B), relatively controlled (Fig. 3C) or eradicated (Fig. 3D). In

several of these scenarios, an increase in the insect population leads to a rapid decline in

plant cover, which in turn leads to a decline in the insect population due to lack of food.

Depending on the prevailing conditions, if this happens during winter, the combination

of lack of food and low temperatures may be enough to make the population of A.

hygrophila unviable. Fig. 3C shows how the insect population fades at the winter severity

of the 8th year, which allows the recovery of A. philoxeroides. On the other hand, if the

insect survives the winter it may be able to reinforce the biocontrol effort the following

year, which can determine plant eradication, as shown in Fig. 3D (the population of A.

hygrophila does not disappear during winter but only after its food source has exhausted).

Table 1. Number of generations produced by A. hygrophila during each year and the probability that the biological control will survive under the prevailing winter conditions,

depending on the initial number of insects and the time of release from 50 independent model simulations. The temperature range on the release date is expressed in ° C.

Number of insects

100 200 500

Temperature Generations Overwinter % Generations Overwinter % Generations Overwinter %

Rel

ease

dat

e (d

ay) 1 4.0 - 15.1 1 - 2 0 1 - 2 0 1 - 3 0

30 3.8 - 15.7 1 - 3 0 1 - 4 2 1 - 5 1

60 4.5 - 19.7 1 - 4 2 1 - 4 2 2 - 5 5

90 6.1 - 21.9 2 - 5 5 2 - 5 5 3 - 5 10

120 8.1 - 24.7 3 - 5 8 3 - 5 10 3 - 5 20

150 10.4 - 28.2 2 - 4 5 2 - 5 5 2 - 5 10

180 12.3 - 29.2 2 - 3 1 2 - 3 1 2 - 4 2


232

1.1.Biocontrol life cycle and viability

The optimal time of the year predicted to perform the release of insects in the area

is the 120th day of the year (April). If the release is carried out earlier, the eggs may suffer

the negative effects of low temperatures, while if it occurs later the number of generations

completed before winter is reduced (Table 1). The greater the number of insects released,

the greater the probability that they will survive the winter. On average, between three

and five biocontrol generations are produced throughout each year in the study area (as

shown in Fig. 4 for a single simulation). Taking into account the stocacity of the

temperatures generated for each simulation, the probabilities that insects survive winter

are scarce, less than 10% when the initial population size is 100 (estimate obtained from

50 independent simulations of annual releases, Table 1).

1.2. Management scenarios and biocontrol optimization

For each of the nine tested scenarios, 50 independent model simulations were

carried out, each with a simulation period of ten years. A direct relationship was found

between the number of insects released into the environment (100, 200 or 500), the

frequency in which these releases are made (annual, biannual or single release) and the

success in the control or eradication of the invasive species (ranging from 2 to 90% of

biocontrol success in different treatments, Table 2). The greater the number of insects

released (up to 500 individuals), the greater the biocontrol efficacy (maximum values of

12% biocontrol success in a single release, 68% in biannual releases and 90% in annual

releases). All three scenarios involving a single insect release in the environment were

unsuccessful for the control of A. philoxeroides, regardless of the number of insects

released (Table 2). Moreover, the lower the frequency of releases made, the greater the

recovery of the invasive plant population between the successive releases. Thus, annual

releases are clearly the most effective in reducing the population size of A. philoxeroides,

while making a single introduction (in the first year of the simulation) is unlikely to

eliminate the invasive plant, with a 12% biocontrol success in the treatment with 500

insects released (Table 2). It must be taken into account that, even if the aerial parts of

the plant were completely destroyed, they retain their ability to resprout from roots, and

plants may reappear during the following year, especially when the insects had probably

succumbed to winter's cold (as shown in Fig. 3C).

Chapter VII

233

2. Discussion

2.1.Biological invasions and modelling tools

Over the past centuries, human activities have drastically modified the structural

and functional attributes of several ecosystems (Crutzen 2002). In this era of major

environmental changes known as Anthropocene, the role of research and science should

not be limited to describing those changes, but also trying to anticipate and/or early detect

potential threats and propose practical mitigation measures (Chapin III & Fernandez

2013). Biological control has proved to be an effective measure for the control of some

invasive alien species, eliminating or maintaining the density of the target populations

within acceptable limits (DeBach & Rosen 1991; Muniappan et al. 2009; Room et al.

1981). Additionally, the current and first legislation of the European Union on the

prevention and management of the introduction and spread of invasive alien species

(Regulation 1143/2014) emphasizes the importance of prevention and early response to

biological invasions in their initial stages.

Figure 3. Illustrative simulations representing possible A. philoxeroides plant cover trends under different

biocontrol (A. hygrophila) treatment intensities: without releasing individuals (A), annual releases of 100

individuals (B), annual releases of 200 individuals (C) and annual releases of 500 individuals (D). The

relative time expressed in “Days” is counted after the start of the simulation on January 1st (t1).


234

Species distribution models (SDM), also known as ecological niche modelling,

uses algorithms to predict a species' spatial distribution (Elith & Leathwick 2009;

Kearney & Porter 2009). The main advantages of SDMs is their simplicity of use and that

they allow to obtain explicit spatial predictions at regional scales, which makes them a

useful tool for environmental management (Vicente et al. 2019). A relevant example of a

correlative SDM is described by Julien et al. (1995), which uses the climate matching

program CLIMEX to model the global theoretical distribution of A. philoxeroides and A.

hygrophila. However, SDMs do not take into account the dynamic ecological processes

and the stochastic drivers associated with the studied species (Gallien et al. 2012; Henry

et al. 2018).

In this scope, ecological

dynamic models can be seen as

alternative useful tools to

support decision making, since

they allow not only to grasp the

functioning of ecosystems under

different sources of

environmental change, but also

to test alternative measures prior

to their implementation at local

scale where conservation

planning and management

actions usually take place

(Schmolke et al. 2010; Bastos et

al. 2018). Therefore, dynamic modeling has been gradually considered in the simulation

of the key processes that anticipate the response of both invasive species and control

agents, supporting the design of optimized and cost-efficient management strategies

(Eiswerth & Johnson 2002; Godfray & Waage 1991; González et al. 2017; Higgins et al.

1997; Liu et al. 2004). On the other hand, while SDMs allow to evaluate the potential

distribution of invasive species or their biocontrol agents in scenarios of climate change

(Julien et al. 1995; Sun et al. 2017), their deterministic assumptions limit the accuracy in

capturing ecological responses under scenarios of more local changes. Dynamic models

make it possible by predicting the ecological effects that these changes, namely

Figure 4. Illustrative pattern for successive generations of A.

hygrophila obtained from a single simulation, considering a

scenario where 100 individuals have been released each year and

for which food is not a limiting factor, i.e. temperature is the only

constrain affecting population dynamics.

Chapter VII

235

introduced by increasing air temperatures and/or atmospheric concentrations of CO2, will

have on intra-specific relationships and population dynamics (Fu et al. 2016; Palanisamy

2013).

2.2.Testing biocontrol efficacy

Over the last decades, the specialist predator A. hygrophila has been widely used

as a biocontrol agent of A. philoxeroides in countries where the plant causes severe

environmental and economic impacts. Previous experiences show that A. hygrophila can

successfully control the distribution of A. philoxeroides, significantly reducing the size of

populations, as well as their ecological impact (Buckingham 1996; Lin et al. 1984; Ma et

al. 2003b; Sainty et al. 1998). Within this scientific arena, we developed and tested a

dynamic modelling approach to simulate the relationship between an invasive plant

species with a short residence time (A. philoxeroides) and its biocontrol agent (A.

hygrophila) within the framework developed for the region of Fisterra (Spain). Our

approach gives particular attention to the influence of temperature on the growth and

survival of both species. As corroborated in our model simulations, previous experiences

in other countries show that low temperatures are a limiting factor in the biocontrol's life

cycle, which may render this method ineffective for the eradication of A. philoxeroides

(Buckingham 1996; Liu et al. 2010; Ma & Wang 2005). According to our simulation

results, a successful control of the initial stages of invasion of A. philoxeroides would be

possible in the study area, provided that periodic insect releases were made to counteract

the probable fading of A. hygrophila during the winter (as shown in Fig. 4). In fact, with

annual or biannual releases of more than 200 individuals, there are high probabilities of

eradicating or significantly reducing the extent size of the population of A. philoxeroides

in the study area (Table 2). Nevertheless, although the precedents in the use of arthropods

(predators or parasitoids) for the biocontrol of invasive species in the EU (Shaw et al.

