Conséquences Des Exondations Pour Les Communautés Végétales Aquatiques Et Le Fonctionnement Des Zones Humides Fluviales Mélissa De Wilde

Conséquences des exondations pour les communautés végétales aquatiques et le fonctionnement des zones humides fluviales Mélissa de Wilde

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Mélissa de Wilde. Conséquences des exondations pour les communautés végétales aquatiques et le fonctionnement des zones humides fluviales. Biodiversité et Ecologie. Université Claude Bernard - Lyon I, 2014. Français. NNT : 2014LYO10275. tel-01142490

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UNIVERSITE CLAUDE BERNARD LYON 1

ECOLE DOCTORALE Evolution Ecosystèmes Microbiologie Modélisation

Doctorat

Ecologie

Mélissa DE WILDE

Conséquences des exondations pour les communautés végétales aquatiques et le fonctionnement des zones humides fluviales.

Thèse dirigée par Gudrun Bornette et Sara Puijalon.

Soutenue le 12 décembre 2014

Jury :

Rapporteur : François Mesléard, Professeur associé, IUT d’Avignon et Directeur de recherches, Fondation de la « Tour du Valat » Rapporteur : Francisca Aguiar, Chercheur, Institut Supérieur d’Agronomie, Centre de recherche forestière Université de Lisbonne Examinateur : Evelyne Franquet, Professeur, UMR CNRS 7263, IRD 237, Université Aix- Marseille Examinateur : Pierre Marmonier, Professeur, UMR CNRS 5023, Université Claude Bernard Lyon1 Examinateur : Marc Philippe, Maitre de Conférence, UMR CNRS 5276, Université Claude Bernard Lyon1 Examinateur : Arnaud Foulquier, Maitre de Conférence, UMR CNRS 5553, Université Joseph Fourier Grenoble Directrice : Gudrun Bornette Directeur de recherche, UMR CNRS 6249, Université de Franche Comté Directrice : Sara Puijalon Chargé de recherche, UMR CNRS 5023, Université Claude Bernard Lyon 1 UNIVERSITE CLAUDE BERNARD - LYON 1

Président de l’Université M. François-Noël GILLY

Vice-président du Conseil d’Administration M. le Professeur Hamda BEN HADID Vice-président du Conseil des Etudes et de la Vie Universitaire M. le Professeur Philippe LALLE Vice-président du Conseil Scientifique M. le Professeur Germain GILLET Directeur Général des Services M. Alain HELLEU

COMPOSANTES SANTE

Faculté de Médecine Lyon Est – Claude Bernard Directeur : M. le Professeur J. ETIENNE Faculté de Médecine et de Maïeutique Lyon Sud – Charles Directeur : Mme la Professeure C. BURILLON Mérieux Directeur : M. le Professeur D. BOURGEOIS Faculté d’Odontologie Directeur : Mme la Professeure C. VINCIGUERRA Institut des Sciences Pharmaceutiques et Biologiques Directeur : M. le Professeur Y. MATILLON Institut des Sciences et Techniques de la Réadaptation Directeur : Mme. la Professeure A-M. SCHOTT Département de formation et Centre de Recherche en Biologie Humaine COMPOSANTES ET DEPARTEMENTS DE SCIENCES ET TECHNOLOGIE

Faculté des Sciences et Technologies Directeur : M. F. DE MARCHI Département Biologie Directeur : M. le Professeur F. FLEURY Département Chimie Biochimie Directeur : Mme Caroline FELIX Département GEP Directeur : M. Hassan HAMMOURI Département Informatique Directeur : M. le Professeur S. AKKOUCHE Département Mathématiques Directeur : M. le Professeur Georges TOMANOV Département Mécanique Directeur : M. le Professeur H. BEN HADID Département Physique Directeur : M. Jean-Claude PLENET UFR Sciences et Techniques des Activités Physiques et Sportives Directeur : M. Y.VANPOULLE Observatoire des Sciences de l’Univers de Lyon Directeur : M. B. GUIDERDONI Polytech Lyon Directeur : M. P. FOURNIER Ecole Supérieure de Chimie Physique Electronique Directeur : M. G. PIGNAULT Institut Universitaire de Technologie de Lyon 1 Directeur : M. le Professeur C. VITON Ecole Supérieure du Professorat et de l’Education Directeur : M. le Professeur A. MOUGNIOTTE Institut de Science Financière et d'Assurances Directeur : M. N. LEBOISNE Thèse réalisée au sein de l’UMR CNRS 5023 Laboratoire d’Ecologie des Hydrosystèmes Naturels et Anthropisés.

Equipe Ecologie Végétale et Zones Humides Université Claude Bernard-Lyon 1 Bât. Forel 2ème étage 43 Boulevard du 11 novembre 1918 69622 Villeurbanne Cedex France Le remède à l'ennui, c'est la curiosité. La curiosité elle, est sans remède. Résumé

L’objectif de cette thèse a été de mesurer comment les modifications des régimes hydrologiques, plus particulièrement les exondations, régissent 1) certains aspects du fonctionnement des zones humides, 2) l’organisation et la dynamique à court terme des communautés végétales aquatiques et 3) la survie et la réponse plastique des végétaux aquatiques. Cette thèse aborde les questions à des échelles spatio-temporelles différentes. Premièrement, à l’échelle de la décennie, j’ai mesuré la conséquence de la baisse de la hauteur d’eau des zones humides péri-fluviales sur leur physico-chimie. Deuxièmement, à l’échelle de la saison, j’ai mesuré l’influence des caractéristiques sédimentaires des zones humides sur la réponse à court-terme des communautés végétales à l’exondation. Enfin, à l’échelle de quelques semaines, je me suis intéressée à l’aptitude des espèces d’angiospermes aquatiques à mettre en place un ajustement plastique face à l’exondation, en conditions expérimentales de laboratoire et in situ, j’ai recherché les déterminismes de cette réponse (écologiques, morphologiques, phylogénétiques). En termes de caractéristiques physico-chimiques des eaux de surface, l’étude sur 15 ans de la dynamique de zones humides péri-fluviales subissant des exondations, ne conclue pas, comme cela est habituellement décrit dans la littérature, à une augmentation de la teneur en nutriments de la masse d’eau, mais plutôt des changements suggérant des variations du fonctionnement hydrogéologique des zones humides, en faveur d’une plus grande influence de la nappe de versant dans leur alimentation. La réponse in situ des communautés végétales à l’exondation diffère selon le type de sédiment. La résistance et la résilience des communautés décroissent toutes deux avec la capacité de rétention d’eau du sédiment. La capacité des plantes aquatiques à tolérer l’exondation, en conditions expérimentales, semble différer selon leur position phylogénétique, mais pas selon leur forme de croissance (rosettes ou caulescentes). Les espèces tolérant l’exondation montrent des ajustements phénotypiques tels que des organes aériens plus denses et une forte plasticité des feuilles, ce qui peut expliquer le maintien d’un taux de croissance similaire en condition terrestre et aquatique chez ces espèces. La comparaison de phénotypes submerses et émerses in situ, suggère également que l’origine phylogénétique et la niche écologique des espèces (amphiphytes ou hydrophytes) gouvernent l’ajustement plastique et la performance des espèces lors de l’exondation.

Mots-clés : Zones humides, Exondation, Ecologie des communautés, Végétation aquatique, Plasticité, Hydrologie, Caractéristiques physico-chimiques Remerciement

Voici venu le moment délicat de la rédaction des remerciements, moment où après n’avoir que peu dormi depuis quelques semaines on se refait le film de ces 4 belles années avec le risque de ressentir de vives émotions mais également d’oublier certaines personnes…

Je tiens tout d’abord à remercier Gudrun Bornette pour m’avoir proposé cette thèse, pour m’avoir transmis son savoir sur le fonctionnement des Lônes et sur l’intérêt des plantes aquatiques. Une grand merci pour son dynamisme, sa bonne humeur et sa disponibilité, pour m’avoir fait confiance durant ces années et également pour ces corrections drastiques mais efficaces.

Je remercie également Sara Puijalon, pour avoir accepté de prendre le train en marche afin de m’épauler au quotidien sur tous les aspects de la thèse.

Je remercie grandement Francisca Aguiar et François Mesléard d’avoir accepté d’évaluer mon manuscrit de thèse ainsi que Evelyne Franquet, Pierre Marmonier, Arnaud Foulquier et Marc Philippe qui ont bien voulu participer à mon jury.

Je tiens également à remercier Elise Buisson qui m’a encadré durant mon M2 à Marseille, qui m’a réconforté dans mon avis de faire de la recherche et qui m’a toujours suivie et soutenue.

Je remercie tous les membres du LEHNA et particulièrement les membres de l’équipe Ecologie Végétale et Zones Humides, pour leur accueil au sein de l’unité. Merci à l’équipe pédagogique du laboratoire pour m’avoir permis d’enseigner durant ma thèse. Un grand merci à Felix Vallier et Antonin Vienney pour leur aide précieuse durant mes missions commandos sur le terrain et leur aide technique. Un grand merci également à Marie-Rose Viricel. Merci à Nadjette et Abdoulaye pour leur disponibilité et leur aide précieuse concernant toutes les tâches administratives qui prennent tout de même une part significative dans la vie du thésard !!

Merci à tous les non-permanents (Marig, Hélène, Laurent) et les stagiaires (Elena, Axel, Quitterie, Marlène, Carole, Laury, Barbara) pour leur aide sur le terrain et, entre autre, toutes les heures passées à étaler des petites feuilles et sur Winfolia, sans quoi ce travail n’aurait pu exister.

J’ai grandement apprécié la cohésion et la partage quotidiens entre doctorants au sein du laboratoire mais également en dehors et notamment les nombreux apéros forts sympathiques au ToïToï !! Merci aux filles (Célia, Soraya, Barbara, Jehanne) et à l’unique garçon EVZH (Florent). Une pensée pour Célia et Barbara, mes supers colocataires de bureau, qui ont dû me supporter.

Un clin d’œil pour Soraya une super colocataire en toute circonstance ; ainsi que pour Philippe et Cagatay !!

Une pensée également à mes partenaires d’entrainement (Coralie, Nathalie, Raphaël et Pierric) et mon coach (Stéphane) du Fudoshin pour tous les moments de douleur passés sur le tatamis à l’entrainement et en compétition. Grâce à vous j’ai pu me libérer, me défouler et apprendre encore plus que c’est grâce à l’acharnement, grâce au « on lâche rien » qu’on y arrive !! J’espère pouvoir remettre de temps en temps le pyjama !!

Merci à ma famille, mes proches, mes amis pour leur soutien permanant. Vive les Belges, vive les Roumains, les Hollandais, Anglais et Sarladais entre autre !! Je m’excuse auprès d’eux pour mes absences répétées ces derniers temps. Je suis de retour !!

Je tiens tout particulièrement à remercier ma maman qui a toujours été présente tout au long de mes études, qui m’a soutenue durant mes moments de doutes et sans qui je n’aurais pu réaliser mes projets.

Enfin, merci à toi Yannick qui aura été là tout le long de cette thèse, au début contraint et forcé et tu as tellement aimé cela que tu es encore là !! Merci pour ton soutien indéniable dans ce travail, pour également avoir passé des heures à étaler des feuilles et certains dimanches sur le terrain. Merci pour ton soutien au quotidien, tu es un super mari, tu es un super papa !!

A mes cacahuètes. Table des matières PREAMBULE ...... 2 CONTEXTE SCIENTIFIQUE ET OBJECTIFS ...... 3 1. ZONES HUMIDES ...... 3 1.1. Caractéristiques ...... 3 1.2. Régimes hydrologiques ...... 5 1.3. Exondation ...... 6 2. IMPACTS DES EXONDATIONS SUR LA STRUCTURE ET LE FONCTIONNEMENT DES ZONES HUMIDES ...... 8 2.1 Fonctionnement biogéochimique ...... 8 2.1.1 Changements lors de l’exondation ...... 8 2.1.2 Remise en eau ...... 11 2.2 Les végétaux aquatiques – plasticité phénotypique et structuration des communautés ...... 12 2.2.1 Plasticité phénotypique ...... 13 2.2.2 Communautés végétales aquatiques ...... 23 2.2.2.1 Résistance ...... 23 2.2.2.2 Résilience ...... 24 2.2.2.3 Stabilité ...... 25 2.2.2.4 Théorie des perturbations intermédiaires ...... 26 3. OBJECTIFS DU PROJET DE THESE ...... 28 METHODOLOGIE ...... 30 1. SITES D’ETUDES ...... 30 2. PLASTICITE PHENOTYPIQUE ...... 34 2.1 Expérimentations contrôlées vs phénotypage in situ ...... 34 2.2 Traits morpho-anatomiques ...... 34 RESULTATS ...... 36 1. DYNAMIQUE A LONG TERME DU FONCTIONNEMENT DES ECOSYSTEMES...... 36 2. AJUSTEMENT PLASTIQUE DES ESPECES ...... 65 2.1 Mise en place des réponses plastiques...... 65 2.2 Comparaison des phénotypes inondés et exondés ...... 94 3. REPONSES A COURT TERME DES COMMUNAUTES VEGETALES AQUATIQUES...... 130 DISCUSSION ...... 157 1. ASPECTS FONCTIONNELS ...... 157 1.1 Fonctionnement hydrologique et biogéochimique ...... 157 1.2 Structuration des communautés végétales aquatiques ...... 159 2. ASPECTS EVOLUTIFS : AJUSTEMENTS PLASTIQUES DES ESPECES ...... 160 PERSPECTIVES ...... 163 1. ASPECTS FONCTIONNELS ...... 163 1.1 Fonctionnement hydrologique et biogéochimique ...... 163 1.2 Structuration des communautés végétales aquatiques ...... 164 2. ASPECTS EVOLUTIFS : AJUSTEMENTS PLASTIQUES DES ESPECES ...... 165 BIBLIOGRAPHIE – INTRODUCTION, METHODOLOGIE, DISCUSSION ET PERSPECTIVES ...... 167

PREAMBULE

Ce manuscrit s’organise en 5 parties. Dans une première partie, je présente l’ état des connaissances sur le sjuet et les problématiques abordées dans ma thèse. La deuxième partie présente les choix méthodologiques, en se focalisant sur les éléments qui ne sont pas détaillés dans les articles. La troisième partie présente mes résultats au travers 4 articles scientifiques rédigés en anglais. Enfin les quatrième et cinquième parties correspondent à une discussion générale de mes résultats et aux perspectives de mes recherches.

CONTEXTE SCIENTIFIQUE ET OBJECTIFS

1. Zones humides

1.1. Caractéristiques

Il existe de nombreuses définitions du terme « zone humide ». Lors de la Convention de Ramsar sur les zones humides (Ramsar, Iran, 1971) une définition extrêmement large a été établie : « les zones humides sont des étendues de marais, de fagnes, de tourbières ou d’eaux naturelles ou artificielles, permanentes ou temporaires, où l’eau est stagnante ou courante, douce, saumâtre ou salée, y compris des étendues d’eau marine dont la profondeur à marée basse n’excède pas six mètres ». Cette définition reste la référence à l’échelle mondiale car reconnue par un grand nombre d’États, mais elle repose uniquement sur le critère de présence d’eau, ce qui pose de nombreux problèmes, tant en termes de diversité des fonctionnements qu’elles recouvrent, qu’en termes de délimitation, en raison de la variabilité spatio-temporelle de la hauteur d’eau.

De fait, le terme « zone humide » englobe des milieux très contrastés en termes de fonctionnement, de dynamique, et de diversité biologique. Le fonctionnement de ces milieux dépend en particulier de l’hydrologie et de l’hydromorphie du sédiment (Vepraskas et al. 2000 ; van der Valk 2006 ; Keddy 2010) :

1) l’hydrologie : la présence d’eau au moins une partie de l’année et ses fluctuations piézométriques. L’amplitude, la durée et la fréquence des évènements hydrologiques dépendent notamment des conditions climatiques (qui contrôlent l’importance et la répartition des précipitations), des caractéristiques du sédiment (qui contrôlent son ressuyage), et de la topographie (qui contrôle les écoulements en surface et en profondeur) ;

2) l’hydromorphie: la saturation en eau des pores du sédiment sur une période plus ou moins longue de l’année. Cette saturation peut entrainer des conditions d’anaérobiose de certains horizons ou du profil entier en fonction de la hauteur de la colonne d’eau, de son taux

CONTEXTE SCIENTIFIQUE ET OBJECTIFS d’oxygénation et de son taux de renouvellement. Ces variations de l’oxygénation du sédiment influencent fortement les conditions d’oxydo-réduction qui contrôlent les processus biogéochimiques (Vepraskas et al. 2000 ; van der Valk 2006).