2016, 2018), the introduction of biocontrol agents does not always result in the correct

establishment of the species or a success as control treatment (Greathead & Greathead

1992). Therefore, the development of protocols for the evaluation of candidate species,

risk assessments, as well as the use of specific predators, are advisable in order to avoid

consequences on non-target species and habitats (Fowler et al. 2012; Jacas et al. 2006).

This is of particular importance when it comes to the comprehension of the suitable

biocontrol thresholds for conservation purposes.


236

2.3. Added-value, caveats, and way forward

Ecological dynamic models are the most useful tools to explain the functioning of

key ecosystem components, since they are simplified representations of complex

phenomena, which allows us to understand the functioning of those ecosystem processes

that are of interest (Jorgensen & Bendoricchio 2001; Schmolke et al. 2010). In this

perspective, it is important that the modelling conceptualization contains almost all the

key components and parameters that characterize the study processes. However, it is not

always possible to obtain all the necessary data for the model construction and simulation.

Although the parameterization and calibration of our model had an unusual quantity and

quality of information available in the literature, parameters such as Daily resprout rate,

Root resprout rate and Root loss rate have not been obtained by any previous studies, as

far as we know. Moreover, the lack of data about the mortality of A. philoxeroides due to

factors other than longevity, herbivory or freezing temperatures, such as the potential

effects caused by predation and/or competition with native species, should be taken into

account and their respective relevance should be assessed in the future.

Table 2. Different scenarios of biocontrol treatments tested from 50 independent model simulations

(frequency of introductions: annual, biannual, single vs. number of insects released: 100, 200, 500). For

each treatment combination, overall success represents both the eradication of the invasive species and the

decrease of its original extension. Color code: light gray for low overall success, bellow 33%; dim gray for

medium success, between 33% and 66%; dark gray for high success, higher than 66%.

Number of insects released

100 200 500

Fre

qu

ency

of

intr

od

uct

ions

Annual

Overall success (%) 48 80 90

Eradication (%) 42 70 86

Bia

nn

ual



Sin

gle



Chapter VII

237

In conclusion, despite the limitations inherent to an academic demonstration, our

main results show the feasibility of constructing dynamic models by focusing on the

interactions between key components of ecosystems affected by biological invasions.

Since the direct relationships between the invasive species and its biocontrol agent can

only partialy assess the efficacy of biocontrol (e.g. Garay et al. 2015; Henry et al. 2018;

Higgins et al. 1997; Shea & Possingham 2000; Withers et al. 2004), this approach

provides also a useful starting point, allowing the development of more complex models,

with the introduction of other pertinent interactions and interferences with a high

applicability potential (Buchadas et al. 2017; Prentice et al. 2007). In fact, the proposed

dynamic model, where the scenarios are changeable to the universe of application

intended, has enormous potential in other regions where A. philoxeroides is present and

traditional control methods, namely by cutting and removing plant biomass, have been

shown to be ineffective. In fact, our framework is suitable and versatile for management

recommendations in the scope of other biocontrol programs, since it allows to test the

efficacy of alternative treatments and mitigation measures under realistic scenarios

associated with early stages of biological invasions, such the case of A. philoxeroides in

the study area. In this perspective, we highlight the interplay between model-based

research and monitoring, where predictive tools can contribute to an increasing efficiency

and usefulness of biocontrol methods to prevent the severe impacts of biological

invasions.

Acknowledgments

This work was supported by the PhD Program in Marine Science, Technology and

Management (DO*MAR) (granted to R. P.), European funds POPH/FSE and national

funds FCT through the Post-Doc grant SFRH/BPD/84044/2012 (granted to J. R. V.) and

FCT - Portuguese Foundation for Science and Technology, under the project

UID/AGR/04033/2019 and INTERACT – Integrative Research in Environment, Agro

Chains and Technology, under the “Programa Norte 2020, FEDER, Aviso Norte 45-2015-

02” (granted to J. A. C.).


238

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Chapter VII – Annex I

243


Annex 1. Diagram of the biocontrol model.

Part 1. Daily temperature.

Part 2. Time counter since the biocontrol introduction.


244

Part 3. Duration of the biocontrol life stages.


245

Part 4. Adult stage, female (upper) and male (bottom).


246

Part 5. Egg stage.


247

Part 6. Larvae and pupa stages.


248

Part 7. Plant sub-model.


249

Part 8. Insect food requirements.


250

Annex 2. State variables’mathematical equations used for the relationships between the biocontrol agent

(Agasicles hygrophila) and the invasive plant species (Alternanthera philoxeroides) in the Fisterra region

(Spain).

State

variables

Inflows/Outflows Description Process equations

Adult_femal

e(t)

Number of

adult female

insects.

Adult_female(t - dt) + (Female_recruitment

+ Female_introduction - Female_mortality -

F_food_deprivation_mortality) * dt

INIT Adult_female = 0

Female_recruitment

(Inflow)

Adult female

increase due to

successful

pupation.

ROUND((1-Sex_ratio)*Pupa_maduration)

Female_introduction

(Inflow)

Adult female

increase due to

new

introductions.

IF Introduction__option = 1 AND

Introduction__cycle = Introduction__timing

THEN Female_number ELSE 0

Female_mortality

(Outflow)

Adult female

decrease due to

daily mortality

or exceeding

lifespan time.

IF

Female__lifespan_cycle=Average_F__lifes

pan_days-1 THEN Adult_female ELSE

Female_daily_mortality_rate*Adult_female

F_food_deprivation_

mortality (Outflow)

Adult female

decrease due to

insufficient

food available.

Adult_female*Adult_food__losses_%*(1-

Sex_ratio)

Adult_male(t

)

Adult_male(t - dt) + (Male_recruitment +

Male_introduction - Male_mortality -

M_food_deprivation_mortality) * dt

INIT Adult_male = 0

Male_recruitment

(Inflow)

Adult male

increase due to

successful

pupation.

ROUND(Pupa_maduration*Sex_ratio)

Male_introduction

(Inflow)

Adult male

increase due to

new

introductions.

IF Introduction__option = 1 AND

Introduction__cycle = Introduction__timing

THEN Male_number ELSE 0

Male_mortality

(Outflow)

Adult male

decrease due to

daily mortality

or exceeding

lifespan time.

IF

Male__lifespan_cycle=Average_M__lifesp

an_days-1 THEN Adult_male ELSE

Male_daily_mortality_rate*Adult_male

M_food_deprivation

_mortality (Outflow)

Adult male

decrease due to

insufficient

food available.

Adult_male*Adult_food__losses_%*Sex_r

atio

Cumulative_

E_days(t)

Used for egg

hatching time

calculation.

Cumulative_E_days(t - dt) + (E_days) * dt

INIT Cumulative_E_days = 0

E_days (Inflow) Used for egg

hatching time

calculation.

IF Egg_maturation>0 THEN 1 ELSE 0

Cumulative_

I1_days(t)

Used for first

instar

maturation time

calculation.

Cumulative_I1_days(t - dt) + (I1_days) * dt

INIT Cumulative_I1_days = 0

I1_days (Inflow) Used for first

instar

IF First_instar_maturation>0 THEN 1

ELSE 0

Chapter VII – Annex II

251

maturation time

calculation.

Cumulative_

P_days(t)

Used for pupa

maturation time

calculation.

Cumulative_P_days(t - dt) + (P_days) * dt

INIT Cumulative_P_days = 0

P_days (Inflow) Used for pupa

maturation time

calculation.

IF Pupa_maturation>0 THEN 1 ELSE 0

Cumulative_

_F_days(t)

Used for

female adult

insect lifespan

calculation.

Cumulative__F_days(t - dt) + (F_days) * dt

INIT Cumulative__F_days = 0

F_days (Inflow) Used for

female adult

insect lifespan

calculation.

IF F_lifespan_days>0 THEN 1 ELSE 0

Cumulative_

_I2_days(t)

Used for

second instar

maturation time

calculation.

Cumulative__I2_days(t - dt) + (I2_days) *

dt

INIT Cumulative__I2_days = 0

I2_days (Inflow) Used for

second instar

maturation time

calculation.

IF Second__instar_maturation>0 THEN 1

ELSE 0

Cumulative_

_I3_days(t)

Used for third

instar

maturation time

calculation.

Cumulative__I3_days(t - dt) + (I3_days) *

dt

INIT Cumulative__I3_days = 0

I3_days (Inflow) Used for third

instar

maturation time

calculation.

IF Third_instar__maturation>0 THEN 1

ELSE 0

Cumulative_

_M_days(t)

Used for male

adult insect

lifespan

calculation.

Cumulative__M_days(t - dt) + (M_days) *

dt

INIT Cumulative__M_days = 0

M_days (Inflow) Used for male

adult insect

lifespan

calculation.

IF M_lifespan_days>0 THEN 1 ELSE 0

Defoliated_p

lant_cover_

m2(t) =

Area covered

by defoliated

plants, m2.

Defoliated_plant_cover_m2(t - dt) +

(Plant__defoliation - Plant__resprout -

Predation__mortality -

Defoliated_plant_temperature_mortality) *

dt

INIT Defoliated_plant_cover_m2 = 0

Plant__defoliation

(Inflow)

Covered area

daily

defoliated.

Defoliated_area_m2

Plant__resprout

(Outflow)

Area of

defoliated

plants

recovered into

area covered by

plants.

IF

Temperature<Resprout_temperature_thresh

old THEN 0 ELSE

Daily_resprout__rate*Defoliated_plant_cov

er_m2

Predation__mortality

(Outflow)

Area of

defoliated

plants lost due

to herbivory

Cover_loss_m2


252

damage on

shoots.

Defoliated_plant_te

mperature_mortality

(Outflow)

Area of

defoliated

plants lost due

to frost.

IF

Temperature<Plant_temperature_threshold

THEN Defoliated_plant_cover_m2 ELSE 0

Egg(t)

Number of

eggs.

Egg(t - dt) + (Egg_laying - Hatch -

Egg_mortality) * dt

INIT Egg = 0

Egg_laying (Inflow) Egg increment

due to egg

laying.

IF Egg_laying__viability=1 AND

Adult_male >0 THEN

Total_egg_number*Laying_losses_%

ELSE 0

Hatch (Outflow) Egg number

decrease due to

hatching.

IF TIME >= Average__laying_day +

Average_Egg__maturation_days THEN

Egg ELSE Egg_maturation_daily_rate*Egg

Egg_mortality

(Outflow)

Egg number

decrease due to

mortality.

(1-Egg_daily__survival__rate)*Egg

Egg_maturat

ion__cumula

tive_days(t)

Used for egg

hatching time

calculation.

Egg_maturation__cumulative_days(t - dt) +

(Egg_maturation) * dt

INIT Egg_maturation__cumulative_days =

0

Egg_maturation

(Inflow)

Used for egg

hatching time

calculation.

Egg_maturation__days

Female_lifes

pan_accumul

ative_days(t)

Used for

female adult

insect lifespan

calculation.

Female_lifespan_accumulative_days(t - dt)

+ (F_lifespan_days) * dt

INIT Female_lifespan_accumulative_days

= 0

F_lifespan_days

(Inflow)

Used for

female adult

insect lifespan

calculation.

IF Female__lifespan_days>0 THEN

Female__lifespan_days ELSE 0

First_instar(t

)

Number of first

instar larvae.

First_instar(t - dt) +

(First_instar_recruitment -

First_instar_maduration -

First_instar_mortality -

FIrst_instar_food_deprivation_mortality) *

dt

INIT First_instar = 0

First_instar_recruitm

ent (Inflow)

First instar

increment due

to egg

hatching.

Hatch

First_instar_madurati

on (Outflow)

First instar

decrease due to

maturation to

second instar.


First_instar_timer THEN First_instar ELSE

First_instar_maturation_daily_rate*First_in

star

First_instar_mortalit

y (Outflow)

First instar

decrease due to

daily mortality

rate.

First_instar_daily_mortality_rate*First_inst

ar

FIrst_instar_food_de

privation_mortality

(Outflow)

First instar

decrease due to

insufficient

food available.

First_instar*Larva_food_losses_%/3

First_instar_

maturation_c

Used for first

instar

First_instar_maturation_cumulative_days(t

- dt) + (First_instar_maturation) * dt


253

umulative_d

ays(t)

maturation time

calculation.

INIT

First_instar_maturation_cumulative_days =

0

First_instar_maturati

on (Inflow)

Used for first

instar

maturation time

calculation.

First_instar__maturation_days

Laying_begi

nning(t)

Used for time

calculations.

Laying_beginning(t - dt) + (Start_1 -

End_1) * dt

INIT Laying_beginning = 0

Start_1 (Inflow) Starts time

calculation

once the first

egg laying

occurs.

IF Egg_laying >0 THEN TIME ELSE 0

End_1 (Outflow) Restarts the

time counter

when new

females are

recruited from

pupas.

IF Female_recruitment >0 THEN

Laying_beginning ELSE 0

Laying_in_d

ays(t)

Used for time

calculations.

Laying_in_days(t - dt) + (Start_2 - End_2)

* dt

INIT Laying_in_days = 0

Start_2 (Inflow) Starts time

calculation

once the first

egg laying

occurs.

IF Egg_laying >0 THEN 1 ELSE 0

End_2 (Outflow) Restarts the

time counter

when new

females are

recruited from

pupas.

IF Female_recruitment >0 THEN

Laying_in_days ELSE 0

Male_lifespa

n__cumulati

ve_days(t)

Used for male

adult insect

lifespan

calculation.

Male_lifespan__cumulative_days(t - dt) +

(M_lifespan_days) * dt

INIT Male_lifespan__cumulative_days = 0

M_lifespan_days

(Inflow)

Used for male

adult insect

lifespan

calculation.

IF Male_lifespan_days>0 THEN

Male_lifespan_days ELSE 0

Plant_cover_

m2(t)

Area covered

by undefoliated

plants, m2.

Plant_cover_m2(t - dt) + (Clonal__growth

+ Plant__resprout + Root_resprout_2 -

Plant_temperature__mortality -

Plant__defoliation) * dt

INIT Plant_cover_m2 = 0

Clonal__growth

(Inflow)

Increase in area

covered by

undefoliated

plants due to

clonal growth.

IF

Temperature>Lower_temperature__thresho

ld AND

Temperature<Upper_temperature__threshol

d AND

(Plant_cover_m2+Defoliated_plant_cover_

m2)<Suitable_area_m2 THEN

((Plant_cover_m2*Plant_density_individual

s_m2)*Clonal_growth_rate)/Plant_density_

individuals_m2 ELSE 0

Plant__resprout

(Inflow)

Increase in area

covered by

IF



254

undefoliated

plants due to

resprout of

defoliated

plants.

old THEN 0 ELSE

Daily_resprout__rate*Defoliated_plant_cov

er_m2

Root_resprout_2

(Inflow)

Increase in area

covered by

undefoliated

plants due to

resprout from

roots.

Root_resporout_1

Plant_temperature__

mortality (Outflow)

Decrease in

area covered by

undefoliated

plants due to

frost.

IF

Temperature<Plant_temperature_threshold

THEN Plant_cover_m2 ELSE 0

Plant__defoliation

(Outflow)

Decrease in

area covered by

undefoliated

plants due to

herbivory.