Les zones humides sont présentes sous toutes les latitudes et représentent seulement 6% de la surface des terres émergées, et pourtant, figurent parmi les écosystèmes les plus riches et les plus diversifiés de notre planète (RAMSAR 2002), biodiversité regroupant des espèces en général hautement spécialisées, et inféodées à ce type d’habitat (Rascio 2002 ; van der

Valk 2006 ; Keddy 2010). Ces milieux reçoivent une attention soutenue des scientifiques, gestionnaires et politiques car elles rendent de nombreux services écosystémiques (par exemple, approvisionnement en eau de qualité et en nourriture, régulation des inondations, services esthétiques et récréatifs ; Engelhardt & Ritchie 2001 ; RAMSAR 2002 ; Millennium Ecosystem

Assessment 2005 ; Keddy et al. 2009). Dans ce travail, je me suis focalisée sur des zones humides continentales caractérisées par la présence d’eau douce. Elles représentent 90% de la surface totale en zones humides et concentrent 40% de la biodiversité mondiale (RAMSAR

2002).

Les zones humides abritent une flore particulière, car dépendante de la présence d’eau au moins une partie de l’année, la végétation hygrophile, dont les espèces présentent des adaptations leur permettant de s'installer, de croître et de se reproduire dans les sédiments inondés ou saturés en eau de manière permanente ou périodique. Les communautés végétales aquatiques jouent un rôle clé dans le fonctionnement des zones humides et fournissent de nombreux services

écosystémiques (Costanza et al. 1997; Engelhardt & Ritchie 2001, 2002). Les zones humides se caractérisent par les plus fortes valeurs de productivité primaire parmi les écosystèmes continentaux, soulignant leur rôle majeur dans le recyclage du carbone et son piégeage

(RAMSAR 2002 ; Millennium Ecosystem Assessment 2005). La forte densité végétale dans les zones humides contribue également à l’épuration des eaux (vis à vis des nutriments en

CONTEXTE SCIENTIFIQUE ET OBJECTIFS particulier). Lorsque leur densité est suffisante, les végétaux limitent le développement du phytoplancton, et permettent le maintien des eaux limpides dans les écosystèmes aquatiques soumis au phénomène d’eutrophisation (Scheffer et al. 1993 ; Brix 1997). La végétation aquatique influence les conditions environnementales locales physiques : lumière, vent, courant, teneur en oxygène ; ou chimiques : pH, teneur en nutriments, CO2, teneur en carbone organique dissous (Madsen et al. 2001). Elles augmentent fortement la complexité de la structure physique des zones humides, créant ainsi des habitats et des refuges pour les communautés animales et végétales (microphytes), et en tant que premier maillon des chaînes trophiques, elles fournissent de la nourriture pour la faune (Schmieder et al. 2006 ; Keddy

2010).

Les espèces végétales des zones humides présentent des exigences contrastées en ce qui concerne les nutriments (sources de carbone, azote, phosphore, Carbiener et al. 1995 ; Lacoul

& Freedman 2006), et les variations hydrologiques et hydrauliques (Brock & Casanova 1997 ;

Bornette et al. 2001 ; Deil 2005). La composition et la structure des communautés végétales sont par conséquent fortement liées à ces paramètres, et constituent un bon indicateur de fonctionnement de ces écosystèmes (Haury et al. 2006 ; Stelzer et al. 2005).

1.2. Régimes hydrologiques

Les activités humaines, la modification des usages et la demande croissante des ressources en eau combinées au changement climatique affectent l’hydrologie des masses d’eau continentales

(lacs, rivières, zones humides ; Winter 2000 ; Brinson & Malvárez 2002).

La chenalisation, la construction de barrages et la constitution de plan d’eaux sont des causes importante de modification physique de l’hydrologie de ces écosystèmes (Döll et al. 2009). Par ailleurs, l’homme, à travers ses besoins en eau (domestique, agriculture, industrie), exerce une pression croissante sur la ressource en eau via l’augmentation des prélèvements des eaux de surfaces et souterraines, diminuant ainsi sa disponibilité (Acreman et al. 2000 ; Döll et al.

CONTEXTE SCIENTIFIQUE ET OBJECTIFS

2009). Enfin, les activités humaines, via la libération de gaz à effet de serre (hausse de 70% entre 1970 et 2004 ; IPCC 2007), entrainent une augmentation de la température des océans et de l’atmosphère (0,85°C au cours de la période 1880-2012 ; IPCC 2013). Par rapport à la période 1986-2005, les émissions de gaz à effet de serre devraient ainsi augmenter de 25 à 90% entre 2000 et 2030 (SRES 2000), ce qui se traduira par un réchauffement de 0,3 à 0,7 °C d’ici

2035 et de 0,3 à 4,8°C d’ici 2100 (IPCC 2013). L’augmentation de la température entraine des modifications des processus hydrologiques avec notamment une augmentation de l’évaporation des masses d’eau et de l’évapotranspiration des végétaux, entrainant des modifications des régimes de précipitations et des écoulements et par conséquent des changements de la recharge en eau des hydrosystèmes (Arnell 1999 ; Meyer et al. 1999 ; Winter 2000).

Les modèles globaux prenant en compte les scénarios de changements climatiques et socio-

économiques prévoient que ces modifications hydrologiques entraineront une augmentation de la fréquence et de l’intensité des événements extrêmes tels que les inondations et les sécheresses

(Arnell 2004 ; Nohara et al. 2006 ; Alcamo et al. 2007). Cependant l’amplitude et la direction de ces extrêmes hydrologiques devraient varier temporellement et spatialement (Kundzewicz et al. 2008). En région tempérée (moyennes latitudes), les modèles montrent une augmentation de la fréquence, de la durée et de l’intensité des déficits hydriques estivaux (Arnell 1999 ;

Vörösmarty et al. 2000 ; Arnell 2004 ; Lehner et al. 2005 ; Nohara et al. 2006 ; IPCC, 2007), conduisant à des exondations au sein des zones humides plus intenses et plus fréquentes (Winter

2000 ; Erwin 2009 ; Junk et al. 2013).

1.3. Exondation

L’exondation correspond à la disparition de la colonne d’eau. Elle débute par l’étiage (baisse de la hauteur d’eau), et se termine par la remise en eau. L’exondation place donc les organismes benthiques, habituellement présents entièrement ou en partie dans la colonne d’eau, dans de nouvelles conditions environnementales.

CONTEXTE SCIENTIFIQUE ET OBJECTIFS

Les exondations peuvent être considérées comme une perturbation au sens de Pickett & White

(1985) pour les organismes aquatiques, dans la mesure où elles constituent un événement relativement discret dans le temps qui perturbe la structure de l’écosystème, de la communauté ou de la population et qui modifie les ressources, la disponibilité du substrat ou l’environnement physique. Une perturbation peut être décrite par 1) des paramètres spatiaux, par exemple l’aire, la taille et la forme de la surface perturbée (Lake 2000 ; van der Valk 2005); 2) des paramètres temporels, le moment (ou timing) qui précise la date à laquelle l’événement a lieu par rapport au cycle de vie des organismes ou au rythme des saisons, la fréquence, la durée de l’évènement, la vitesse à laquelle la condition perturbante se met en place (Lake 2000 ; van der Valk 2005) ; et 3) l’intensité (sévérité de l’effet), qui peut être appréhendé par la valeur prise par la variable physique associée à la contrainte environnementale, ou par la réponse des communautés vivantes : proportion de biomasse détruite, ou taux de changement des communautés (Grime

1979 ; Sousa 1984 ; White & Pickett 1985 ; White & Jentsch 2001).

L’exondation est une perturbation dont l’intensité augmente avec le temps, et au cours de laquelle les habitats propices et les refuges peuvent être réduit ou éliminés (Lake 2000, 2003 ;

Boulton 2003). Les réponses écologiques peuvent se faire graduellement, ou par étapes lorsque certains seuils hydrologiques sont passés et peuvent varier selon les caractéristiques de l’habitat

(Lake 2000, 2003 ; Boulton 2003 ; Humphries & Baldwin 2003). Lors de l’exondation, le sédiment peut conserver une quantité relativement importante d’eau permettant le maintien de certains organismes benthiques et interstitiels (par exemple les plantes aquatiques, les macro- invertébrés ; Lake 2000, 2003 ; Boulton 2003 ; Brock et al. 2003). La capacité des sédiments à retenir l’eau dépend principalement de leur teneur en matière organique et de leur granulométrie

(Walczak et al. 2002 ; Rawls et al. 2003). Les sédiments riches en matière organique ont une forte capacité à retenir l’eau. Les sédiments limoneux ou argileux du fait de leur densité élevée ont également une capacité à retenir l’eau relativement importante comparée aux sédiments

CONTEXTE SCIENTIFIQUE ET OBJECTIFS grossiers avec des particules de grande taille et une teneur en matière organique faible (Walczak et al. 2002 ; Saxton & Rawls 2006).

2. Impacts des exondations sur la structure et le fonctionnement des zones humides

Dans les zones humides méditerranéennes les exondations sont fréquentes (Humphries &

Baldwin 2003 ; Bond et al. 2008) et les communautés sont adaptées à ces évènements récurrents

(Bonis et al. 1995 ; Boulton 2003 ; Brock et al. 2003 ; Amalfitano et al. 2008 ; Gómez et al.

2012). En revanche, ces exondations imprévisibles entrainent des situations inédites pour la plupart des zones humides continentales situées en région tempérée car celles-ci sont en général en eau permanente tout au long de l’année. Ces exondations peuvent donc avoir des conséquences fortes sur la diversité spécifique, la structure des communautés et le fonctionnement de ces écosystèmes (Greening & Gerritsen 1987 ; Holmes et al. 1999)

2.1 Fonctionnement biogéochimique

Les processus biogéochimiques se déroulant au niveau des sédiments des zones humides dépendent de 1) la disponibilité des composés organiques 2) de la nature de la matrice inorganique 3) de l’oxygénation, 4) de l’humidité et 5) de la température du sédiment (Baldwin

& Mitchell 2000). Lors de l’exondation ces paramètres peuvent être modifiés entrainant des changements des processus biogéochimiques (Qiu & McComb 1996 ; Baldwin & Mitchell

2000 ; Olde Venterink et al. 2002).

2.1.1 Changements lors de l’exondation

Au début d’un épisode d’exondation (étiage), on peut observer au sein d’une zone humide la constitution d’une ou plusieurs vasques et l’exposition à l’air d’une partie du sédiment. Dans les deux cas, une augmentation de la température et un sédiment encore saturé tendent à stimuler la croissance et le métabolisme bactérien, et les activités enzymatiques responsables de la minéralisation de la matière organique (Freeman et al. 1996 ; Baldwin & Mitchell 2000 ;

Koschorreck 2005 ; Racchetti et al. 2011 ; Fig. 1). L’augmentation de la minéralisation dans

CONTEXTE SCIENTIFIQUE ET OBJECTIFS ces deux conditions (vasques et sédiments exondés) entraine la libération d’importantes quantités d’ammonium (via l’ammonification) et de phosphore inorganique dans le milieu (De

Groot & Van Wijck 1993 ; Baldwin 1996 ; Olila et al. 1997). Lorsque l’exondation se prolonge, le sédiment exposé à l’air et oxydé pourrait montrer un taux de minéralisation plus important que dans les vasques, car les microorganismes aérobies utilisent certains substrats organiques

(tels que la lignine) plus efficacement que les microorganismes anaérobies (Baldwin & Mitchell

2000 ; Olde Venterink et al. 2002). Au niveau des sédiments exondés, l’augmentation de la minéralisation de la matière organique engendre une augmentation des émissions de CO2 (De

Groot & Van Wijck 1993 ; Mitchell & Baldwin 1998, 1999).

La teneur en oxygène est physiquement liée à la température de l’eau, ainsi l’augmentation de la température de l’eau fréquemment observée dans les vasques entraine une diminution de sa teneur en oxygène (Wetzel 1983 ; Sipkay et al. 2009). En condition de température élevée, la carence en oxygène augmente également du fait de l’augmentation de l’activité des microorganismes. Dans ce contexte, la décomposition anaérobie de la matière organique conduisant à la libération de méthane (méthanogénèse) peut augmenter.

Lorsque le sédiment voit son exondation s’intensifier, les conditions de températures élevées et la présence importante d’oxygène dans les sédiments exondés, vont stimuler l’oxydation de l’ammonium (nitrification) issu de l’ammonification et donc la libération de nitrates dans le milieu (Olila et al. 1997 ; Olde Venterink et al. 2002 ; Gómez et al. 2012), alors que les conditions anoxiques au sein des vasques entrainent une accumulation d’ammonium. En parallèle, la dénitrification diminue au niveau des sédiments exondés en raison de la disparition des zones anoxiques (Olde Venterink et al. 2002 ; Gómez et al. 2012). Ainsi, une augmentation de la nitrification et une diminution de la dénitrification entrainerait une accumulation de nitrate au niveau des sédiments. Cependant dans les premiers temps de l’exondation, la faible dessiccation du sédiment permet la présence simultanée de zones aérobies et anaérobies. Ainsi

CONTEXTE SCIENTIFIQUE ET OBJECTIFS une diminution de l’azote minéral peut être observée au niveau des sédiments exondés du fait du couplage nitrification-dénitrification (Gómez et al. 2012). Une assimilation des nutriments par les plantes et les microorganismes est possible limitant ainsi leur disponibilité pour le milieu. Cavanaugh et al. (2006) montrent une diminution de la nitrification à cause de la compétition entre les bactéries nitrifiantes et les macrophytes pour l’ammonium.

En ce qui concerne le phosphore, le phosphore inorganique peut être rapidement immobilisé dans le sédiment du fait de son adsorption rapide et difficilement réversible par la phase minérale du sédiment oxydé (Ca principalement dans les sédiments alcalins, Fe et Al principalement dans les sédiments neutres et acides) diminuant ainsi sa disponibilité (De Groot

& Fabre 1993 ; Baldwin 1996 ; Baldwin & Mitchell 2000).

Figure 1. Schéma synthétisant les principaux changements biogéochimiques lors de la mise en place de l’exondation.

Dans le cas d’une exondation très intense et/ou prolongée, la teneur en eau du sédiment diminue et la température augmente fortement, entrainant une diminution de la biomasse et/ou de l’activité microbiennes (De Groot & Van Wijck 1993 ; Larned et al. 2007 ; Austin & Strauss

2011 ; Fig.2). Austin & Strauss (2011) observent une diminution de la nitrification et de la dénitrification pour des teneurs en eau de 15 et 5% respectivement. En cas d’extrême exondation, lorsque le sédiment devient “sec” (teneur en eau inférieure à 5%), une forte mortalité microbienne et donc un relargage d’azote et de phosphore causé par la lyse cellulaire

CONTEXTE SCIENTIFIQUE ET OBJECTIFS peuvent être observés (De Groot & Van Wijck 1993 ; Qiu & McComb 1996 ; Austin & Strauss

2011). L’affinité de la phase minérale du sédiment pour le phosphore tend alors à diminuer car elle devient plus cristalline (Sah et al. 1989 ; Baldwin 1996 ; Baldwin & Mitchell 2000),

3- entrainant ainsi la libération d’une partie du phosphore (PO4 ).

Figure 2. Schéma synthétisant les principaux changements biogéochimiques lors de l’assèchement du sédiment.

2.1.2 Remise en eau

La remise en eau s’accompagne de la diminution de la température de la masse d’eau et du sédiment, et de l’augmentation de l’anaérobiose dans le sédiment.