Defoliated_area_m2

Pupa(t)

Pupa number. Pupa(t - dt) + (Third_instar_maturation -

Pupa_maduration - Pupa_mortality) * dt

INIT Pupa = 0

Third_instar_maturat

ion (Inflow)

Increase in

pupa number

due to third

instar larvae

maturation.


Third_instar__timer THEN Third_instar

ELSE

Third_instar_maturation_daily_rate*Third_

instar

Pupa_maduration

(Outflow)

Decrease in

pupa number

due to pupa

maturation.


Pupa_timer THEN Pupa ELSE

Pupa_maturation_daily_rate*Pupa

Pupa_mortality

(Outflow)

Decrease in

pupa number

due to pupa

mortality.

Pupa_daily_mortality_rate*Pupa

Pupa_matura

tion__cumul

ative_days(t)

Used for pupa

maturation time

calculation.

Pupa_maturation__cumulative_days(t - dt)

+ (Pupa_maturation) * dt

INIT Pupa_maturation__cumulative_days =

0

Pupa_maturation

(Inflow)

Used for pupa

maturation time

calculation.

Pupa_maturation_days

Root_reserv

oir_m2(t)

Area where the

roots of the

plant remain

after the lost of

aerial

structures, m2.

Root_reservoir_m2(t - dt) +

(Aerial_part_losses - Root_resporout_1 -

Root_loss) * dt

INIT Root_reservoir_m2 = 0

Aerial_part_losses

(Inflow)

Root area

increase due to

decrease in

area covered by

plants, both

defoliated and

undefoliated.

Defoliated_plant_temperature_mortality+Pr

edation__mortality+Plant_temperature__m

ortality

Root_resporout_1

(Outflow)

Root area

decrease due to

plant resprout,

IF



255

i.e. increase in

covered area.

old THEN 0 ELSE

Root_reservoir_m2*Root_resprout_rate

Root_loss (Outflow) Root area loss

due to

mortality.

IF Root_reservoir_m2 <

Root_resprout_required_area_m2 THEN

Root_reservoir_m2 ELSE

Root_loss_rate*Root_reservoir_m2

Second_insta

r(t)

Number of

second instar

larvae.

Second_instar(t - dt) +

(First_instar_maduration -

Second_instar_maturation -

Second_instar__mortality -

Second_instar_food_deprivation_mortality)

* dt

INIT Second_instar = 0

First_instar_madurati

on (Inflow)

Second instar

increment due

to first instar

maturation.


First_instar_timer THEN First_instar ELSE

First_instar_maturation_daily_rate*First_in

star

Second_instar_matur

ation (Outflow)

Second instar

decrease due to

maturation to

third instar.


Second_instar_timer THEN Second_instar

ELSE

Second_instar_maturation_daily_rate*Seco

nd_instar

Second_instar__mort

ality (Outflow)

Second instar

decrease due to

daily mortality

rate.

Second_instar_daily_mortality_rate*Secon

d_instar

Second_instar_food_

deprivation_mortalit

y (Outflow)

Second instar

decrease due to

insufficient

food available.

Second_instar*Larva_food_losses_%/3

Second_insta

r_maturation

_cumulative

_day(t)

Used for

second instar

maturation time

calculation.

Second_instar_maturation_cumulative_day(

t - dt) + (Second__instar_maturation) * dt

INIT

Second_instar_maturation_cumulative_day

= 0

Second__instar_mat

uration (Inflow)

Used for

second instar

maturation time

calculation.

Second_instar_maturation_days

Third_instar(

t)

Number of

third instar

larvae.

Third_instar(t - dt) +

(Second_instar_maturation -

Third_instar_maturation -

Third_instar_mortality -

Third_instar_food_deprivation__mortality)

* dt

INIT Third_instar = 0

Second_instar_matur

ation (Inflow)

Third instar

increment due

to second instar

maturation.


Second_instar_timer THEN Second_instar

ELSE

Second_instar_maturation_daily_rate*Seco

nd_instar

Third_instar_maturat

ion (Outflow)

Third instar

decrease due to

maturation to

pupa.


Third_instar__timer THEN Third_instar

ELSE

Third_instar_maturation_daily_rate*Third_

instar

Third_instar_mortalit

y (Outflow)

Third instar

decrease due to

daily mortality

rate.

Third_instar_daily_mortality_rate*Third_in

star


256

Third_instar_food_d

eprivation__mortalit

y (Outflow)

Third instar

decrease due to

insufficient

food available.

Third_instar*Larva_food_losses_%/3

Third_instar

_maturation_

cumulative_

day(t)

Used for third

instar

maturation time

calculation.

Third_instar_maturation_cumulative_day(t

- dt) + (Third_instar__maturation) * dt

INIT

Third_instar_maturation_cumulative_day =

0

Third_instar__matur

ation (Inflow)

Used for third

instar

maturation time

calculation.

Third_instar_maturation_days

Chapter VII – Annex III

257

Annex 3. Description of parameters, associated variables, composed variables and, when applicable, the

respective sources used into model construction in order to recreate the relationships between the

biocontrol agent (Agasicles hygrophila) and the invasive plant species (Alternanthera philoxeroides).

Unless otherwise is specified, time is expressed in days and temperature in ⁰C.

Parameters/associate

d variables/composed

variables

Description Value / Calculation Source

Adult_food__losses_

%

Estimation of

total adult

insect losses

due to food

shortage.

(Expressed

in %)

IF Total_food__available_% = 1 THEN 0

ELSE (1-Total_food__available_%)

N/A

Annual_cycle Used as daily

counter within

each year of

simulation.

COUNTER(0,365) N/A

Available_food__leav

es_%

Estimation of

total leaf area

available for

insect feeding.

(Expressed

in %)

IF Total_leaf_area_mm2 >

Total_food_needs_mm2 THEN 1 ELSE IF

Total_food_needs_mm2 =0 THEN 0 ELSE

1-(Total_food_needs_mm2-

Total_leaf_area_mm2)/Total_food_needs_

mm2

N/A

Average_Egg__matur

ation_days

Average time

required for egg

maturation.

IF Cumulative_E_days=0 THEN 0 ELSE

ROUND(Egg_maturation__cumulative_da

ys/Cumulative_E_days)

N/A

Average_F__lifespan_

days

Average female

adult insect

lifetime.

IF Cumulative__F_days=0 THEN 0 ELSE

ROUND(Female_lifespan_accumulative_d

ays/Cumulative__F_days)

N/A

Average_leaves_per_p

lant

Average

number of

leaves on each

plant.

300 Field measured

Average_M__lifespan

_days

Average male

adult insect

lifetime.

IF Cumulative__M_days=0 THEN 0 ELSE

ROUND(Male_lifespan__cumulative_days

/Cumulative__M_days)

N/A

Average_Pupa__matur

ation_days

Average time

required for

pupa

maturation.

IF Cumulative_P_days=0 THEN 0 ELSE

ROUND(Pupa_maturation__cumulative_d

ays/Cumulative_P_days)

N/A

Average_shoot__leng

ht_per_plant_cm

Average total

shoot length per

plant, cm.

2500 Field measured

Average__First_instar

__maturation_days

Average time

required for

first instar

maturation.

IF Cumulative_I1_days=0 THEN 0 ELSE

ROUND(First_instar_maturation_cumulati

ve_days/Cumulative_I1_days)

N/A

Average__laying_day Used for

calculation of

first egg laying

time.

IF Laying_in_days =0 THEN 0 ELSE

ROUND(Laying_beginning/Laying_in_da

ys)

N/A

Average__Second_ins

tar__maturation_days

Average time

required for

second instar

maturation.

IF Cumulative__I2_days=0 THEN 0 ELSE

ROUND(Second_instar_maturation_cumul

ative_day/Cumulative__I2_days)

N/A

Average__Third_insta

r__maturation_days

Average time

required for

IF Cumulative__I3_days=0 THEN 0 ELSE

ROUND(Third_instar_maturation_cumulat

ive_day/Cumulative__I3_days)

N/A


258

third instar

maturation.