Immédiatement après la remise en eau, un relargage important de nutriments (nitrate et/ou ammonium et phosphates) dans la colonne d’eau peut avoir lieu et se traduire potentiellement par une eutrophisation du milieu (Olila et al. 1997 ; Corstanje & Reddy 2004 ; Song et al. 2007).

La remise en eau peut en effet solubiliser l’azote et le phosphore issus de la minéralisation et de la mort des microorganismes. Le retour des conditions anaérobies peut également entrainer le relargage des phosphates précédemment adsorbés dans le sédiment oxydé (Ardon et al.

2010). Le relargage des nutriments pourrait être d’autant plus important que la remise en eau

CONTEXTE SCIENTIFIQUE ET OBJECTIFS se ferait par la zone interstitielle comparée à une remise en eau de surface (Corstanje & Reddy

2004).

La forte disponibilité en nutriments peut stimuler la croissance et l’activité microbienne et ainsi maintenir, au moins les premiers temps suivant la remise en eau, une activité de minéralisation et de nitrification importante dans les zones aérobies (Olde Venterink et al. 2002 ; Corstanje &

Reddy 2004), et donc de fortes concentrations en nutriments dans la lame d’eau.

Le retour progressif de conditions anaérobies dans le substrat après la remise en eau, et la forte disponibilité en nitrates devraient permettre la reprise des activités de dénitrification (Olde

Vinterink et al. 2002 ; Corstanje & Reddy 2004) et donc potentiellement une élimination des nitrates et une émission de N2, d’autant plus importantes que la nitrification diminuera après la remise en eau (Cavanaugh et al. 2006).

Figure 3. Schéma synthétisant les principaux changements biogéochimiques lors de la remise en eau après l’exondation.

2.2 Les végétaux aquatiques – plasticité phénotypique et structuration des communautés

Les végétaux aquatiques possèdent de nombreuses adaptations à la vie aquatique (Rascio 2002) qui pourraient limiter leur capacité à tolérer l’exondation. Le passage d’un environnement aquatique à un environnement émergé représente des changements majeurs des conditions subies par les plantes aquatiques: l’eau devient une ressource limitante, la diffusion de la lumière et des gaz n’est plus contrainte, l’accès aux nutriments est modifié du fait de la

CONTEXTE SCIENTIFIQUE ET OBJECTIFS disparition de la colonne d’eau et des changements des processus biogéochimiques intervenant au niveau des sédiments et les contraintes gravitationnelles sont plus fortes (Rattray et al. 1991 ;

Niklas 1998 ; Baldwin & Mitchell 2000 ; Wells & Pigliucci 2000).

Certaines plantes aquatiques ne survivent pas à l’exondation alors que d’autres peuvent montrer différentes stratégies pour faire face à l’exondation (Kautsky 1988 ; Arthaud et al. 2012).

Certaines espèces peuvent établir une banque de graines via la reproduction sexuée de façon récurrente ou lors de l’exondation (Casanova & Brock 2000) ou entrer en dormance dans le sédiment sous forme de propagules végétatives (Barrat-Segretain 2001) qui permettront la résilience des espèces lors de la remise en eau (Brock et al. 2003 ; van der Valk 2005). Certaines espèces peuvent montrer une stratégie d’attente en persistant sous forme “chétive”, sans ajustement morphologique spécifique, leur permettant de survivre un certain temps mais pas de croitre. Enfin, certaines espèces peuvent montrer des ajustements plastiques leur permettant de croitre verticalement et de se reproduire (Bradshaw 1965). Les caractéristiques de la perturbation, la sensibilité et les différentes stratégies mise en place par les espèces face à l’exondation vont déterminer la sensibilité des communauté à court terme et la composition, la structure et la dynamique des communautés à plus long terme.

2.2.1 Plasticité phénotypique

La plasticité phénotypique est la capacité d’un génotype à produire plusieurs phénotypes selon l’environnement dans lequel il se développe (Sultan 2000). Elle est très fréquente chez les plantes aquatiques (Givnish 2002 ; Pigliucci et al. 2006 ; Ghalambor et al. 2007). Ainsi, certaines espèces sont capables de modifier leur phénotype et de produire une forme émergée en modifiant leurs allocations de biomasse aux différents organes et certains traits phénologiques, morphologiques, anatomiques et physiologiques (Bradshaw 1965 ; Sculthorpe

1967 ; Wells & Pigliucci 2000). La plante, en modifiant plusieurs de ces traits qui interagissent les uns avec les autres, peut maintenir sa performance dans les conditions exondées temporaires

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(survie et croissance, Robe & Griffiths 1998 ; Ackerly et al. 2000 ; Romanello et al. 2008 ; Li et al. 2011).

De nombreux ajustements morpho-anatomiques ont été observés chez les Angiospermes aquatiques lorsqu’ils sont soumis à des exondations (Tableau 1):

Régulation des échanges hydriques et gazeux

Les végétaux aquatiques possèdent généralement des feuilles fines dépourvues de cuticule et parfois fortement disséquées favorisant le prélèvement des nutriments et du carbone inorganique dans l’eau (CO2 diffuse 10000 fois plus lentement dans l’eau que dans l’air ; Rascio et al. 1999). La réduction de la cuticule est liée au fait que les organes baignent dans l’eau, et elle s’accompagne en général de la réduction du nombre de stomates. La surface foliaire est généralement importante car la pénétration de la lumière est réduite dans l’eau. La plasticité foliaire est très développée chez les plantes aquatiques avec notamment des possibilités importantes d’ajustements morpho-anatomiques (Bodkin et al. 1980 ; Bruni et al. 1996 ; Wells

& Pigliucci 2000). Certaines espèces d’angiospermes ont la capacité de modifier la morphologie et l’anatomie des feuilles après leur initiation (Golibert 1989 ; Bruni et al. 1996 ;

Kuwabara et al. 2001) ou même de remplacer leurs feuilles aquatiques par des feuilles aériennes

(Robe & Griffiths 1998). En conditions émergées, la surface des feuilles produites tend à diminuer, réduisant les pertes en eau par évapotranspiration (Bruni et al 1996 ; Loreti &

Oesterheld 1996 ; Geng et al. 2006 ; Li et al. 2011). On observe également une modification de la forme des feuilles qui deviennent moins disséquées en milieu terrestre diminuant ainsi les surfaces d’échange (Bruni et al. 1996 ; Schmidt & Millington 1968 in Wells & Pigliucci 2000 ;

Wanke 2011). La taille et la densité des stomates tendent à diminuer chez certaines espèces en condition exondée, limitant par conséquent les pertes en eau par la transpiration (Romanello et al. 2008 ; Yu et al. 2014) alors que pour d’autres espèces une augmentation de la densité des stomates améliorant la régulation des échanges gazeux et hydriques a été observée (Kane &

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Albert 1987 ; Hostrup & Wiegleb 1991 ; Bruni et al. 1996 ; Robe & Griffith 1998 ; Wells &

Pigliucci 2000 ; Germ & Gaberscik 2003 ; Iida et al. 2007). La production d’une cuticule ou l’augmentation de son épaisseur et la diminution de sa perméabilité permettent une meilleur régulation des échanges gazeux et hydriques (Kane & Albert 1987 ; Pedersen & Sand-Jensen

1992 ; Robe & Griffiths 1998 ; Wells & Pigiucci 2000 ; Frost-Christensen et al. 2003)

Absorption de l’eau et des nutriments

La croissance rapide du système racinaire, résultant en un système racinaire plus profond et/ou une augmentation du nombre de poils racinaires, augmentant la surface d’échange racinaire améliorent l’accès à l’eau et aux nutriments en condition exondée (Vasellati et al. 2001 ;

Hussner et al. 2009).

Contrainte gravitationnelle

Les plantes aquatiques ont la spécificité d’avoir un aérenchyme développé facilitant la diffusion des gaz, la respiration racinaire dans les sols anaérobies et la flottaison (Colmer 2003 ; Jackson

& Colmer 2005). La forte densité de l’eau permet le maintien du port de la plante et donc la réduction des tissus de soutien. Certaines espèces peuvent ainsi avoir une tige longue et flexible

(e.g Ranunculus fluitans, du fait de l’absence de tissus de soutien et de la présence d’un aérenchyme) et des feuilles longues ou disséqués offrant une résistance minimale au courant

(Usherwood et al. 1997). En réponse à des contraintes mécaniques différentes en milieu terrestre, la réduction du volume de l’aérenchyme dans l’ensemble des organes (limbe, pétiole, racine ; Hostrup & Wiegleb 1991 ; Loreti & Oesterheld 1996 ; Robe & Griffith 1998 ; Šraj-

Kržič et al. 2006 ; Li et al. 2011) et l’augmentation de la rigidité des organes porteurs par la production de tissus de soutien (tige, feuilles ; Golibert 1989 ; Usherwood et al. 1997 ; Hamann

& Puijalon 2013) a été observée. Les tissus de soutien étant caractérisés par des parois cellulaires épaisses, contenant de la lignine ou de la cellulose, leur présence se traduit donc par une teneur des organes en matière sèche plus élevée (Garnier & Laurent 1994). La lignification

CONTEXTE SCIENTIFIQUE ET OBJECTIFS des tissus augmente leur rigidité et donc leur capacité d’autoportance, et peut également diminuer les risques d’embolisme et améliorer le transport de la sève brute au sein des vaisseaux de xylème (Pedersen & Sand-Jensen 1993 ; Vasellati et al. 2001). Au niveau des feuilles, une augmentation de l’épaisseur du limbe est souvent observée en conditions exondée (Bruni et al.

1996 ; Iida et al. 2007), correspondant à une augmentation de la rigidité des tissus et à des changements structurels du mésophylle, alors que chez d’autres espèces aucun changement n’est observé (Germ & Gaberscik 2003). La diminution de la surface foliaire et une augmentation de la teneur en matière sèche, résultant en une diminution de la surface foliaire spécifique (Iida et al. 2007) s’accompagne d’une augmentation de la durée de vie des feuilles en condition terrestre, améliorant ainsi l’utilisation des ressources (Li et al. 2011).

L’augmentation de la contrainte gravitationnelle peut également entrainer une diminution de la longueur du pétiole et/ou du limbe (Iida et al. 2007). Certaines études mettent en évidence une diminution de la longueur des entrenœuds chez les plantes soumises à l’exondation (Geng et al.

2006 ; Iida et al. 2007 ; Hussner & Meyer 2009) ou de la hauteur de la plante ou des organes porteurs (Loreti & Oesterheld 1996 ; Hamann & Puijalon 2013).

Allocation de biomasse

Certaines de ces modifications entrainent des changements d’allocation de biomasse. Chez certaines espèces, une augmentation de l’allocation au système racinaire peut optimiser le prélèvement ou le stockage des ressources souterraines (eau, nutriments ; Li et al. 2011). Une augmentation du ratio biomasse souterraine/ biomasse aérienne peut également correspondre à une augmentation de l’allocation aux organes de reproduction végétative (rhizome par exemple) qui permettra la production de nouvelles tiges et feuilles lors du retour des conditions favorables

(Geng et al. 2006 ; Touchette et al. 2008). La diminution de la biomasse aérienne peut aussi

être interprétée comme une stratégie pour diminuer les pertes en eau en condition terrestre

(Romanello et al. 2008). Par ailleurs, une augmentation de l’allocation de biomasse à la partie

CONTEXTE SCIENTIFIQUE ET OBJECTIFS souterraine mène à une diminution de la surface photosynthétique et donc potentiellement à la réduction du taux de croissance.

CONTEXTETableau 1. SCIENTIFIQUEPrincipaux ajustements ET OBJECTIFS morpho-anatomiques observés chez les Angiospermes aquatiques lorsqu’ils sont soumis à des exondations

Stratégies Contraintes Variation des traits Espèces Forme de Phylogénie Niche Références croissance écologique Résistance Régulation échanges Réduction surface foliaire Nymphoides peltata Rosette Eudicot Astérales Hydrophyte Li et al. 2011 gazeux et hydriques Menyanthaceae Alternanthera Caulescente Eudicot Caryophyllales Amphiphyte Geng et al. 2006 philoxeroides Amaranthaceae Paspalum dilatatum Rosette Monocot Poales Poaceae Hélophyte Loreti & Oesterheld 1996 Ranunculus Caulescente Eudicot Ranunculales Amphiphyte Bruni et al. 1996 flabellaris Ranunculaceae Réduction biomasse aérienne Acorus americanus Rosette Monocot Acorales Hélophytes Romanello et al. 2008 Acoraceae Réduction taille et densité des stomates Acorus americanus Rosette Monocot Acorales Hélophytes Romanello et al. 2008 Acoraceae Nymphoides peltata Rosette Eudicot Astérales Hydrophyte Yu et al. 2014 Menyanthaceae Augmentation densité des stomates Ranunculus Caulescente Eudicot Ranunculales Amphiphyte Bruni et al. 1996 flabellaris Ranunculaceae Littorella uniflora Rosette Eudicot Lamiales Amphiphytes Hostrup & Wiegleb 1991; Plantaginaceae Robe & Griffith 1998 Hippuris vulgaris Caulescente Eudicot Lamiales Hydrophyte Kane & Albert 1987; Wells Plantaginaceae & Pigliucci 2000 Ranunculus Caulescente Eudicot Ranunculales Hydrophyte Germ & Gaberscik 2003 trichophyllus Ranunculaceae Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 malaianus Potamogetonaceae Production ou augmentation épaisseur Hippuris vulgaris Caulescente Eudicot Lamiales Hydrophyte Kane & Albert 1987 cuticule et diminution perméabilité Plantaginaceae Lobelia dortmanna Rosette Eudicot Asterales Amphiphyte Pedersen & Sand-Jensen Campanulaceae 1992 Mentha aquatica Caulescente Eudicot Lamiales Amphiphyte Frost-Christensen et al. Lamiaceae 2003 Littorella uniflora Rosette Eudicot Lamiales Amphiphyte Robe & Griffith 1998 Plantaginaceae Absorption eau et Prolifération système racinaire Myriophyllum Caulescente Eudicot Saxifragales Amphiphyte Hussner et al. 2009 nutriments aquaticum Haloragaceae Paspalum dilatatum Rosette Monocot Poales Poaceae Hélophyte Vasellati et al. 2001 Littorella uniflora Rosette Eudicot Lamiales Amphyphite Robe & Griffith 1998 Plantaginaceae Augmentation allocation biomasse Nymphoides peltata Rosette Eudicot Astérales Hydrophyte Li et al. 2011 système racinaire Menyanthaceae Gravité Réduction volume aérenchyme Littorella uniflora Rosette Eudicot Lamiales Amphiphyte Hostrup & Wiegleb 1991, Plantaginaceae Robe & Griffith 1998 Paspalum Rosette Monocot Poales Poaceae Hélophyte Loreti & Oesterheld 1996 dilatatum Nymphoides peltata Rosette Eudicot Astérales Hydrophyte Li et al. 2011 Menyanthaceae Myosotis Caulescente Eudicot Boraginaceae Amphiphyte Šraj-Kržič et al. 2006 scorpioides

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Production tissus de soutien Hippuris vulgaris Caulescente Eudicot Lamiales Hydrophyte Golibert 1989 ; Plantaginaceae Hamann & Puijalon 2013 Mentha aquatica Caulescente Eudicot Lamiales Amphiphyte Hamann & Puijalon 2013 Lamiaceae Myosotis Caulescente Eudicot Boraginaceae Amphiphyte Hamann & Puijalon 2013 scorpioides Augmentation épaisseur feuilles Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 malaianus Potamogetonaceae Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 anguillanus Potamogetonaceae Ranunculus Caulescente Eudicot Ranunculales Amphiphyte Bruni et al. 1996 flabellaris Ranunculaceae Diminution longueur pétiole et/ou du Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 limbe malaianus Potamogetonaceae Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 anguillanus Potamogetonaceae Berula erecta Rosette Eudicot Apiales Apiaceae Amphiphyte Hamann & Puijalon 2013 Mentha aquatiqua Caulescente Eudicot Lamiales Amphiphyte Hamann & Puijalon 2013 Lamiaceae Diminution longueur entre nœud/ hauteur Alternanthera Caulescente Eudicot Caryophyllales Amphiphyte Geng et al. 2006 philoxeroides Amaranthaceae Hydrocotyle Caulescente Eudicot Apiales Araliaceae Amphiphyte Hussner & Meyer 2009 ranunculoides Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 malaianus Potamogetonaceae Potamogeton Caulescente Monocot Alismatales Hydrophyte Iida et al. 2007 anguillanus Potamogetonaceae Paspalum dilatatum Rosette Monocot Poales Poaceae Hélophyte Loreti & Oesterheld 1996 Hippuris vulgaris Caulescente Eudicot Lamiales Hydrophyte Hamann & Puijalon 2013 Plantaginaceae Régénérative Augmentation du ratio biomasse Peltandra virginica Rosette Monocot Alismatales Amphiphytes Touchette et al. 2008 souterraine/ biomasse aérienne Araceae Alternanthera Caulescente Eudicot Caryophyllales Amphiphyte Geng et al. 2006 philoxeroides Amaranthaceae

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Cette synthèse bibliographique montre qu’un nombre relativement important de travaux portent sur l’étude de la réponse plastique des plantes aquatiques faisant face à l’exondation. Cependant les travaux existant sont souvent des études de cas, décrivant des espèces isolées, dans des conditions expérimentales très différentes d’une étude à l’autre, et étudient des traits qui peuvent être très variés, rendant difficile l’établissement de règles générales en ce qui concerne les patrons de réponse des végétaux à la contrainte. De surcroit, le déterminisme et la valeur adaptative de cette réponse ne sont que rarement abordés.

Patrons de réponse et déterminisme de la plasticité

Certaines études montrent ou suggèrent que la réponse des plantes à certaines contraintes

(courant, exondation, vent) peut différer selon leur forme de croissance car les organes les subissant diffèrent (Puijalon et al. 2005 ; Puijalon et al. 2011 ; Hamann & Puijalon 2013). Les plantes aquatiques présentent des formes de croissance différentes (rosette vs. caulescente).

Pour les espèces aquatiques présentant un port en rosette, la capacité à produire un phénotype terrestre repose principalement sur la production de feuilles autoportantes (Hamann & Puijalon

2013). La tige très courte portant le méristème apical ne subit pas ou peu la gravité et du fait de son enfouissement partiel dans le sédiment reste protégé de la dessiccation, ce qui devrait permettre à la plante de produire rapidement des feuilles adaptées aux conditions terrestres

(Wells & Pigliucci 2000 ; Grime & Mackey 2002 ; Puijalon & Bornette 2006). Pour les espèces caulescentes, la production d’une tige autoportante est nécessaire pour maintenir une posture

érigée et permettre un positionnement efficace des feuilles par rapport à la lumière (Hamann &

Puijalon 2013). Les tiges tombent sur le sédiment lors de l’exondation, et si elles survivent, doivent émettre secondairement une nouvelle tige auto-portante à partir du bourgeon terminal ou d’un bourgeon axillaire (Shimizu-Sato & Mori 2001).

Des études comparatives de la plasticité en fonction de la position phylogénétique ont été réalisées (Pigliucci et al. 1999 ; Van Buskirk 2002 ; Mommer et al. 2006 ; Richter-Boix et al.

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2006). L’ensemble des ces études montrent que les espèces (animales et végétales) ont des patrons de réponses similaires et indépendants de la phylogénie. Cependant l’amplitude de la réponse plastique peut différer selon la niche des espèces, les espèces colonisant une grande variété d’habitat ou des habitats imprévisibles montrent une forte plasticité, suggérant une plasticité adaptative (Van Buskirk 2002 ; Mommer et al. 2006 ; Richter-Boix et al. 2006). La plasticité phénotypique des Angiospermes aquatiques n’a jamais été abordée dans une étude comparative, or les différentes espèces présentent des origines phylogénétiques et des histoires

évolutives variées ainsi que des niches écologiques différentes (hydrophytes qui sont strictement aquatiques, qui développent la totalité de leur appareil végétatif dans l'eau ou à la surface vs. amphiphytes qui passent au moins une partie de leur cycle de vie dans l’eau, capables de supporter l’exondation). Les plantes aquatiques montrent une évolution convergente et parallèle pour les traits liées à la vie aquatique au sein de nombreux groupes d’Angiospermes

(Les et al. 1991 ; Cook 1999 ; Fig. 4). Le passage de la vie terrestre à la vie aquatique semble s’être fait à plusieurs reprises au cours de l’évolution (Cook 1999 ; Rascio 2002). Pour certains groupes, il est probable que plusieurs passages entre vie terrestre et aquatique aient eu lieu au cours de leur histoire évolutive (e.g Ranunculus, Johansson 1998 ; Cook 1999). La vie aquatique commence très tôt dans l’histoire évolutive des angiospermes (Friis et al. 2003 ;

Soltis et al. 2008) comme le montre l’existence des groupes basaux fossiles et actuels, entièrement ou majoritairement aquatiques (e.g Archaefructaceae, Nympheales,

Ceratophyllales, Alismatales ; Cook 1999 ; Sun et al. 2002). Il existe également des groupes plus récents comprenant à la fois des espèces terrestres et aquatiques (e.g. Apiales). La présence d’espèces complètement ou partiellement aquatiques, isolées dans des groupes taxonomiques majoritairement terrestres, suggère un passages récent à la vie aquatique. Au contraire, des espèces appartenant à des groupes taxonomiques (ordre ou famille) entièrement aquatiques, suggèrent que les groupes taxonomiques auxquels elles appartiennent sont retournés

CONTEXTE SCIENTIFIQUE ET OBJECTIFS précocement à la vie aquatique dans leur histoire évolutive, peu après la divergence de leur lignée (Cook 1999 ; Chambers et al. 2008).

Figure 4. Arbre phylogénétique des ordres et quelques famille des angiospermes (APGIII 2009). Les traits rouges pleins indiquent les groupes entièrement aquatiques, le trait rouge pointillé indique un groupe majoritairement aquatique et les traits bleus pleins indiquent les groupes majoritairement terrestres comprenant des taxons aquatiques.

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2.2.2 Communautés végétales aquatiques

La stabilité des communautés subissant des perturbations est reliée à leur capacité de résistance et de résilience (Leps et al. 1982). La résistance d’une communauté représente sa capacité à tolérer une perturbation, sans montrer de changements significatifs dans sa structure et sa composition (Leps et al. 1982 ; Lake 2000). La résilience d’une communauté fait référence à sa capacité à revenir à l’état pré-perturbation et la vitesse avec laquelle ce processus se réalise

(Leps et al. 1982 ; Pimm 1984, 1991 ; Westman 1986 ; Lake 2000). Les communautés montrant une forte résistance ou une forte résilience aux perturbations sont plus stables.

La dynamique, la structure et la composition des communautés végétales aquatiques dépendent de la niche des individus adultes (effet de l’environnement sur la survie et la croissance des adultes), et de la niche de régénération (recrutement ; effet de l’environnement sur la dispersion, la germination et la mortalité des germinules) des différentes espèces (Grubb 1977 ; Seabloom et al. 1998, 2001 ; Van der Valk 1981, 2005; Van Geest et al. 2005) ; c’est-à-dire des conditions environnementales historiques et actuelles. Ainsi les caractéristiques des cycles exondation/ remise en eau tel que leur amplitude, leur rythmicité et leur durée peuvent mener à différentes réponse des communautés végétales en terme de dynamique, de composition et de richesse spécifique (Van der Valk 1981, 2005 ; Wilcox & Meeker 1991 ; Riis & Hawes 2002 ; Van

Geest et al. 2005 a, b ; Wilcox & Nichols 2008).

2.2.2.1 Résistance

La résistance des communautés végétales aquatiques face à l’exondation repose sur l’aptitude des espèces à la tolérer via notamment des ajustements phénotypiques (Bradshaw 1965 ;

Sculthorpe 1967 ; Robe & Griffiths 1998 ; Wells & Pigliucci 2000). Lorsque l’exondation atteint un certain seuil d’intensité, la résistance des communautés aquatiques peut décroitre, et la perturbation peut entrainer leur destruction partielle ou totale (Boschilla et al. 2012).

L’exondation peut changer les conditions environnementales et lever la dormance des graines.

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Ainsi certaines graines vont pouvoir germer lorsqu’une trouée dans la végétation sera créée, augmentant ainsi la quantité de lumière ou la température au niveau du substrat menant à la germination (Seabloom et al. 1998). Dans ce cas, l’établissement d’espèces amphiphytes rudérales facultatives et d’hélophytes peut se faire à partir de la banque de propagules ou par dispersion (Coops & van der Velde 1995 ; Abernethy & Willby 1999 ; Hudon 2004 ; Havens et al. 2005).

2.2.2.2 Résilience

Lors de la remise en eau, la résilience de la communauté peut se faire grâce à la banque de propagules (germination de graines et croissance de propagules végétatives tels que rhizomes, bourgeons, turions, fragments non spécialisés ; Barrat-Segretain & Amoros 1996 ; Barrat-

Segretain et al. 1999 ; Brock et al. 2003 ; Combroux & Bornette 2004), par recolonisation à partir de zones moins perturbées (effet bordure), ou par dispersion de propagules provenant d’autres écosystèmes (Barrat-Segretain 1996). Cette dernière option nécessite l’existence de connexions permanentes ou temporaires entre les milieux aquatiques. La résilience de la communauté va dépendre également de la capacité des plantes établies, pendant la phase d’exondation, à supporter la remise en eau. Certaines espèces d’amphiphytes, d’hélophytes ou des espèces terrestres ayant été recrutés pendant l’exondation peuvent tolérer des périodes d’inondation plus ou moins longues (Casanova & Brock 2000).

Dans les milieux aquatiques permanents, la banque de graines n’est pas considérée comme essentielle dans la dynamique des communautés, du fait de l’importance de la multiplication végétative chez les végétaux aquatiques (Combroux et al. 2001 ; Capers 2003). Ainsi si l’intensité et la fréquence des exondations sont suffisamment modérées pour ne pas détruire les propagules végétatives, les plantes peuvent rapidement régénérer à partir de ces dernières lors de la remise en eau (Doyle & Smart 2001 ; Liu et al. 2006). La part de propagules végétatives dans la banque peut s’accroitre pour les espèces ne tolérant pas l’exondation et produisant des

CONTEXTE SCIENTIFIQUE ET OBJECTIFS propagules végétatives leur permettant d’entrer en dormance pour passer l’épisode d’exondation. Cependant la baisse de hauteur d’eau au cours de l’exondation pourrait favoriser la floraison et donc la production de graines pour certaines espèces menant à leur accumulation dans la banque. De plus le recrutement de certaines espèces submerses requiert une période d’exondation (Casanova & Brock 2000 ; Bonis & Grillas 2002 ; Warwick & Brock 2003) permettant d’augmenter leur biomasse et leur étendue lors de la remise en eau (Havens et al.

2005). Dans les zones humides temporaires en climat méditerranéen, les communautés sont dominées par des espèces annuelles (Aponte et al. 2010). Les communautés présentent majoritairement des stratégies de résilience, reposant largement sur la reproduction sexuée

(Grillas et al. 1993 ; Bonis et al. 1995 ; Brock & Rogers 1998 ; Casanova & Brock 1996 ; Brock et al. 2003).

2.2.2.3 Stabilité

Van der Valk (1981, 2005) a développé un modèle décrivant 2 types de changement de la végétation au cours des cycles exondation/remise en eau : la fluctuation qui correspond à un changement de l’abondance relative des différentes espèces et la succession qui correspond à un changement de la composition spécifique. La succession peut être cyclique avec l’alternance de communautés différente entre la phase aquatique et exondée.

Lorsque la communauté végétale aquatique montre une forte résistance à l’exondation, peu de changements sont observé (fluctuations) et cette résistance permet la stabilité de la communauté au fil du temps. Si la communauté montre une faible résistance à l’épisode d’exondation mais une résilience importante lors de la remise en eau, van der Valk parle de succession cyclique, et celle ci permettra le maintien de la communauté au fil du temps (Van der Valk 1981 ; Stroh et al. 2008). Une faible résistance accompagnée d’une faible résilience peut mener à un changement important de la communauté végétale avec notamment l’augmentation de la part

CONTEXTE SCIENTIFIQUE ET OBJECTIFS d’espèces émergentes comme les hélophytes, ou d’espèces terrestres tolérant l’inondation (Stroh et al. 2008).

2.2.2.4 Théorie des perturbations intermédiaires

Selon la théorie des perturbations intermédiaires (Connell 1977 ; Huston 1979), la richesse spécifique est la plus élevée pour les perturbations d’intensité et de fréquence intermédiaires. En cas de perturbations peu intenses et/ou peu fréquentes, les espèces compétitives dominent la communauté et empêchent le recrutement de nouvelles espèces

(exclusion compétitive), avec pour conséquence une spécialisation des niches et une richesse spécifique faible. Dans le cas de perturbations intenses et fréquentes, seules les espèces rudérales persistent, conduisant à une faible spécialisation des niches et une faible richesse spécifique. Dans le cas de perturbations intermédiaires en fréquence et en intensité, un équilibre dynamique est atteint, permettant la coexistence d’espèces compétitives, rudérales, et d’aptitude compétitive intermédiaire, conduisant à une richesse spécifique maximale et l’établissement d’une mosaïque changeante. Dans le cas d’exondations peu fréquentes et/ou intenses, les hydrophytes et quelques amphiphytes dominent la communauté ; une faible richesse spécifique est alors observée (Fig 5). Dans le cas d’exondations d’intensité et/ou de fréquence intermédiaires, on attend ainsi une richesse spécifique élevée du fait de la coexistence d’espèces hydrophytes et amphiphytes tolérant l’exondation, et d’espèces recrutées pendant les phases d’exondation (Abernethy & Willby 1999 ; Havens et al. 2005 ; Fig 5). Dans le cas d’exondations intenses et/ou fréquentes, seulement certaines espèces d’amphiphytes et les hélophytes seront capables de survivre, conduisant à une faible richesse spécifique (Fig 5).

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Figure 5. Distribution théorique de la richesse spécifique le long d’un gradient de fréquence et d’intensité de perturbations par exondations (modifié de Grime 2002).

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3. Objectifs du projet de thèse

Cette thèse s’inscrit dans le programme ANR WETCHANGE dont l’objectif était d’établir des prévisions à l’horizon 2030 - 2050 de la réponse des zones humides péri-fluviales aux exondations induites par le changement global, en utilisant des descripteurs physiques

(fonctionnement hydrologique et caractéristiques physiques et chimiques des habitats) et biologiques (communautés végétales, animales et microbiennes).

L’objectif de cette thèse a été de comprendre comment les déficits hydriques saisonniers (liés au changement climatique et/ou aux activités anthropiques) régissaient le fonctionnement des zones humides, et la structure et la dynamique des communautés végétales aquatiques. Cette thèse posait 3 grandes questions qui se situent à des échelles spatio-temporelles différentes auxquelles j’ai répondu par des approches corrélatives et expérimentales.

La première question se situe à l’échelle de l’écosystème et de la décennie. Suite à un épisode d’exondation, la remise en eau, entrainant le relargage des nutriments, la reprise de l’activité biologique et des processus biogéochimique à des taux élevés, est qualifiée de « hot moment » et correspond à une période relativement courte (McClain et al. 2003). Le premier objectif a

été d’évaluer l’impact potentiel de ces événements à l’échelle décennale, en étudiant, à l’échelle de l’écosystème (Gutknecht et al. 2006), l’effet de la dynamique à long terme des exondations sur la teneur en nutriments de la masse d’eau des zones humides. Pour atteindre cet objectif, j'ai mesuré les changements temporels de la physico-chimie des eaux de zones humides péri- fluviales rangées sur un gradient de fréquence et d’intensité d’exondations croissantes sur une période de 15 ans.

A l’échelle de l’organisme et de quelques semaines, le deuxième objectif a été d’étudier l’aptitude des espèces d’angiospermes aquatiques à mettre en place un ajustement plastique face aux exondations, en conditions expérimentales et in situ. J’ai cherché à déterminer quels types d’ajustements morphologiques les espèces peuvent développer en réponse à l’exondation,

CONTEXTE SCIENTIFIQUE ET OBJECTIFS mais également quels facteurs peuvent influencer cette réponse, à savoir la position phylogénétique, la niche écologique (hydrophyte vs. amphiphyte) ou encore la forme de croissance des différentes espèces (rosette vs. caulescente).