Clonal_growth_rate Rate of daily

clonal plant

growth.

According to

Shen et al.

2005, the

optimum

temperature for

the

development of

A.

philoxeroides is

30°C. Also,

Julien et al.

1995 indicate

that the limits

for optimal

development of

the plant are

25-32°C.

IF Temperature <25 OR Temperature >32

THEN

(Plant_daily__growth_cm/Average_shoot_

_lenght_per_plant_cm)/2 ELSE

(Plant_daily__growth_cm/Average_shoot_

_lenght_per_plant_cm)

(Julien et al.

1995; Shen et

al. 2005)

Cover_loss_m2 Plant cover lost

due to plant

mortality due to

herbivory, m2.

Devoured_plants/Plant_density_individual

s_m2

N/A

Daily_consumed__lea

ves

Number of

leaves lost due

to insect

feeding.

IF Available_food__leaves_% = 1 THEN

Total_food_needs_mm2/Leaf_area_mm2

ELSE

Total_food_needs_mm2/Leaf_area_mm2*

Available_food__leaves_%

N/A

Daily_defoliated__pla

nts

Number of

plants

defoliated daily

due to

herbivory.

Daily_consumed__leaves/Average_leaves_

per_plant

N/A

Daily_resprout__rate Area covered

by defoliated

plants daily

recovered into

area covered by

plants, m2.

0.05 Estimated

value.

Defoliated_area_m2 Area covered

by plants

transformed

into area

covered by

defoliated

plants due to

herbivory, m2.

Daily_defoliated__plants/Plant_density_in

dividuals_m2

N/A

Devoured_plants Plants dead due

to herbivory

damage on their

shoots.

Food_requirements_mm2/Shoot_area_per_

plant_mm2

N/A

Egg_daily__survival_

_rate

Daily survival

rate for eggs.

IF Temperature <

Egg_survival_min_temperature OR

((1+Egg_mortality_rate)^(1/Egg_maturatio

n__days)-1) <0 THEN 0 ELSE

(1+Egg_mortality_rate)^(1/Egg_maturatio

n__days)-1

N/A


259

Egg_laying_viability Egg laying

feasibility due

to temperature.

IF

Temperature<Egg_laying__min_temperatu

re THEN 0 ELSE 1

N/A

Egg_laying__min_tem

perature

Minimum

temperature

required for egg

laying.

10 (Stewart 1996)

Egg_maturation_daily

_rate

Daily rate for

egg maturation.

IF Average_Egg__maturation_days=0

THEN 0 ELSE

(1+1)^(1/Average_Egg__maturation_days)

-1

N/A

Egg_maturation__day

s

Required time

for egg

maturation,

depending on

temperature.

IF Temperature < 15 THEN 21 ELSE

ROUND (0.1132*Temperature^2 -

6.1776*Temperature + 87.206)

(Stewart et al.

1999)

Egg_mortality_rate Egg mortality,

depending on

temperature.

IF (-0.0064*Temperature^2 +

0.2972*Temperature - 3.077) <0 THEN 0

ELSE -0.0064*Temperature^2 +

0.2972*Temperature - 3.077

(Stewart 1996)

Egg_survival_min_te

mperature

Temperature

threshold for

egg survival.

10 (Stewart 1996)

Egg_survival_max_te

mperature

Temperature

threshold for

egg survival.

27 (Stewart 1996)

Female_daily_mortalit

y_rate

Female adult

insect daily

mortality rate.

IF Female__lifespan_days <= 0 THEN 1

ELSE

(1+Female_mortality_rate)^(1/Female__lif

espan_days)-1

N/A

Female_mortality_rate Female adult

insect mortality

rate when

lifespan time is

exceeded.

1 N/A

Female_number Number of

female adult

insects released

in each

introduction.

Variable N/A

Female__lifespan_cyc

le

Counter of

female adult

insect lifespan

time since

introduction

time.

IF Adult_female >0 THEN

COUNTER(0,Average_F__lifespan_days)

ELSE 0

N/A

Female__lifespan_day

s

Female adult

insect lifespan

time, depending

on temperature.

ROUND (-0.0579*Temperature^3 +

2.1133*Temperature^2 -

17.02*Temperature + 51.188)

(Guo et al.

2012)

First_instar_daily_mor

tality_rate

Daily mortality

rate for first

instar larvae.

(1+Larva_mortality_rate)^(1/First_instar__

maturation_days)-1

N/A

First_instar_maturatio

n_daily_rate

Daily

maturation rate

for first instar

larvae.

IF

Average__First_instar__maturation_days=

0 THEN 0 ELSE

(1+1)^(1/Average__First_instar__maturati

on_days)-1

N/A

First_instar_timer Used for

display

Average_Egg__maturation_days+Average

__First_instar__maturation_days

N/A


260

simplification

of the model.

First_instar__maturati

on_days

Required time

for first instar

larvae

maturation,

depending on

temperature.



3.2562*Temperature + 46.978)

(Stewart et al.

1999)

Food_needs_per_adult

_per_day_mm2

Surface area of

leaf tissue

required daily

for each adult,

mm2.

49.95 (Fu et al.

2016)

Food_needs_per_larva

_per_day_mm2

Surface area of

leaf tissue

required daily

for each larva,

mm2. Estimated

as the average

of the daily

food needs of

the three larval

stages.

19.38 (Fu et al.

2016)

Food_requirements_m

m2

Used for the

estimation of

shoot tissue

area that will be

lost due to

herbivory if

leaves are an

insufficient

food source,

mm2.

IF Available_food__leaves_%<1 THEN

Total_food_needs_mm2-

Total_leaf_area_mm2 ELSE 0

N/A

Introduction__cycle Used for the

implementation

of different

insect

introductions.

COUNTER(0,Introduction__periodicity) N/A

Introduction__option Allows or

prevents insect

introduction on

the simulation.

Variable, 0 or 1 N/A

Introduction__periodic

ity

Number of

insect

introductions

performed

during the

simulation.

Variable N/A

Introduction__timing Day of the year

when the

insects are

introduced on

the system.

Variable, from 1 to 365 N/A

Larva_food_losses_% Estimation of

larva mortality

due to food

shortage.

(Expressed

in %)

IF Total_food__available_% = 1 THEN 0

ELSE (1-Total_food__available_%)

N/A


261

Larva_mortality_rate Larva mortality

rate depending

on temperature.

Is the same

value for the

three larval

stages.

IF (0.0052*Temperature^2 -

0.2447*Temperature + 2.9399) <0 THEN

0 ELSE 0.0052*Temperature^2 -

0.2447*Temperature + 2.9399

(Stewart et al.

1999)

Laying_available__lea

ves_%

Estimation of

leaf availability

for egg laying.

(Expressed

in %)

IF Laying_leaves__required=0 THEN 0

ELSE

Total_leaves/Laying_leaves__required

N/A

Laying_leaves__requir

ed

Number of

leaves required

for egg laying.

We estimate

that one leaf

can hold up to

three egg

batches.

IF Number_of_eggs_per_batch = 0 THEN

0 ELSE

(Total_egg_number/Number_of_eggs_per_

batch)/3

N/A

Laying_losses_% Estimation of

the egg losses

due to not

enough leaves

available for

egg laying.

(Expressed

in %)

IF Laying_available__leaves_% > 1 THEN

1 ELSE Laying_available__leaves_%

N/A

Leaf_area_mm2 Surface area of

each leaf, mm2.

430 (Jia et al.

2010)

Lower__temperature_t

hreshold

Minimum

temperature

required for

clonal growth

of A.

philoxeroides.

We have

selected an

average value

between 7 and

12⁰C, which are

cited by

different

sources.

9.5 (Julien et al.

1995; Shen et

al. 2005)

Male_daily_mortality_

rate

Male adult

insect daily

mortality rate.

IF Male_lifespan_days <= 0 THEN 1

ELSE

(1+Male__mortality_rate)^(1/Male_lifespa

n_days)-1

N/A

Male_lifespan_days Male adult

insect lifespan

time, depending

on temperature.