Enfin, à l’échelle de la communauté et de la saison, le troisième et dernier objectif a été d’identifier in situ la réponse à court terme (sur une année) des communautés végétales à un

épisode d’exondation en fonction des caractéristiques sédimentaires de la zone humide, en mesurant la résistance et la résilience de ces communautés après retour des conditions aquatiques.

METHODOLOGIE

Dans cette partie j’ai pris le parti de ne pas présenter les methodologie utilisées dans les différents articles qui suivent, mais plutôt de me focaliser sur les choix méthodologiques préliminaires, à savoir la sélection des sites d’étude, et le fil directeur des études portant sur les ajustements plastiques en réponse aux exondations.

1. Sites d’études

Les suivis, les expérimentations, et les prélèvements de végétation ont été réalisés en majeure partie dans les zones humides de la basse vallée de l’Ain et du haut Rhône (Tableau 2).

L’échantillonnage des plantes pour la comparaison in situ des phénotypes inondés et exondés a

été réalisé conjointement dans ces zones humides fluviales et dans les étangs de la Dombes

(Tableau 2). Ces milieux humides sont proches géographiquement et situés au nord-est de Lyon, assurant ainsi qu’ils soient soumis au même climat.

METHODOLOGIE

Tableau 2. Sites d’études dans lesquels les suivis et prélèvements ont été réalisées.

Sites d’étude Coordonnées Suivi Echantillonnage des Echantillonnage Etude in situ de la réponse géographiques physico- plantes utilisées pour comparaison in situ des des communautés végétales chimique l’expérimentation en phénotypes inondés et à un épisode d’exondation laboratoire exondés ZHP Rhône, La Chaume 05°06’23’’E 45°49’35’’N × Etangs de la Dombes 05°06’01’’E 46°05’45’’N et 05°07’18’’E × 46°03’18’’N ZHP Ain, Albarine 05°15’40’’E, 45°58’21’’N × × ZHP Ain, Bellegarde 05°18’10’’E 46°00’39’’N × ZHP Ain, Brotteaux 05°12’31’’E 45°48’09’’N × ZHP Ain, Carronière 05°30’80’’E 46°02’93’’N × ZHP Ain, Creux de 05°13’52”E Fouchoux 45°50’08”N × × ZHP Ain, Gourdans 05°22’86’’E 45°82’20’’N, × ZHP Ain, Planet amont 05°24’19’’E 45°84’34’’N, × ZHP Ain, Petits Peupliers 05°16’25’’E 45°58’36’’N × × ZHP Ain, Port Galland 05°21’44’’E 45°81’51’’N × ZHP Ain, Ricotti 05°23’81’’E 45°82’83’’N, × × ZHP Ain, Sous Bresse 05°14’43”E 45°49’24”N × × ZHP Ain, Vers la Borne 05°17’13’’E 45°59’02’’N × ZHP Ain, Vers la Borne 05°27’71’’E Ouest 45°98’14’’N, × ZHP Ain, Villette 05°16’57”E 45°59’08”N × × × ×

Les étangs de la Dombes sont des étangs artificiels peu profonds, situés sur un plateau d’altitude

moyenne de 280 m à l’ouest de la basse vallée de l’Ain et au nord du haut Rhône après sa

confluence avec l’Ain (Arthaud et al. 2013 ; Wezel et al. 2013).

Les zones humides péri-fluviales sont d’anciens bras fluviaux abandonnés par le cours d’eau

en marge du chenal actif. Ces milieux sont créés et entretenus par la dynamique des fleuves

dans leur plaine alluviale. Elles ont un intérêt particulier à être étudiées du fait de leur formation

et de leur contextes hydrogéologiques variés qui conduisent à la co-occurrence de nombreux

METHODOLOGIE fonctionnements et habitats contrastés dans le même paysage fluvial, et par conséquent à une biodiversité particulièrement élevée et originale (Ward & Tockner 2001 ; Amoros & Bornette

2002 ; Ward et al. 2002). Elles sont diverses en termes de caractéristiques sédimentaires, dépendantes des phénomènes d’érosion-déposition et donc de la dynamique du fleuve

(notamment la force des crues; Bornette et al. 1998 ; Bornette et al. 2008). Ces zones humides ont des fonctionnements hydrologiques variés et peuvent être alimentées par différentes sources en eau (rivière par débordement, nappe d’accompagnement de la rivière et/ou nappe de versant ;

Bornette & Amoros 1991 ; Bornette et al. 1996) qui ont des caractéristiques physico-chimiques différentes.

La rivière d’Ain prend sa source dans le Jura et parcours 200 km avant de se jeter dans le Rhône.

Au niveau de la basse vallée de l’Ain (50 km avant le confluence), la rivière coule avec une pente assez faible et présente une morphologie active qui crée une diversité de milieux labellisés

Natura 2000 « Milieux alluviaux de la basse vallée de l’Ain ». Les zones humides péri-fluviales de la basse vallée de l’Ain sont des modèles d’étude adaptés à notre problématique car elles subissent de fortes modifications de leur régime hydrologique et connaissent des diminutions de hauteur d’eau fréquentes et plus longues du fait 1) de l’augmentation des besoins en eau, 2) des activités anthropiques (modification des usages) et 3) du changement climatique.

Besoins en eau

L’Homme exerce une pression croissante sur la ressource en eau via l’augmentation des prélèvements (agricole, industriels et d’eau potable) des eaux de surfaces et souterraines, diminuant ainsi la disponibilité en eau pour les zones humides. La basse vallée de l’Ain est globalement peu urbanisée, mais la proximité de Lyon crée une pression démographique grandissante sur le secteur avec comme conséquence des prélèvements croissants pour l’eau potable. L’activité dominante reste l’agriculture tournée vers les grandes cultures, et notamment

METHODOLOGIE la culture intensive de maïs irrigué. Le bassin compte également un pôle industriel bien développé (Le Parc Industriel de la Plaine de l’Ain).

Modification des usages

La construction de barrages (les barrages de Vouglans, de Coiselet, de Cize-Bolozon et d’Allement) et la constitution de plans d’eau le long du cours d’eau modifient son régime hydrologique et donc la disponibilité et l’alimentation en eau des zones humides péri-fluviales.

L’augmentation du couvert forestier, la construction de barrages hydroélectriques, la chenalisation et l’extraction de sédiments entrainent un fort déficit sédimentaire dans le lit principal de nombreuses rivières (Peiry et al. 1994) et par conséquent l’enfoncement du lit de la rivière (incision). L’incision conduit à l’enfoncement du lit de la rivière et l'abaissement du niveau d'eau par rapport à la plaine alluviale (Bravard et al. 1997 ; Amoros & Bornette 2002 ;

Buijse et al. 2002). En été, cette baisse atteint fréquemment un maximum en raison de pluies plus faibles et une demande en eau plus importante. L’incision entraine une diminution de la hauteur d’eau plus fréquente et plus longue dans les zones humides péri-fluviales du fait de la diminution de la connectivité latérale et verticale avec la rivière. L’alimentation en eau des zones humides se fait alors principalement par la nappe de versant (Amoros & Bornette 2002).

L’Ain a été soumis à un phénomène d’incision très marqué depuis le 19ème siècle, du fait notamment de la construction entre 1931 et 1968 des 5 barrages hydroélectriques qui retiennent les sédiments (Bravard et al. 1989 ; Marston et al. 1995 ; Bornette et al. 1996 ; Rollet et al.

2013).

Changement climatique

Le rapport Jouzel (Ouzeau et al. 2014) donne les tendances d’évolution du climat en France.

Selon les modèles et les scénarios, les résultats montrent une hausse des températures moyennes de 0,6 à 1,3°C et du nombre de jours de vagues de chaleur l’été (entre 0 et 5 jours sur l’ensemble du territoire et jusqu’à 10 jours dans le quart Sud-Est) à l’horizon 2021-2050 et de 0,9 à 5,3°C

METHODOLOGIE et 5 à 20 jours de vague de chaleur estivale (jusqu’à 40 jours dans le sud est) à l’horizon 2071-

2100 par rapport à la période de référence (1976-2005).

2. Plasticité phénotypique

2.1 Expérimentations contrôlées vs phénotypage in situ

L’étude en laboratoire de la mise en place d’une réponse plastique par les espèces aquatiques soumises à l’exondation permet de contrôler les conditions tel que le type de sédiment, la cinétique de mise en place de l’exondation, la luminosité et l’humidité entre autres. Cependant, en conditions contrôlées, certains stimuli accompagnant la baisse de hauteur d’eau in situ, tel que les modifications de lumière et de température, sont absents ou limités. De surcroit, les individus utilisés sont standardisés, ou peuvent être démunis de leur organes de réserve les empêchant de produire de nouvelles feuilles ou tiges. Troisièmement, en conditions expérimentales les espèces ne sont pas forcément dans une niche écologique optimale

(limitation de l’espace, de la lumière, de certains nutriments).

L’échantillonnage d’individus in situ permet de se libérer des contraintes logistiques liés à l’expérimentation en laboratoire sur un grand nombre d’espèces, en terme notamment d’espace, et ainsi de pouvoir étudier un nombre relativement important d’espèces. Néanmoins, cette méthodologie ne permet pas de contrôler les conditions expérimentales, en particulier en ce qui concerne la cinétique et la durée de l’exondation subie par les organismes.

2.2 Traits morpho-anatomiques

Les traits morpho-anatomiques mesurés sur les plantes sont résumés sur la Figure 6. Le choix des traits découle de leur signification fonctionnelle (Tableau 1). Les traits sont faciles et rapides à mesurer (soft traits, Hodgson et al. 1999 ; Weiher et al. 1999 ; Cornelissen et al. 2003) sur un grand nombre d’espèces et de conditions de croissance.

METHODOLOGIE

Traits morphologiques mesurés: - hauteur (h) - surface foliaire (A) - masse sèche et fraiche: racine (R), tige (T), feuilles (F) - longueur (l) et largeur (L) du limbe, longueur du pétiole (p) et périmètre du limbe (P)

Figure 6. Traits morpho-anatomiques mesurés pour étudier la plasticité phénotypique des plantes aquatiques à l’exondation.

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

RESULTATS

1. Dynamique à long terme du fonctionnement des écosystèmes.

Mélissa De Wilde, Sara Puijalon, Felix Vallier et Gudrun Bornette. Physico-chemical consequences of water-level decreases in wetlands (soumis à WETLANDS).

Problématique

L’objectif de cet article est d’évaluer dans quelle mesure les baisses des niveaux d’eau dans les zones humides péri-fluviale sont liées aux changement de leur caractéristiques physico- chimiques à long terme. Les zones humides péri-fluviales sont particulièrement complexes dans ce cadre car elles peuvent être alimentées par des eaux d’origines différentes.

Les hypothèses testées sont :

1) les baisses des niveaux d'eau n'ont pas d'effet sur la teneur en nutriments de l'eau, en

raison du renouvellement de l'eau élevé dans ces écosystèmes, qui peut lessiver les

nutriments libérés par les sédiments après la phase d’exondation.

2) les caractéristiques physico-chimiques de l'eau de surface des zones humides péri-

fluviale deviennent de plus en plus semblables à celles de la nappe de versant, en raison

de l'augmentation de l'alimentation en eau des zones humides par la nappe de versant.

Pour tester ces hypothèses nous avons comparé les caractéristiques physico-chimiques de l'eau de surface de 9 zones humides péri-fluviales situées dans la plaine alluviale de la rivière d'Ain

(France), qui ont connu des baisses des niveaux d’eau contrastées entre 1993 et 2011.

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Physico-chemical consequences of water-level decreases in wetlands

Mélissa De Wilde*, Sara Puijalon, Felix Vallier, and Gudrun Bornette

UMR 5023 «Laboratoire d’Ecologie des Hydrosystèmes Naturels et Anthropisés »; Université de Lyon; Université Lyon 1; CNRS; ENTPE; 43 boulevard du 11 novembre 1918, 69622 Villeurbanne Cedex, France; *e-mail: [email protected]; phone: +33 (0)4 72 43 12 54

Abstract

In the context of global change, many wetlands experience decreases in water levels and changes in their hydrological functioning that modify their physico-chemical characteristics.

Riverine wetlands are particularly complex in this framework, because they may be supplied by different water sources. The objective of this study was to assess the extent to which long- term water level decreases in riverine wetlands relate to changes in their physico-chemical characteristics. We tested the hypothesis that water level decreases have no effect on water nutrient contents, because of the high water renewal in these ecosystems, which may leach nutrients released from the sediment after the dewatered phase. We expected also that water physico-chemical characteristics of riverine wetlands should become increasingly similar to those of the hillslope groundwater, because of increasing water supplies of hillslope groundwater to wetlands. We compared the surface water physico-chemical characteristics of

9 riverine wetlands located in the floodplain of the Ain River (France), which experienced contrasting water level decreases between 1993 and 2011. Over the studied period, no increase in water nutrient contents occurred; but significant physico-chemical changes happened, which suggest either an increased connectivity with the hillslope groundwater or a lower rate of water renewal in the wetlands. In 3 wetlands, water physico-chemical changes suggested an increase in river water supplies, which may be related to the lateral migration of the river towards the wetlands. These functional changes are related to the magnitude of the water level decrease,

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes but may also relate to the hydrogeological and geomorphological contexts of the riverine wetlands.

Keywords Riverine wetlands, River incision, Water level decrease, Hydrological connectivity,

Physico-chemical characteristics

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Introduction

Continental wetlands are among the ecosystems that contribute the most to the species richness of Earth and contain 40% of the world’s biodiversity (RAMSAR 2002). These wetlands also perform many ecosystem services, such as water purification, carbon sequestration, flood control, and the regulation of nutrient cycles (Millennium Ecosystem Assessment 2005; Keddy et al. 2009). Temperate wetlands undergo frequent decreases in water level induced by several factors that often act in synergy (Brinson and Malvàrez 2002), including drainage, river incision, rainfall decrease, and groundwater abstraction for human needs (Bravard et al. 1997;

Acreman et al. 2000; IPCC 2001; van Diggelen et al. 2006). Globally, hydrological changes are a major threat to wetlands because these can potentially lead to functional alterations and ultimately to the disappearance of these ecosystems (Bunn and Arthington 2002, Millennium

Ecosystem Assessment 2005). Consequently, wetlands are currently recognized as the most threatened ecosystems in the context of global change (Coe 1998; Meyer et al. 1999;

Millennium Ecosystem Assessment 2005).

Many studies have investigated the effects of decreases in the water level on the biogeochemical processes in wetlands. At low water levels, an increase in the water ammonium content is frequently reported because of increasing ammonification due to the increasing temperature, and resulting hypoxia (Racchetti et al. 2011). During dewatering (complete absence of water above the sediment), the absence of water causes warmer and aerobic conditions in the sediment, which results in the promotion of microbial activities (Freeman et al. 1996) and increases in nitrification and P mineralization (Olila et al. 1997; Olde Venterink et al. 2002; Gómez et al. 2012). Immediately after rewetting, nutrients (nitrate or ammonium and phosphates) are often released from the sediment, which leads to high nutrient levels in the water column and eventually to eutrophication (Corstanje and Reddy 2004; Song et al. 2007).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

The effect of water level decrease may strongly differ among wetlands. Indeed, these ecosystems may have contrasting geomorphology, hydrogeology, and resulting successional dynamics that lead to contrasting sediment (in terms of nutrient, carbon content, and grain-size) and water characteristics and consequently geochemical responses to water level decrease

(Johnston et al. 2001; Thoms 2006). Among wetlands, riverine wetlands are particularly diverse. In fact, even in the same floodplain reach, these wetlands can be supplied by water originating from different sources, e.g. river surface water (when the river backflows and/or overflows in the wetland, Amoros and Bornette 2002), river seepage water (when river water infiltrates in coarse sediment and exfiltrates in the wetland, Bornette et al. 1994a), and shallow groundwater from any hillslope (Bornette and Amoros 1991; Bornette et al. 1996). The way these different water sources supply the wetlands also varies with time, and this variation depends on the river discharge and the elevation of the groundwater table (Bornette and Heiler

1994, Bornette et al. 1998), and modifies consequently the water characteristics in wetlands.