ROUND (-0.0356*Temperature^3 +

1.2587*Temperature^2 -

9.2283*Temperature + 33.258)

(Guo et al.

2012)

Male_number Number of

male adult

insects released

in each

introduction.

Variable N/A

Male__lifespan_cycle Counter of

male adult

insect lifespan

IF Adult_male>0 THEN

COUNTER(0,Average_M__lifespan_days)

ELSE 0

N/A


262

time since

introduction

time.

Male__mortality_rate Male adult

insect mortality

rate when

lifespan time is

exceeded.

1 N/A

Number_of_eggs_per_

batch

Number of eggs

in each

individual

batch,

depending on

temperature.


5.91*Temperature - 41.85)<0 THEN 0

ELSE (-0.13*Temperature^2 +

5.91*Temperature - 41.85)

(Stewart et al.

1999)

Number_of_eggs__lai

d_per_female

Total number

of eggs each

female adult

insect lays

during its

lifespan,

depending on

temperature.


640.33*Temperature - 6252)<0 THEN 0

ELSE (-14.35*Temperature^2 +

640.33*Temperature - 6252)

(Stewart 1996)

Plant_daily__growth_

cm

Daily growth

for each plant,

expressed as

shoot length

increase in cm.

Using data

from previous

works and field

measurements

of the length of

the stems, we

estimate a

growth of

9.7mm/week in

each stem,

which

translates into

2.8cm/day per

plant taking

into account

that in average

each plant has

20 stems.

2.8 (Pan et al.

2011)

Plant_density_individ

uals_m2

Number of

individual

plants present

per square

meter.

5 Field

measurement.

Plant_temperature_thr

eshold

Temperature

threshold for

survival of

aboveground

plant structures,

frost destroys

aerial parts of

the plant.

0 (Anderson et

al. 2016)

Pupa_daily_mortality_

rate

Daily rate for

pupa mortality.

(1+Pupa_mortality_rate)^(1/Pupa_maturati

on_days)-1

N/A


263

Pupa_maturation_dail

y_rate

Daily rate for

pupa

maturation.

IF Average_Pupa__maturation_days=0

THEN 0 ELSE

(1+1)^(1/Average_Pupa__maturation_days

)-1

N/A

Pupa_maturation_days Required time

for pupa

maturation,

depending on

temperature.



6.6886*Temperature + 104.93)

(Stewart et al.

1999)

Pupa_mortality_rate Pupa mortality

rate.

IF (0.0099*Temperature^2 -

0.454*Temperature + 5.164) <0 THEN 0

ELSE 0.0099*Temperature^2 -

0.454*Temperature + 5.164

(Stewart et al.

1999)

Pupa_timer Used for

display

simplification

of the model.


__First_instar__maturation_days+Average

_Pupa__maturation_days+Average__Seco

nd_instar__maturation_days+Average__Th

ird_instar__maturation_days

Resprout_temperature

_threshold

Temperature

threshold for

the plant

resprout from

roots.

5 (Shen et al.

2005)

Root_loss_rate Rate of loss

from the root

reservoir

0.01 Estimated

value.

Root_resprout_rate Resprout rate of

plants from the

root reservoir.

0.01 Estimated

value.

Root_resprout_require

d_area_m2

Minimum area

of the root

reserve required

for the resprout

to happen, m2.

10 (Clements et

al. 2014)

Second_instar_daily_

mortality_rate

Daily rate for

second instar

mortality.

(1+Larva_mortality_rate)^(1/Second_insta

r_maturation_days)-1

N/A

Second_instar_maturat

ion_daily_rate

Daily rate for

second instar

maturation.

IF

Average__Second_instar__maturation_day

s=0 THEN 0 ELSE

(1+1)^(1/Average__Second_instar__matur

ation_days)-1

N/A

Second_instar_maturat

ion_days

Required time

for second

instar

maturation,

depending on

temperature.




(Stewart et al.

1999)

Second_instar_timer Used for

display

simplification

of the model.



__Second_instar__maturation_days

N/A

Sex_ratio Sex ratio

between male

and female

adult insects.

0.5 (Guo et al.

2014; Maddox

1968)

Shoot_area_per_plant

_mm2

Shoot surface

area for each

individual

plant, mm2.

IF Total_plant_number > 0 THEN

Total_shoot__area_mm2/Total_plant_num

ber ELSE 0

N/A


264

Suitable_area_m2 Suitable area

for plant

growth within

the study area,

m2.

Variable N/A

Temperature For each day of

the year, a

random value is

generated

between the

minimum and

maximum

temperature for

that month.

RANDOM(Minimun_mean__temperature,

Maximum_mean__temperature)

Climatological

data was

obtained from

a weather

station of the

Spanish

Meteorological

Agency

(AEMET),

located at

coordinates

42° 55' 29'' N -

9° 17' 29'' W.

Third_instar_daily_mo

rtality_rate

Daily mortality

rate for third

instar.

(1+Larva_mortality_rate)^(1/Third_instar_

maturation_days)-1

N/A

Third_instar_maturati

on_daily_rate

Maturation

daily rate for

third instar.

IF

Average__Third_instar__maturation_days

=0 THEN 0 ELSE

(1+1)^(1/Average__Third_instar__maturat

ion_days)-1

N/A

Third_instar_maturati

on_days

Required time

for third instar

maturation,

depending on

temperature.




(Stewart et al.

1999)

Third_instar__timer Used for

display

simplification

of the model.



__Second_instar__maturation_days+Avera

ge__Third_instar__maturation_days

N/A

Total_adults Used for the

estimation of

the adult insect

food needs.

Adult_female+Adult_male N/A

Total_adult_food_nee

ds_mm2

Used to

estimate the

leaf area that

adult insects

require daily to

feed, mm2.

Food_needs_per_adult_per_day_mm2*Tot

al_adults

N/A

Total_available_food_

mm2

Total leaf area

and shoot

surface area

used for

calculation of

plant losses due

to herbivory,

mm2. Previous

works estimate

that only 40%

of the shoot

surface area is

available for

insect feeding.

Total_leaf_area_mm2+Total_shoot__area_

mm2*0.4

(Schooler et al.

2006)


265

Total_egg_number Potential

number of eggs

that will be

daily generated

if female adult

insects are

present.

IF Female__lifespan_days > 0 THEN

Adult_female*Number_of_eggs__laid_per

_female/Female__lifespan_days ELSE 0

N/A

Total_food_needs_m

m2

Leaf surface

needed for the

daily feeding of

all adult insects

and larvae,

mm2.

Total_adult_food_needs_mm2+Total_larva

_food_needs_mm2

N/A

Total_food__available

_%

Ratio between

total food needs

and total

available food.

(Expressed

as %)

IF Total_available_food_mm2 >

Total_food_needs_mm2 THEN 1 ELSE IF

Total_food_needs_mm2 =0 THEN 0 ELSE

1-(Total_food_needs_mm2-

Total_available_food_mm2)/Total_food_n

eeds_mm2

N/A

Total_larva Used for the

estimation of

the larva food

needs.

First_instar+Second_instar+Third_instar N/A

Total_larva_food_nee

ds_mm2

Used to

estimate the

leaf area that

larva require

daily to feed,

mm2.

Food_needs_per_larva_per_day_mm2*Tot

al_larva

N/A

Total_leaf_area_mm2 Total leaf

surface area

available for

insect feeding,

mm2.

Leaf_area_mm2*Total_leaves N/A

Total_leaves Total number

of leaves.

Average_leaves_per_plant*Total_plant_wi

th_leaves

N/A

Total_plant_cover_m2 Surface covered

by plants, both

defoliated and

not, m2.

Defoliated_plant_cover_m2+Plant_cover_

m2

N/A

Total_plant_number Total number

of plants.

Plant_density_individuals_m2*Total_plant

_cover_m2

N/A

Total_plant_with_leav

es

Total number

of undefoliated

plants.