Temperate riverine wetlands experience a decrease in water levels that is frequently induced by river incision and/or groundwater abstraction for human needs. River incision leads to the degradation of the river bed and the lowering of the river water level relative to the floodplain (Bravard et al. 1997; Buijse et al. 2002). In summer, this decline frequently reaches a maximum because of lower rainfalls and greater water demands. In such situations, riverine wetlands experience more extensive and longer periods of low water levels and more frequent dewatering. In addition, these ecosystems may experience changes in the water physico- chemical characteristics due to modifications of connectivity with the different sources of water

(river surface water, river seepage water, and hillslope groundwater; Bornette and Heiler 1994;

Bravard et al. 1997). In riverine wetlands, changes in the water physico-chemical characteristics may consequently result from a complex combination of the responses to changes in

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes hydrological connectivity and to water level decreases, but such changes have never, to our knowledge, been investigated.

The objective of this study was thus to assess the impact of water level decreases on the physico-chemical characteristics of riverine wetlands at a decennial scale. We focused on rarely studied calcareous oligotrophic or slightly mesotrophic riverine wetlands. We hypothesized that water nutrient level should not be correlated to water level decreases or dewatering because groundwater supplies may leach nutrients (ammonium, nitrate, and phosphate) released from the sediment in the course of dewatering events. In accordance with the model developed by

Bravard et al. (1997), because of an expected increased drainage of the hillslope groundwater by the wetlands in the course of dewatering, we also expected 1) a positive correlation between the magnitude of water level decrease and proximity between the wetland and the hillslope groundwater water characteristics and 2) an negative correlation between the magnitude of water level decrease and proximity between the wetland and the river water characteristics.

To test these hypotheses, we compared 9 riverine wetlands located in the same floodplain that encountered contrasting water level decreases over 20 years. We first tested how changes in water physico-chemical characteristics correlated with water level decrease over time. Second, we determined whether the physico-chemical similarity between wetland water and hillslope groundwater increased with increasing water level decrease.

Materials and methods

Study area

The study sites were located in the lower part of the Ain River floodplain, which is a tributary of the Rhône River upstream of Lyon (France; Figure 1; Table 1). The Ain River is a relatively natural and freely meandering river, but its downstream reach is impacted by the Vouglans dam, built in 1968, which regulates maximum annual peak flows and led to sediment deficit and

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes consequently conspicuous progressive river bed incision (Bravard et al. 1989; Marston et al.

1995; Bornette et al. 1996; Rollet et al. 2013). The river bed incision led to the lowering of the river water level relative to the floodplain (Bravard et al. 1997; Amoros and Bornette 2002), and consequently affected the water level in wetlands. As the Ain is a piedmont river the riverine wetlands are cut-off river channels that were naturally created by river dynamics. Some of them are still permanently connected to the river at their downstream end, and some are completely disconnected at both ends. All the wetlands have been recently cut-off, they are all highly supplied by groundwater (either seepage or hillslope groundwater, in different proportions according to the season and the river discharge; Bornette et al. 1998). During floods, which usually last a few days, the river can overflow the wetlands. All these riverine wetlands experience a decrease in water levels over the last years, induced by river incision, and groundwater abstraction for human needs.

Fig. 1 Locations of the wetland, groundwater, and the Ain River water sampling stations.

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Table 1 Summer water level decrease between 1993 and 2011 (%),the water sampling years, the total number of sampling dates, and the groundwater reference sampling station for each riverine wetland and the Ain River.

Site Site designation Summer water level Water sampling years Total number of Groundwater reference decrease (%) sampling dates sampling site (Site designation) Port PGAwet 1.8 1993, 1994,1995, 40 Leyment (LEYgro) Galland 1996, 2003, 2007, 2009, 2010, 2011 Gourdans GOUwet 15 1994, 1995, 1996, 36 Dombes (DOMBgro) 2003, 2007, 2009, 2010, 2011 Ricotti RICwet 51.1 1994, 1995, 1996, 36 Leyment (LEYgro) 2003, 2007, 2009, 2010, 2011 Sous SBRwet 55.7 1993, 1994, 1995, 55 Leyment (LEYgro) Bresse 1996, 2003, 2005, 2006, 2007, 2009, 2010, 2011 Villette VILMwet 66.4 1996, 2002, 2003, 37 Bugey (BUGgro) Amont 2007, 2009, 2010, 2011 Creux de CFOwet 69.5 1994, 1995, 1996, 46 Dombes (DOMBgro) Fouchoux 2002, 2003, 2007, 2009, 2010, 2011 Planet PLAMwet 100 1993, 1994, 1995, 32 Leyment (LEYgro) Amont 1996, 2003, 2007, 2009, 2010 Petits PEUwet 100 1993, 1994, 1995, 35 Bugey (BUGgro) Peupliers 1996, 2003, 2006, 2007, 2009, 2010, 2011 Vers La BORwet 100 1993, 1994, 1995, 58 Bugey (BUGgro) Borne 1996, 2002, 2003, 2005, 2006, 2007, 2009, 2010, 2011 Ain River AINriv 1993, 1994, 1995, 34 1996, 2003, 2009, 2010, 2011

Among the 9 wetlands studied (Table1; Figure1), three of them, “Villette amont”

(VILMwet), “Vers la Borne” (BORwet), and “Petits Peupliers” (PEUwet), are located on the upstream left bank of the river. In this area, the shallow hillslope groundwater originates from the Bugey karstic mountains that is contaminated by nitrates by agriculture activity in the floodplain (BUGgro) (Fig. 1, Bornette et al. 1998, unpublished report). A second set of wetlands, which includes “Planet amont” (PLAMwet), “Ricotti” (RICwet), “Sous Bresse”

(SBRwet), and “Port Galland” (PGAwet), is located downstream along the left bank of the river. In this area, the shallow hillslope groundwater (LEYgro) originates from the karstic

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes mountains close to the Leyment village, also contaminated by nitrates from agriculture. The last two wetlands, “Creux de Fouchoux” (CFOwet) and “Gourdans” (GOUwet), are located in the downstream right bank of the river. The hillslope groundwater in the Dombes area concerned by this study is derived from the moraine Dombes plateau, and the geological substratum is dominated by clay and coarse alluviums (DOMBgro).

PEUwet (length = 250 m; mean width = 6.5 m) and RICwet (length = 375 m; mean width = 13 m) are fairly uniform in terms of water depth with a low summer depth (maximum summer depth = 0.8 m). GOUwet (length = 375 m; mean width = 14 m) is morphologically heterogeneous with an upstream basin that is 2 m in depth followed by a uniform linear shallow channel (depth ≤ 0.5 m). BORwet (length = 575 m; mean width = 23 m) and PLAMwet (length

= 200 m, mean width = 22 m) consist of two rather homogeneous channels that are less than

1.5 m in depth. CFOwet (length = 600 m; mean width = 11 m), VILMwet (length = 650 m; mean width = 19.5 m), SBRwet (length = 700 m; mean width = 15 m), and PGAwet (length =

850 m; mean width = 42 m) are composed of alternating riffles and pools (0.8 to 3 m in depth).

Apart from flood events, the water in these wetlands flows toward the river through surface or seepage runoff. The flow may stop in summer, when the riffles and the shallowest parts of the channels become dewatered.

The riverine wetland substrate at the date of the cut-off consisted of pebbles like river bed load. Presently, the substrate is highly heterogeneous and consists of a mosaic of several sediment types. The pebbles either stay uncovered by fine sediment because of high groundwater drainage, riffle occurrence, or floods that scour fine sediment (Bornette et al.

1994b) or are covered by fine sediments in the deepest or more preserved situations. Silt and clay dominate in the areas subjected to river water inputs with low energy (mainly downstream confluences). Shallow layers of organic matter (a few cm) dominate in the deepest areas and in the areas where the river energy during floods is low. These wetlands comprise plant

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes communities composed mainly of submerged aquatic vegetation and emergent macrophytes along the banks and in the shallowest parts of the channels. Among the more frequent aquatic macrophytes are mesotraphent species, such as Berula erecta, Mentha aquatica, Sparganium emersum, Potamogeton natans, and Chara major (see Bornette et al. 2001 for more details).

Plant communities are stable over the last decade, and have little control over the physico- chemical changes in these systems, which are mostly governed by the connectivity with river and groundwater fluxes (Bornette and Amoros 1991, Bornette et al. 1998).

Water level decreases in riverine wetlands

The summer maximum water depth of the 9 riverine wetlands ranged from 0.8 to 3 m in 1993 and from 0 to 3 m in 2011. This great variability of water depth between and within riverine wetlands, due to their complex geomorphology, makes necessary to express water level changes as a mean percentage decrease. For each riverine wetland, the water depth was collected during the summer (low water levels) using a gauge at several points (every 25 to 50 meters so 7 to 33 points depending on the wetland) distributed along the upstream-downstream continuum of the wetland channel at several years (from 3 to 11 years, distributed between 1993 to 2011, depending on the wetland). For each sampling date, the measurements were averaged to create a single value of the summer water level. The magnitude of the change in the mean summer water level between 1993 and 2011 was assessed by the slope of the measurements performed at the different dates. The magnitude of the water level decrease was then used to calculate the relative water level decrease (%) in the wetlands. The average percentage decrease in the summer water level between 1993 and 2011 differed among the wetlands and ranged from less than 2% for PGAwet to 100% for PLAMwet, PEUwet, and BORwet (Table 1). Two wetlands exhibited low summer water level decreases over the period (1.8% and 15% for PGAwet and

GOUwet, respectively), four exhibited an intermediate water level decrease (from 51.1% to

69.5%: RICwet, SBRwet, VILMwet, and CFOwet), and three wetlands exhibited a high

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes summer water decrease and became temporary wetlands over the period (100% water level decrease: PLAMwet, PEUwet and BORwet).

Physico-chemical data

The water physico-chemical characteristics of the wetlands and of the Ain River have been surveyed monthly or bimonthly from April to October from 1993 to 2011 to minimize seasonal variations (at least 7 years of sampling with 32 to 58 sampling dates depending on the wetland;

Table 1). To minimize within-day chemical variations, water samples were collected at all dates in the same order for a given sampling event. Water samples were collected near the water surface in each wetland and in the river (Figure 1), filtered with glass microfiber filters with a mesh size of 1.2 μm, brought to the laboratory in an icebox, stored at 4°C, and analyzed within

48 h after collection. The N-nitrate (mg L-1), N-ammonium (mg L-1), and P-phosphate (mg L-

1) concentrations, which are indicative of the nutrient levels, were measured by colorimetry using standard HACH procedures (HACH Company, Loveland, CO, USA) for the 1993-2002 period and through a colorimetric method with a spectrometer (Easychem Plus SysteaTM) for the 2002-2011 period. For the HACH procedures, the cadmium reduction method was used to quantify the N-nitrate concentration (lower limit of detection = 0.01 mg L-1), the salicylate method was used to quantify the N-ammonium concentration (lower limit of detection = 0.02 mg L-1), and the ascorbic acid method was used to quantify the P-phosphate concentration

(lower limit of detection = 0.006 mg L-1). For the colorimetric method with a spectrometer, the cadmium reduction method was used to quantify the N-nitrate concentration (lower limit of detection = 0.03 mg L-1), the blue indophenol method (analogue of the salicylate method) was used to quantify the N-ammonium concentration (lower limit of detection = 0.016 mg L-1), and the ascorbic acid method was used to quantify the P-phosphate concentration (lower limit of detection = 0.006 mg L-1).The two types of protocols for nutrient dosing were comparable, and the detection limits allowed the detection of low nutrient values. The water P-phosphate and N-

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes ammonium contents of the river and river seepage are high, and the Bugey hillslope groundwater is characterized by high N-nitrate levels when it reaches the river floodplain

(Bornette et al. 1996; Bornette et al. 1998). The temperature (°C) and conductivity (μS cm-1) were measured in situ using a WTW conductometer. A cool and stable wetland water temperature throughout the seasons is indicative of a high contribution of hillslope groundwater

(Bornette and Amoros 1991). The hillslope groundwater originating from the Bugey mountains is characterized by high conductivity (Bornette et al. 1996, Bornette et al. 1998). The oxygen level (mg L-1) and pH were measured in situ using a WTW oximeter and a pH meter, respectively. The oxygen content increases when the hillslope groundwater supplies are high, which results in the maintenance of cool oxygenated conditions in the wetland (hillslope groundwater has a rather high oxygen content, see Table 2). The pH increases when river water supplies the wetland (seepage or surface river water; Bornette and Amoros 1991; Bornette et al. 1996; Bornette et al. 1998).

The physico-chemical characteristics of the Bugey, Leyment, and Dombes groundwater were documented using the database Access to Data on Groundwater

(http://www.ades.eaufrance.fr).

Data analysis

The significance of the temporal changes in the water physico-chemical characteristics throughout the studied period (from 1993 to 2011) was first tested using all the chemical data available for a given wetland using linear regressions. The data were log transformed when necessary to better meet the assumption of normality. Given the large number of tests that was performed, we applied a sequential Bonferroni-Holm correction.

Second, for each parameter, an ANCOVA was performed on the set of wetlands that presented significant changes over time to test whether the slope of the regression (i.e., the rate of change) was similar among the wetlands for this parameter. Each ANCOVA was performed

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes using the physico-chemical parameter as the dependent variable, the time (i.e. the whole study period) as a covariate, and site (factor), and their interaction as the main effects.

Third, to assess whether the wetland physico-chemical characteristics became closer to those of the river or hillslope groundwater through time in the framework of possible concomitant changes in the physico-chemical characteristics of these two water masses, it was necessary to cope with the heterogeneity of temporal data, and compare data for similar sampling dates and sampling periods, for avoiding any differences that may be due to year-to- year stochastic variation of chemical data. For this purpose, we selected two groups of data with identical sampling dates in 1995-1996 and 2009-2010 periods (4 dates per year so 16 dates for each site). The data were analyzed using a between-class normalized Principal Component

Analysis (nPCA), in which each sampled site at each period (1995-1996 or 2009-2010) was considered a class, in order to consider the average chemical change between the two periods

(see Bornette et al. 1994c and Puijalon and Bornette 2004 for more details). Then, for expressing these changes relative to the river or groundwater characteristics (which may have change also during the same period, Table 1) we subtracted at each period the factorial coordinates of the river or groundwater to factorial coordinates of wetlands. We then tested whether the changes in the wetland water characteristics between the two periods (1995-1996 and 2009-2010) (factorial distance provided by the between-class nPCA, corrected by the river or groundwater trajectory) were correlated to the percentage of the decrease in the water level using a Spearman correlation test.

All of the statistical analyses were performed using the R.2.10.1 software (R-

Development-Core-Team, 2009).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Results

Changes in the water physico-chemical characteristics over time

Some of the physico-chemical parameters varied significantly over time, although these changes were not always in the same direction (Table 2; Table 3). The water temperature increased significantly in 3 of the riverine wetlands that showed intermediate to high water level decreases (BORwet, PLAMwet, and VILMwet; Table 3). The water oxygen content decreased significantly in GOUwet, PGAwet, and RICwet, which experienced low to moderate water level decreases, and increased in VILMwet (intermediate water level decrease; Table 3). The water pH decreased in SBRwet and increased in VILMwet, both of which exhibited intermediate water level decreases (Table 3). The conductivity increased significantly in 4 riverine wetlands with low to intermediate water level decreases (GOUwet, PGAwet, RICwet, and SBRwet) and decreased in VILMwet (Table 3). The N-nitrate contents decreased in

BORwet, PLAMwet (both were fully dewatered at the end of the summer period), and PGAwet

(had the lowest water level decrease over the same period), and the N-ammonium contents decreased in BORwet, PLAMwet (both were fully dewatered at the end of the summer period),

SBRwet (the summer water level decreased intermediately over the period), and the Ain River.

No significant variation in the P-phosphate concentrations was observed (Table 3).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Table 2 Physico-chemical characteristics (mean ± sd) of the river, wetlands, and groundwater stations for the 1995-1996 and 2009-2010 periods.