Plant_cover_m2*Plant_density_individuals

_m2

N/A

Total_shoot_lenght_c

m

Total length of

shoots in the

population, cm.

Total_plant_number*Average_shoot__leng

ht_per_plant_cm

N/A

Total_shoot__area_m

m2

Total shoot

surface area in

the population,

mm2. The

surface was

calculated as

that of a

cylinder with

height equal to

Total_shoot_le

nght_cm and

0.41cm width.

Total_shoot_lenght_cm*100*1.29 Field

measured.


266

Total_study__area_m2 Total extension

of the study

area where the

A.

philoxeroides

population is

located.

Variable N/A

Upper__temperature_t

hreshold

Maximum

temperature

limit for clonal

growth of A.

philoxeroides.

36 (Julien et al.

1995)

References of Annex 3



Clements, D., Dugdale, T. M., Butler, K. L., & Hunt, T. D. (2014). Management of aquatic alligator weed

(Alternanthera philoxeroides) in an early stage of invasion. Management of Biological Invasions, 5, 327-

339.

Fu, J. W., Shi, M. Z., Wang, T., Li, J. Y., Zheng, L. Z., & Wu, G. (2016). Demography and population

projection of flea beetle, Agasicles hygrophila (Coleoptera: Chrysomelidae), fed on alligator weed under

elevated CO2. Journal of Economic Entomology, 109(3), tow037.

Guo, J. Y., Fu, J. W., Shi, M. Z., Li, J. Y., & Wan, F. H. (2014). Sex ratio effects on copulation, fecundity

and progeny fitness for Agasicles hygrophila, a biological control agent of alligator weed. Biocontrol

Science & Technology, 24(11), 1321-1332.

Guo, J. Y., Fu, J. W., Xian, X. Q., Ma, M. Y., & Wan, F. H. (2012). Performance of Agasicles hygrophila

(Coleoptera: Chrysomelidae), a biological control agent of invasive alligator weed, at low non-freezing

temperatures. Biological Invasions, 14(8), 1597-1608.

Jia, X., Pan, X. Y., Sosa, A., Li, B., & Chen, J. K. (2010). Differentiation in growth and biomass allocation

among three native Alternanthera philoxeroides varieties from Argentina. Plant Species Biology 25(2),

85-92.






and resistance to a stem‐boring insect in Alternanthera philoxeroides. Plant Species Biology, 26(1), 58-

65. doi: 10.1111/j.1442-1984.2010.00307.x

Schooler, S., Baron, Z., & Julien, M. (2006). Effect of simulated and actual herbivory on alligator weed,

Alternanthera philoxeroides, growth and reproduction. Biological Control, 36(1), 74–79.

Shen, J., Shen, M., Wang, X., & Lu, Y. (2005). Effect of environmental factors on shoot emergence and

vegetative growth of alligatorweed (Alternanthera philoxeroides). Weed Science, 53(4), 471-478.



philoxeroides). In Moran V.C., Hoffman, J.H. (Eds.), Proceedings on the IX International Symposium on

Biological Control of Weeds (pp. 393-398). Stellenbosch, South Africa: University of Cape Town.




Conclusions

267

Conclusions

Understanding the underlying mechanisms of biological invasions is key to

predicting future invasion scenarios and to designing efficient strategies for the control

and restoration of invaded areas. The general objective of this doctoral thesis is to

contribute to determine the role that clonal plant growth, and different attributes

associated with it, play in biological invasions. With this aim, a series of experiments

were carried out in which the benefit of various characteristics associated with clonal

reproduction was tested in two invasive species, Carpobrotus edulis and Alternanthera

philoxeroides. The main results of these experiments are described below:

Chapter I

Physiological integration was beneficial for clonal systems of C. edulis when

apical ramets were subjected to a burial stress, allowing the apical ramets of C. edulis to

survive burial, and prevented the loss of biomass for the clonal fragments. In addition, a

non-local plastic response was found in the basal ramets, induced by the conditions

experienced by their corresponding apical ramets. Thus, when apical ramets remained

unburied, there was a division of labor, with basal ramets specializing in the acquisition

of resources through their roots, while apical ramets developed their aerial part. On the

contrary, when apical ramets were buried, basal ramets changed their biomass allocation

pattern and increased the production of photosynthetic structures. In conclusion,

physiological integration may have important consequences for understanding the

invasive success of this clonal species in coastal sand dunes.

Chapter II

The results of the experiment showed the benefit of physiological integration for

both C. edulis and C. acinaciformis when growing in environments with heterogeneous

nutrient distribution. It was found a division of labour between basal ramets, which grew

under high nutrient conditions and developed their root system, and apical ramets, which

grew on dune sand and did not develop roots. The results also indicate that C. edulis may

have a better ability to buffer the negative effect of fragmentation compared to C.

Conclusions

268

acinaciformis. C. edulis has been considered more invasive than C. acinaciformis due to

the wider distribution of the former in Europe. If this is the case, a greater capacity to

withstand fragmentation of the clonal system may be advantageous in already established

populations, favouring the expansion of C. edulis in the habitats it invades.

Chapter III

It was found that the distribution of biomass in response to nutrient availability in

C. edulis differs between native and non-native populations. Thus, plants from the non-

native range (Iberian Peninsula) had a greater plastic response to nutrient scarcity than

plants from the native range (South Africa), consisting of a greater root development, thus

suggesting that this trait has undergone adaptive selection during the invasion process.

This plastic forage response can contribute to the optimization of nutrient uptake by plants

and could therefore be considered as an adaptation strategy. However, this response was

not observed in the other experimental treatments. The lack of response to drought may

be due to the good adaptation of this species to hydric stress, so the short duration of the

experiment did not allow a response to be appreciated. As for the shadow stress, a plastic

response was found consisting on a lesser root development, but this effect was identical

between both populations.

Chapter IV

The results of this experiment indicate the presence of adaptive selection during

the invasion process of C. edulis. Thus, populations from non-native areas (Iberian

Peninsula, California and Australia) showed significantly higher growth in response to an

increase in nutrients than populations from the native range (South Africa). However, the

differences detected in plant growth were not transferred to greater competitive ability in

non-native range populations. On the other hand, a greater benefit was found from the

addition of nutrients, in terms of increased total biomass, in C. edulis from California than

in the less invasive congener C. chilensis, suggesting that the plastic response to soil

nutrient content might explain the differences in invasiveness of both species. On the

other hand, the comparison of competitive ability between congeners did not show a clear

relationship between this trait and the invasiveness of C. edulis. A further study of the

Conclusions

269

relevance of competitive ability in coastal habitats invaded by C. edulis will be necessary

to elucidate the role this trait plays in biological invasions of this species.

Chapter V

After analysing the defensive responses of A. philoxeroides to various real and

simulated herbivory treatments, no increase in defensive chemical compounds (phenols

or tannins) was found. Damage to leaves caused by the specialist predator A. hygrophila

caused the same response in plants as the simulated herbivory treatment which implied

damage to leaves along with the application of jasmonic acid. This response consisted in

an increase of root biomass in the basal part of the clonal systems and was only possible

when physiological integration was maintained throughout the experiment. This response

was positive for the plants, partly compensating for the loss of foliar biomass in the apical

ramets. Thus, the results of the experiment show that jasmonic acid plays a role in the

compensatory response of A. philoxeroides to herbivory, and that this response does not

consist in the production of defensive chemical compounds, but in a non-local change in

biomass allocation (at least in plants from the studied population, which is in the non-

native range, in China). This experiment highlights the importance of physiological

integration in defensive responses to herbivory by clonal plants.

Chapter VI

When studying the epigenetic mechanisms associated with phenotypic plasticity

in A. philoxeroides, a transgenerational effect was found between first-generation plants

that grew under high nutrient conditions and second-generation plants that grew under

low nutrient conditions. In the populations of the native distribution range (Brazil) the

variables associated with growth (number of ramets, stem biomass, root biomass and total

biomass) were affected, while in the population of the non-native distribution range

(Iberian Peninsula) biomass allocation between aerial and underground structures was

altered. The transgenerational effect observed in the populations of the native range may

be due to a "silver spoon" effect (i.e. an advantage due to access to better resources during

an early stage of clonal system development), whereas the observed changes in plants

from the non-native range appear to be regulated by DNA methylation. This experiment

Conclusions

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highlights the importance of transgenerational effects and epigenetic DNA regulations on

the development of an invasive clonal plant, which may help to understand the

mechanisms underlying its invasiveness.

Chapter VII

Finally, one of the most remarkable traits of clonal plants should not be

overlooked: the ability to propagate from vegetative fragments. In this final chapter, a

dynamic simulation model was elaborated for the development of A. philoxeroides in a

model population located in the NW of the Iberian Peninsula. The model also includes

the life cycle of the insect A. hygrophila, a predator of the plant in its native range, which

has been used as a biocontrol agent in several countries. The ability of A. philoxeroides

to resprout once aerial structures have been completely eliminated is, together with the

insect's limited tolerance to cold, the main obstacle for the biological control of A.

philoxeroides. The proposed model allows the development of successful strategies for

the control of this aggressive invasive species, with high applicability potential to other

regions where it is present.

Overall, the different experiments carried out during this doctoral thesis highlight

the benefits of physiological integration in invasive clonal plants, as they allow survival

under conditions of severe stress (such as burial or herbivory). Furthermore, capacity for

phenotypic plasticity shown by our model species is remarkable, including both biomass

partitioning of individual ramets and coordinate division of labour of clonal systems in

response to different environmental factors. This plastic ability, together with the

epigenetic regulation of DNA, can compensate for the lack of genetic variability inherent

in clonal reproduction, turning an apparent handicap into an advantage once the invasive

plant has established itself in a new habitat.

Acknowledgments (Spanish)

271

Agradecimientos

Hay muchas personas sin las cuales esta tesis doctoral no hubiera sido posible, pero

principalmente quiero agradecerle al profesor Sergio Rodríguez Roiloa todo lo que ha

hecho por mí durante estos cuatro años. Ha sido un largo camino desde que empecé mi

TFM hasta que he terminado la primera versión de la tesis y escribo estas líneas. Gracias

por toda la ayuda durante los experimentos y sobre todo por tu positividad y alegría

contagiosa. El laboratorio no sería lo mismo sin ti.

También estoy muy agradecido al profesor Fei-Hai Yu por haberme acogido en su

laboratorio durante mi estancia en China, y por todas las facilidades que tuve durante la

estancia. También por su inestimables comentarios en otros experimentos en los cuales

ha colaborado. Al doctor Bi-Cheng Dong por hacerme un hueco en su despacho y por

toda la ayuda que me ofreció a lo largo del experimento. Gracias también por enviarnos

plantas de China para incluirlas en el experimento de epigenética, una pena que no

llegaron a tiempo. Y por la traducción del abstract de la tesis a chino, claro. A Aini por

todo lo que hizo por mí durante los seis meses que estuve viviendo en Beijing. A Wang,

Bella y Aika por todos los sitios que visitamos juntos. A Victoria, por la música.

A la profesora Dalva Matos por su exagerada amabilidad durante mi estancia en Brasil,

desde la fiesta de bienvenida en su casa a la fiesta de despedida en el laboratorio, y otras

tantas no menos memorables. También por todas las facilidades que tuve en el

experimento y por la oportunidad de realizar trabajos con sus estudiantes. A Rosane y

Augusto por acogerme amablemente en su casa durante los primeros días de mi estancia.

Al resto de alumnos del laboratorio por su ayuda. A todos mis compañeros en la

República A Moita por lo bien que pasamos juntos. A Anabelly por los buenos momentos

y el açaí.

A los profesores João Cabral y Joana Vicente por su inestimable ayuda durante la estancia

en Portugal, tanto dentro como fuera del laboratorio. Llegué allí sin tener ni idea de

simulaciones de ecosistemas y volví con un paper bajo el brazo. A Nazareth e Ivo por su

hospitalidad y por las risas juntos.

A mis compañeros de laboratorio por los buenos momentos en las comidas y las cenas.

Al profesor Rodolfo Barreiro por sus valiosos comentarios en cada uno de los

Acknowledgments (Spanish)

272

experimentos en los que participó y por compartir sus conocimientos en estadística. A

Cris Pardo por el buen rollo y todos los consejos cuando empecé la tesis. A Brais, Ana,

Emma, Érika y el resto de alumnos cuyos TFGs fueron parte de los experimentos de esta

tesis doctoral, por la ayuda.

A todos los amigos que han estado ahí durante estos años. A los Olimos por las cenas y

especialmente a Juanjo por los cafés. Mentiría si dijese que no han sido fundamentales

para mantener la cordura en los largos días haciendo análisis estadísticos o pasando cosas

a limpio. A los demás, perdonad que no os mencione a todos, pero me quedaría sin papel

antes que sin palabras de agradecimiento.

A Daniel, por haber estado ahí desde que empezó esta aventura, allá por el lejano 2010.

Todo lo que dijera se quedaría corto.

A mi familia, por el apoyo y la paciencia en todos estos años. Porque una cosa es que al

niño le dé por estudiar en la Universidad, queda bien para presumir delante de los vecinos

y tal, pero otra muy diferente es que siga estudiando ad infinitum después de la carrera.

De verdad, gracias por la paciencia. En cuanto termine esto del doctorado empezaré a

buscar trabajo (aunque no sea en España). Gracias especialmente a mis abuelos, que

siempre han estado ahí animándome desde pequeño. Y a mi hermana, lo mismo pero

desde que ella era pequeña. Muchísimas gracias por todo.


273


In accordance with the regulations of the Doctoral School of the University of A Coruña,

as part of a doctoral thesis by compendium of research articles, a full copy of the following

articles is included below:

1. Portela, R., & Roiloa, S. R. (2017). Effects of clonal integration in the

expansion of two alien Carpobrotus species into a coastal dune system – a

field experiment. Folia Geobotanica, 52(3-4), 327-335. doi: 10.1007/s12224-

016-9278-4

275

2. Portela, R., Barreiro, R., & Roiloa, S. R. (2019). Biomass partitioning in

response to resources availability: a comparison between native and invaded

ranges in the clonal invader Carpobrotus edulis. Plant Species Biology, 34(1),

11-18. doi: 10.1111/1442-1984.12228 285

3. Portela, R., Dong, B. C., Yu, F. H., Barreiro, R., & Roiloa, S. R. (2019). Effects

of physiological integration on defense strategies against herbivory by the

clonal plant Alternanthera philoxeroides. Journal of Plant Ecology, 12(4),

662-672. doi: 10.1093/jpe/rtz004 293

4. Portela, R., Dong, B. C., Yu, F. H., Barreiro, R., Roiloa, S. R., & Silva Matos,

D. M. (2019). Trans-generational effects in the clonal invader Alternanthera

philoxeroides. Journal of Plant Ecology. doi: 10.1093/jpe/rtz043 (print proof) 305


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Biological invasions are one of the main causes of global biodiversity loss. The reason

why only a few alien species become invasive has yet to be clarified. In this doctoral

thesis, a series of experiments have been conducted to elucidate the role played by

different traits associated with clonal reproduction in biological invasions. In chapters I

and II, field experiments were carried out to investigate the benefit of physiological

integration in Carpobrotus spp. Chapters III and IV delve into the selection of

phenotypic plasticity and the competitive ability of Carpobrotus spp. throughout the

processes of biological invasions. Chapter V focuses on the role of physiological

integration in the defensive response to real and simulated herbivory by the invasive

plant Alternanthera philoxeroides. Chapter VI evaluates the role of DNA methylation as

an epigenetic transmission mechanism of phenotypic plasticity for this species. Finally,

in chapter VII a dynamic simulation model for the biocontrol of A. philoxeroides is

proposed, using the insect Agasicles hygrophila in a model population located in

Fisterra, Galicia (NW Spain).

an experimental approach with Carpobrotus edulis (L.) NE Br.

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