Site Period Temperature Oxygen pH Conductivity [N-NO3] [N-NH3] [P-PO4] designation (°C) (mg L-1) (μS cm-1) (mg L-1) (mg L-1) (mg L-1) PGAwet 1995-1996 17.6±2.7 9.2±2.9 7.6±0.2 419±34 1.25±0.4 0.04±0.01 0.008±0.01 GOUwet 1995-1996 15.7±1.3 8.03±1.6 7.6±0.2 449±44 0.7±0.2 0.05±0.05 0.005±0.007 RICwet 1995-1996 14.2±3.0 6.5±0.9 7.6±0.1 438±59 0.8±0.5 0.05±0.05 0.004±0.004 SBRwet 1995-1996 15.8±2.3 6.6±0.9 7.5±0.1 558±49 5.5±1.3 0.05±0.03 0.004±0.005 VILMwet 1995-1996 11.5±0.8 3.6±0.9 7.3±0.03 488±6 2.1±0.3 0.03±0.01 0.002±0 CFOwet 1995-1996 16.4±2.4 8.2±1.9 7.7±0.1 363±38 0.6±0.2 0.03±0.03 0.005±0.005 PLAMwet 1995-1996 12±1.7 6.7±1.3 7.6±0.2 512±56 3.3±0.5 0.05±0.02 0.01±0.01 PEUwet 1995-1996 14.6±2.4 4.1±1.6 7.6±0.1 508±39 2.3±2.2 0.08±0.07 0.004±0.005 BORwet 1995-1996 13.7±2.4 8.5±1.1 7.5±0.08 555±25 5.1±0.7 0.03±0.01 0.006±0.006 BUGgro 1995-1996 13.7±0.8 7.5±0.7 7.4±0.07 521±18 4.2±1.9 0±0 0±0 DOMBgro 1995-1996 13.0±0.9 9.5±0.9 7.3±0.1 647±22 9.3±0.7 0±0 0±0 LEYgro 1995-1996 12.5±2.2 9.05±0.9 7.3±0.2 507±38 4±1.9 0±0 0±0 AINriv 1995-1996 15.2±2.1 10.9±1.1 8.3±0.1 365±24 0.9±0.2 0.05±0.03 0.006±0.01 PGAwet 2009-2010 17.9±5.1 7.1±1.8 7.9±0.1 442±15 0.7±0.4 0.03±0.02 0.005±0.004 GOUwet 2009-2010 15.5±3.9 5.9±1.7 7.6±0.1 476±25 0.6±0.3 0.03±0.02 0.006±0.004 RICwet 2009-2010 14.2±3.7 4.3±1.9 7.6±0.1 461±21 0.8±0.7 0.04±0.03 0.005±0.004 SBRwet 2009-2010 14.4±2.1 5.7±1.3 7.2±0.1 581±19 6.8±2.5 0.02±0.02 0.005±0.003 VILMwet 2009-2010 13.7±1.2 5.8±1.1 7.6±0.1 454±15 1.8±0.5 0.02±0.01 0.005±0.004 CFOwet 2009-2010 17.2±4.7 7.9±2.5 7.9±0.3 360±27 0.5±0.4 0.04±0.04 0.005±0.003 PLAMwet 2009-2010 14.7±1.3 7.0±2.6 7.7±0.3 482±45 1.6±1.1 0.01±0.002 0.006±0.003 PEUwet 2009-2010 13.5±3.1 4.8±1.9 7.8±0.3 489±22 1.9±1.4 0.06±0.04 0.004±0.002 BORwet 2009-2010 19.4±4.6 9.1±2.6 7.8±0.4 520±65 2.4±2.1 0.02±0.01 0.005±0.004 BUGgro 2009-2010 13.2±0.9 7.3±1.1 7.3±0.2 544±67 2.9±0.5 0±0 0±0 DOMBgro 2009-2010 13.3±2.2 7.9±2.1 7.3±0.1 637±33 8.3±0.7 0±0 0±0 LEYgro 2009-2010 11.9±0.8 7.9±1.3 7.4±0.3 488±18 2.8±1.2 0±0 0±0 AINriv 2009-2010 15±4.3 10.2±1.4 8.2±0.1 362±26 0.8±0.1 0.01±0.01 0.005±0.004

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Table 3 Summer water level decrease (%) for each riverine wetland and temporal changes in the physico-chemical characteristics of the water from the Ain River and the wetlands analyzed by linear regression.

Physico-chemical parameters - - - Site Summer water Temperature Oxygen pH Conductivity [N-N03 ] [N-NH3 ] [P-PO43 ] designation level decrease (%) (°C) (mg L-1) (μS cm-1) (mg L-1) (mg L-1) (mg L-1) 1 1 1 1 1 1 1 PGAwet 1.8 F 38=1.22 ns F 38=7.82**(-) F 37=2.95 ns F 37=6.40*(+) F 37=4.20*(-) F 37=3.48 ns F 36=0.051 ns

1 1 1 1 1 1 1 GOUwet 15 F 34=1.79 ns F 31=8.87**(-) F 32=0.840 ns F 33=5.95*(+) F 33=0.852 ns F 33=3.88 ns F 32=0.818 ns

1 1 1 1 1 1 1 RICwet 51.1 F 33=0.0837 ns F 32=7.86**(-) F 33=0.136 ns F 33=6.52*(+) F 33=0.181 ns F 33=0.280 ns F 33=0.359 ns

1 1 1 1 1 1 1 SBRwet 55.7 F 52=2.55 ns F 51=3.13 ns F 51=5.04*(-) F 51=6.75*(+) F 52=0.0146 ns F 52=4.73*(-) F 50=1.08 ns

1 1 1 1 1 1 1 VILMwet 66.4 F 34=14.4***(+) F 32=4.51*(+) F 34=15.2***(+) F 32=4.53*(-) F 34=0.913 ns F 34=1.69 ns F 34=1.24 ns

1 1 1 1 1 1 1 CFOwet 69.5 F 43=0.990 ns F 44=0.328 ns F 43=0.93 ns F 42=0.0262 ns F 44=1.14 ns F 43=0.207 ns F 43=0.107 ns

1 1 1 1 1 1 1 PLAMwet 100 F 29=8.14**(+) F 29=0.534ns F 30=0.00295 ns F 29=1.53 ns F 30=9.59**(-) F 27=5.76*(-) F 27=2.88 ns

1 1 1 1 1 1 1 PEUwet 100 F 33=0.637 ns F 33=2.08 ns F 33=0.729 ns F 32=0.403 ns F 32=0.251 ns F 32=2.02 ns F 30=1.87 ns

1 1 1 1 1 1 1 BORwet 100 F 56=8.98**(+) F 55=1.55 ns § F 55=0.0227 ns F 55=1.46 ns F 55=7.95** (-) F 50=5.99*(-) F 54=0.309 ns

1 1 1 1 1 1 1 AINriv F 31=0.119 ns F 31=2.92 ns F 32=0.0152 ns F 32=0.000218 ns F 31=0.0497 ns § F 31=11.1 **(-) F 31=0.000977 ns

The Fdf values, significance levels, and sense of variation (if significant) of each parameter over time are presented. Significance levels: ***P < 0.001; **P < 0.01; *P < 0.05; ns, not significant; § = data were log-transformed. After correction using the Bonferroni-Holm sequential method, none of the tests were significant.

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

The first two axes of the between-class nPCA explained 77% of the variance of the dataset. The conductivity (EC) and N-nitrate concentration were positively correlated with the first axis. The pH and the temperature (T) were negatively correlated with the first axis (Fig.

2a). The N-ammonium and P-phosphate concentrations were also negatively correlated, although to a lesser extent, with the first axis (Fig. 2a). The oxygen and N-ammonium contents were negatively and positively correlated, respectively, with the second axis (Fig. 2a).

Two groups of wetlands were discriminated according to their trajectory corrected by the river trajectory (Fig. 2b). For the first group (PEUwet, VILMwet, PLAMwet, BORwet, and

CFOwet), the wetland water characteristics changed toward the origin of the F1 and F2 axes, i.e., increased in similarity with the river physico-chemical characteristics. This trend mainly corresponded to a decrease in the conductivity and N-nitrate content in the wetlands over time.

A second set of wetlands (PGAwet, RICwet, and GOUwet) changed less over time and displayed a trajectory perpendicular to the first set, which indicates a lack of increasing similarity with the river over time. The last wetland, SBRwet, did not change over time.

When the factorial changes in the wetland water characteristics were corrected by the factorial changes of the groundwater sampled in their surroundings (Fig. 2c-e), 3 wetlands

(BORwet, PEUwet, and VILMwet; Fig. 2c) changed toward negative values on the first axis over time, which indicates an increasing contrast between the wetland water characteristics and the groundwater characteristics in their surroundings. Two other wetlands (CFOwet and

GOUwet) changed from highly positive values to lower values on the second axis (Fig. 2d).

The water characteristics of CFOwet changed to a greater extent than those of GOUwet on the second axis and did not change on the first axis, whereas the water characteristics of GOUwet moved toward the origin of the factorial axes, i.e., closer to the groundwater characteristics, over time. PGAwet, PLAMwet, and SBRwet also moved toward lower values on the second axis, and only RICwet factorial position changed positively on the first axis (Fig. 2e).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Consequently, the similarity between the PLAMwet and PGAwet water characteristics with the characteristics of their groundwater reference increased over time, whereas both SBRwet and

RICwet, which are located close to each other in the floodplain, did not increase in similarity with the groundwater over time.

Fig. 2 Between-class normalized PCA of physico-chemical changes in the wetland, groundwater, and Ain River water between 1995-1996 and 2009-2010. In this analysis, one wetland at one period was considered a class. a: Correlation circle between the parameters and the factorial axes 1 and 2 (N.NH3: N-ammonium; N.NO3: N-nitrate; P.PO4: P-phosphate; Cd: conductivity; T: temperature; O2: oxygen). b: Factorial changes in the wetlands corrected by the river trajectory. The dashed arrows correspond to the sites that exhibit a trajectory that goes to the river characteristics. c to e: Factorial changes in the wetlands corrected by their groundwater reference trajectory (c: Bugey groundwater; d: Dombes groundwater; e: Leyment groundwater). The numbers on the axes of b-e are the factorial coordinates. The length of the arrows represents the amplitude of the changes between the two periods. The origin of the diagrams represents the river (b) and the groundwater source (c to e).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Changes in the physico-chemical parameters and water level decreases

For each physico-chemical parameter, ANCOVAs were used to test whether the rate of temporal change from 1993 to 2011 was similar among those wetlands that showed a similar change direction over time. For the temperature, oxygen, conductivity, and nitrate contents, the effects of time and site were significant, and the interaction term (time x site) was not significant

(Table 4). This finding indicates different ranges of values but a similar rate of change over time between the wetlands for each parameter (Table 4). For the N-ammonium content, neither the wetland nor the interaction (time x site) effects were significant, which indicates that all of the wetlands have the same N-ammonium content and changed in the same way for this parameter (Table 4).

Table 4 Effects of time, wetland, and their interaction on the physico-chemical parameters that varied significantly over time, tested by ANCOVA. Parameter (sites) Time Site Time u Site 1 2 2 Temperature F 119=16.01*** F 119=12.60*** F 119=0.3435 ns (BORwet, PLAMwet, VILMwet) 1 2 2 Oxygen content F 101=26.74*** F 101=22.99*** F 101=0.3633 ns (GOUwet, PGAwet, RICwet) 1 2 2 Conductivity F 154=32.68*** F 154=133.7*** F 154=0.1686 ns (GOUwet, PGAwet, RICwet, SBRwet) 1 1 1 Nitrate nitrogen content F 122=10.48** F 122=54.21*** F 122=2.578 ns (BORwet, PGAwet, PLAMwet) 1 1 1 Ammonium nitrogen content F 129=19.90*** F 129=1.375 ns F 129=0.6334 ns (BORwet, PLAMwet, SBRwet) The Fdf values and the significance levels of the effects (time as a co-variable, main effect of the site, and their interaction) are presented. Significance levels: ***P < 0.001; **P < 0.01; *P < 0.05; ns, not significant.

The distances between the wetland factorial coordinates in 1995-1996 and in 2009-2010 obtained from the between-class PCA corrected by the river changes over the same period were positively correlated with the wetland water level decreases (S = 34.56, rho = 0.71, p = 0.03;

Figure 3a). The factorial distances corrected by the groundwater factorial changes were also positively correlated with the water level decreases (S = 36.6, rho = 0.7, p = 0.04; Figure 3b).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Fig. 3 Relationship between the summer water level decreases and the factorial distances adjusted by a) the river trajectory (rho = 0.71; p = 0.03) and b) the groundwater trajectory (rho = 0.7; p = 0.04) between 1995-1996 and 2009-2010.

Discussion

As hypothesized, we did not observe any increase in the water nutrient content in the wetlands, which may have resulted from nutrient solubilization during the course of dewatering-rewetting events (no increase in P-phosphate, N-ammonium, and N-nitrate water contents over time). For several parameters, we demonstrated directional changes that suggest modifications in the wetland connectivity with the river and groundwater.

Significant and similar conductivity increases and oxygen concentration decreases

(except for SBRwet) were observed in all of the wetlands that experienced a low to moderate summer water level decrease (i.e., PGAwet, GOUwet, RICwet, and SBRwet; from 1.8 to

55.7%). An increase in the water conductivity may result from an increasing connectivity with the calcareous hillslope groundwater and/or a decreasing connectivity with the river over time.

The decreasing water oxygen content may result from a lower water renewal rate in these wetlands, which may favor stagnation and subsequent hypoxia in pools remained with water.

In two of the three wetlands that experienced the highest summer water level decrease over time (PEUwet, BORwet, and PLAMwet; 100%), the water temperature increased

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes significantly (PLAMwet and BORwet), suggesting that, the groundwater influence decreased over time in these wetlands.

One of the wetlands that was intermediately impacted by the summer water level decrease (VILMwet) changed significantly, although in a different way, compared with the other wetlands that experienced a similar water level decrease (RICwet, SBRwet, and CFOwet).

Physico-chemical changes observed suggests an increase (rather than a decrease) in the proportion of river water supplying the wetland (because the changes in temperature, pH, and conductivity indicates an increasing similarity between the wetland water characteristics with those of the river water). Such an increase may be related to the river migration toward the wetland during the study period. It means that for a given water level decrease, the connectivity with the hillslope groundwater may not increase if the mobility of the river guaranties connections between river and the wetland through seepage.

Dewatering-related eutrophication was not observed in our study. This may result from the exportation of nutrients by groundwater fluxes. Eutrophication may also be impeded by biological and geochemical processes (Cavanaugh et al. 2006). Indeed, even if we did not directly measure these processes, we observed N-nitrate and N-ammonium content decreases in several wetlands (both the N-nitrate and the N-ammonium contents decreased in PLAMwet and BORwet, the N-nitrate content decreased in PGAwet, and the N-ammonium content decreased in SBRwet), which may be related to either a higher consumption rate by primary producers or to increasing nitrogen loss as N2 gas due to alternating dry and inundated phases

(Baldwin and Mitchell 2000; James et al. 2004). It is difficult to exclude the possibility that this type of nutrient decrease may be related to vegetation changes over time. However, the dominant species remained rather unchanged during the study period (with a low progression of helophytes because of rather high water levels in winter). Furthermore, the trophic status of the plant communities either remained mesotrophic or became more oligotrophic (Arthaud and

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Bornette 2012, unpublished report). Consequently, the observed nitrogen decrease may be related to microbial activity.

An increase in the wetland connectivity with calcium- and nitrate-rich hillslope groundwater may also prevent phosphate-induced eutrophication (Smolders et al. 2006).

Indeed, dewatering favors iron oxidation and leads to the formation of Fe-PO4 and Al-PO4 complexes in the sediment (Baldwin and Mitchell 2000). Sediment rewetting with NO3-rich groundwater maintains Fe-PO4 complexes because nitrate limits Fe reduction and phosphate release to the water column (Lucassen et al. 2005). In calcareous wetlands, sediment calcium may also contribute to the sorption of P in the sediment by Ca-P binding that is considered unavailable (P co-precipitates with CaCO3; De Groot and Van Wijck 1993).

The magnitude of the water physico-chemical changes observed in the wetlands between 1995-1996 and 2009-2010 was correlated with the magnitude of the water level decrease, which indicates that the water level decrease is related to modifications in the wetland functioning. However, the direction of the functional changes differed greatly between the wetlands, and, contrary to what we expected, we did not observe a systematic increase in the physico-chemical similarity between the wetlands that exhibited a high water level decrease and the hillslope groundwater. In fact, although the summer water levels in BORwet, PEUwet,

PLAMwet, CFOwet, and VILMwet strongly decreased over time, their water physico-chemical characteristics was increasingly similar to that of the river. During the course of river incision, the river may increasingly drain the hillslope groundwater, and one should observe an increasing similarity between the river and the hillslope groundwater. In the present situation, we observed a decreasing similarity between the characteristics of BORwet, PEUwet, and

VILMwet and the hillslope groundwater and an increasing similarity with the river, which suggests that river migration may have led to an increasing connectivity between the wetlands and the river.

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

Most studies that focused on the effects of water level decrease on wetland physico- chemical characteristics used an experimental approach and studied cycles of dewatering and rewetting over a short time scale without considering the influence of different water sources

(Olde Venterink et al. 2002; Corstanje and Reddy 2004; Song et al. 2007). Such short-term responses may poorly reflect the long-term functioning of wetlands if some hydrogeological processes interact with biochemical processes and limit or impede wetland eutrophication. In fact, many wetlands are located in a floodplain context and are influenced by groundwater

(Brinson 1993; Winter 1999). In such situations, alkaline groundwater fluxes may prevent eutrophication through either leaching or sediment-binding of nutrients. Moreover, spate floods in piedmont rivers may contribute to the washing out of fine sediment and nutrients from wetlands (Bornette et al. 1994b, c, in prep.). However, in riverine wetlands located far from the river, the lack of scouring floods may lead to clogging, which would limit the vertical connectivity with groundwater and lead to ecosystems that are ruled by internal functional processes (Tockner et al. 1999; Cabezas et al. 2008; Cabezas et al. 2009).

Conclusion

In the present study, we demonstrated that the magnitude of functional changes in riverine wetlands is correlated with the magnitude of the summer water level decrease over time. To improve the predictions of the responses of wetlands to dewatering induced by global change, one must consider the geomorphological and geological context of wetlands. Indeed, these parameters, even if somewhat difficult to quantify precisely, may strongly affect the geochemical responses to dewatering and the ecosystem resilience after the disturbing event.

In the particular case of riverine wetlands, our results did not demonstrate any increase in the nutrient levels of oligo-mesotrophic wetlands in response to dewatering, if these wetlands are connected to groundwater through coarse substrate. For eutrophic and/or wetlands where the

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes substrate is clogged by fine sediments, one may expect a higher nutrient release during the course of dewatering. Substrate characteristics may consequently also strongly influence the responses of wetlands to dewatering because they modify the nutrient-binding capacity of the substrate and nutrient leaching through groundwater or surface water fluxes.

Acknowledgments

This work was performed under the aegis of the LTER “Zone Atelier Bassin du Rhône” and was funded by the Wetchange Program (ANR-09-CEP-006-01) of the French National

Research Agency (Agence Nationale de la Recherche-ANR).

RESULTATS Dynamique à long terme du fonctionnement des écosystèmes

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RESULTATS Mise en place des réponses plastiques

2. Ajustement plastique des espèces

2.1 Mise en place des réponses plastiques.

Mélissa De Wilde, Nadia Sebei, Sara Puijalon, et Gudrun Bornette. Responses of macrophytes to dewatering: effects of phylogeny and phenotypic plasticity on species performance (sous presse dans Evolutionary Ecology).

Problématique

L’objectif de cet article est de 1) mesurer l’effet de l’exondation sur la performance des plantes aquatiques, 2) déterminer de quelle manière la forme de croissance et la position phylogénétique affecte la performance et 3) relier la performance à la plasticité.

Les hypothèses testées dans ce travail sont :

1) les espèces appartenant à des groupes taxonomiques ayant colonisés l’habitat aquatique

tôt dans l’histoire évolutive des Angiospermes ont une survie et une croissance plus

faible en condition exondée comparé aux espèces appartenant à des groupes

taxonomiques ayant colonisés le milieu aquatique plus récemment,

2) pour les espèces proches phylogénétiquement, celles ayant une forme de croissance en

rosette ont une survie supérieure et un taux de croissance moins impacté en condition

exondée comparé aux espèces caulescentes,

3) en terme de réponse plastique, les plantes aquatiques en condition exondée comparée

aux planes aquatiques en conditions submergées, devraient avoir une teneur en matière

sèche des organes aériens (tiges, feuilles) plus élevée, une surface foliaire plus faible,

résultant en une surface foliaire spécifique plus faible et une durée de vie des feuilles

plus élevée et ces changements devraient permettre aux plantes de maintenir leur

performance en condition exondée.

Pour tester ces hypothèses, la réponse à l’exondation a été mesurée en condition expérimentale en mesurant la survie, la croissance et un ensemble de traits décrivant la morphologie des

RESULTATS Mise en place des réponses plastiques plantes et l’économie en ressource des feuilles de 8 espèces de plantes aquatiques ayant des positions phylogénétiques et des formes de croissances contrastées.

RESULTATS Mise en place des réponses plastiques

Responses of macrophytes to dewatering: effects of phylogeny and phenotypic plasticity on species performance

Mélissa De Wilde*, Nadia Sebei, Sara Puijalon, Gudrun Bornette

UMR CNRS 5023 « Laboratoire d’Ecologie des Hydrosystèmes Naturels et Anthropisés »; Université de Lyon; Université Lyon 1; ENTPE; 43 boulevard du 11 novembre 1918, 69622 Villeurbanne Cedex, France; * e-mail: [email protected]; phone: +33 (0)4 72 43 12 54

Running title: responses of macrophytes to dewatering

Keywords: aquatic plants, dewatering, plasticity, growth form, phylogeny, plant traits

Total word count: 4214

Number of cited references: 52

Number of figures and tables: 3 figures and 6 tables

RESULTATS Mise en place des réponses plastiques

Abstract

Temporary dewatering constitutes a drastic change in conditions for aquatic vegetation.

Species’ sustained performance under these conditions relies partly on their ability to produce a terrestrial phenotype. Such adaptations may include the development of self-supporting aboveground organs with higher dry matter content enabling plants to withstand gravity and smaller leaves with thicker cuticle to reduce evapotranspiration, leading to lower specific leaf area, higher leaf-construction costs and consequently higher leaf life span. The ability of aquatic plant species to produce a terrestrial-adapted phenotype may differ according to growth form and evolutionary history. The objectives of this study were to 1) measure the effects of dewatering on aquatic plant performance, 2) determine how growth form and phylogenetic position affect performance, and 3) relate plant performance to plasticity. To meet these objectives, we experimentally studied aquatic plant responses to dewatering by measuring survival, growth, and a set of traits describing the morphology and leaf-resource economy of eight aquatic plant species with contrasting phylogeny and growth forms. The ability of aquatic plants to withstand dewatering differed according to phylogeny but not to growth form. The eudicots presented high survival and similar growth rates under terrestrial compared to aquatic conditions, while monocots generally did not survive dewatering. These species produced phenotypic adjustments, such as denser aboveground organs and leaf plasticity, which can explain the maintenance of similar growth rates under terrestrial conditions. The relatively strong plasticity and performance of eudicots in terrestrial habitats suggests that their optimal niche is the interface between aquatic and terrestrial ecosystems.

RESULTATS Mise en place des réponses plastiques

Introduction

Temporary dewatering is one of the most drastic changes that may affect vegetation in aquatic ecosystems, because it represents a major habitat shift from aquatic to terrestrial environment

(Maberly and Spence, 1989; Madsen and Sand-Jensen, 1991). Water becomes limiting, while the solar radiation and gravitational forces encountered by the plants increase substantially, and access to nutrients is modified (Rattray et al. 1991; Niklas 1998; Baldwin and Mitchell 2000;

Rascio 2002). As aquatic species often have thin leaves with reduced cuticles and stomata, and low support tissues compared to terrestrial ones, dewatering frequently induces increased evapotranspiration and rapid leaf desiccation, preventing gas exchange, and leaving the plants unable to support themselves (Nielsen and Sand-Jensen 1997; Wells and Pigliucci 2000; Rascio

2002). Simultaneously, dewatering increases light and CO2 availability, potentially enhancing photosynthesis (Nielsen and Sand-Jensen 1989; Sand-Jensen and Frost-Christensen 1999).

Several strategies are known for allowing plants to withstand temporary dewatering

(Kautsky 1988; Arthaud et al. 2012): seed bank establishment (Casanova and Brock 2000); the ability to remain dormant in the sediment as vegetative propagules (Barrat-Segretain 2001) or to persist as stunted forms without any specific developmental strategy, enabling them to survive but not to grow. However there are species that can develop phenotypes that are better adapted to the new environment (Bradshaw 1965; Sculthorpe 1967; Wells and Pigliucci 2000), enabling them to maintain performance (Robe and Griffiths 1998). The performance (i.e., the capacity to maintain biomass over many generations; McGill et al. 2006) of aquatic species during dewatering relies on their ability to survive, maintain growth (Robe and Griffiths 1998) and reproduce (e.g., through allocation to vegetative or sexual reproduction; Volder et al. 1997), which together determine plant fitness (Geber and Griffen 2003).

Among the functional traits related to plant performance is dry matter content of aboveground organs (stems, leaves). Under terrestrial conditions, higher dry matter content of

RESULTATS Mise en place des réponses plastiques aboveground organs corresponds to a lower proportion of aerenchyma and a greater investment in strengthening tissues, leading to self-supporting aboveground organs able to withstand the forces they encounter (Niklas 1998; Vasellati et al. 2001; Jackson and Colmer 2005; Mommer and Visser 2005; Li et al. 2011; Hamann and Puijalon 2013). This self-supporting growth form permits light harvesting and gas exchange for efficient photosynthesis. Higher dry matter content of leaves may also relate to a higher investment in the cuticle leading to reduced evapotranspiration. The production of leaves with high dry matter content and smaller areas reducing evapotranspiration may lead to a reduction in specific leaf area (SLA). Due to the need to produce more dry matter from a reduced photosynthetic area, the leaf-construction costs may be higher under terrestrial conditions. A longer leaf life span which increases nutrient retention may partially compensate for high construction cost (Ryser and Urbas 2000; Navas et al. 2003;

Wright et al. 2004).

Growth form may also affect plant’s capacity to tolerate dewatering because plant organs differ in size and form and thus interact differently with terrestrial conditions. For submerged rosette species, the ability to maintain an erect position when dewatered mainly relies upon the production of self-supporting leaves, having a higher leaf dry-matter content (Hamman and

Puijalon 2013). The apical meristem of rosette aquatic plants is frequently partly buried in the sediment, and protected from desiccation allowing the plant to produce leaves adapted to terrestrial conditions (Wells and Pigliucci 2000; Grime and Mackey 2002). For caulescent species, increasing stem dry-matter content may facilitate an erect posture and efficient leaf positioning for light capture (Hamman and Puijalon 2013). However, aquatic caulescent plants generally collapse onto the substrate during dewatering. Although the apex may secondarily produce an erect stem and/or axillary buds may be reactivated (Shimizu-Sato and Mori 2001), the efficiency and kinetics of the response are likely to be lower than in rosette species.

RESULTATS Mise en place des réponses plastiques

Finally, all these responses may differ according to evolutionary history (Keeley 1998;

Sultan 2007). Aquatic species belong to a wide range of families, and the shift from terrestrial to aquatic habitats (and vice versa) likely has occurred many times in the course of evolution

(Cook 1999; Rascio 2002). If plant lineages tend to lose the ability to produce ancestral traits when their production is not required for a long time, then taxonomic groups that evolved toward aquatic forms early in angiosperm history may have lost more traits related to terrestrial life compared to taxonomic groups that shifted toward aquatic environments more recently (Les and Sheridan 1990; Germ and Gaberščik 2003). Consequently, such early aquatic groups may exhibit poorer performance under dewatered conditions than groups that shifted toward aquatic environments more recently.

The few studies focusing on the adaptations of aquatic plants to dewatering have often addressed a single species (Volder et al. 1997; Robe and Griffiths 1998; Li et al. 2011), or did not consider the role of growth form and phylogeny (Touchette et al. 2007, 2008; Luo et al.

2008). Consequently, this study aimed as comparing several species for 1) measuring the effect of dewatering on aquatic-plant performance (using survival and growth as surrogates), 2) determining how growth form and phylogeny affect plant performance in dewatered conditions, and 3) determining how plant performance relates to plasticity. To meet these objectives, we studied aquatic-plant responses to dewatering by experimentally measuring survival, growth, and a set of traits describing the morphology and leaf-resource economy of eight aquatic plant species with contrasting phylogeny and growth forms. We hypothesized that species belonging to taxonomic groups that appeared in or colonized aquatic habitats early in angiosperm history would exhibit lower survival and growth under dewatered conditions compared to species belonging to taxonomic groups that shifted toward the aquatic environment more recently. We also hypothesized that for closely related species, those with rosette growth forms would have higher survival and less-impacted growth rates under dewatering conditions compared to

RESULTATS Mise en place des réponses plastiques caulescent species. Finally, in terms of plastic responses, we expected that dewatered conditions would lead to higher aboveground-organ dry matter content, lower leaf area (possibly resulting in lower SLA) and longer leaf life span and that these changes would maintain the performance of aquatic plants under terrestrial conditions.

Materials and methods

Studied species

Eight aquatic plant species were chosen for their contrasting phylogeny and growth forms (Fig.

1). Four species were herbaceous monocots belonging to monophyletically aquatic taxa and consequently were considered as old aquatics. Four species were herbaceous eudicots

(Asterids), considered as more recent aquatic species, because these are isolated species in otherwise terrestrial taxa (Cook 1999; Chambers et al. 2008; APG III 2009). Of the selected monocots, Baldellia ranunculoides (L.) Parl. (Alismataceae) and Sparganium emersum

Rehmann (Typhaceae) have a rosette growth form, while Elodea canadensis Michaux

(Hydrocharitaceae) and Potamogeton coloratus Hornem. (Potamogetonaceae) are caulescent.

The eudicots included two rosette species, Berula erecta (Huds.) Coville (Apiaceae) and

Samolus valerandi L. (Primulaceae), and two caulescent species, Mentha aquatica L.

(Lamiaceae) and Veronica anagallis-aquatica L. (Plantaginaceae). All the species can grow under water.

RESULTATS Mise en place des réponses plastiques

Figure 1. Phylogeny (APG III 2009) and growth forms of the eight species studied.

Plant material

For rosette species, an individual was defined as an entire rosette. For caulescent species, an individual was defined as an anchored erect shoot for V. anagallis-aquatica and P. coloratus, as a 10-cm-long leafy apical-stem fragment without roots for E. canadensis and as a rooted, leafy apical-stem fragment with two internodes for M. aquatica. Each species was collected from a single underwater population in wetlands of the Rhône and Ain Rivers (05°06'00'' E-

45°49'30'' N for E. canadensis; 05°16’53’’E-45°59’06’’N for B. erecta, M. aquatica and P. coloratus; 05°18’2’’E-45°58’24’’N for S. emersum and 05°11’31’’E-45°48’6’’N for S. valerandi). Individuals of B. ranunculoides and V. anagallis-aquatica were obtained by germinating seed collected from two riverine wetlands of the Ain River (05°16’25’’E-

45°58’36’’N and 05°16’53’’E-45°59’06’’N, respectively). The seeds were mixed with natural

RESULTATS Mise en place des réponses plastiques substrate and germinated under controlled conditions, and the seedlings were transplanted in the same manner as the other species for the experiment.

Experimental design

For all species, individuals were cleaned of any sediment particles and weighed (initial fresh mass). Hundred and forty individuals per species were transplanted at random into plastic boxes

(length x height x width: 235 x 95 x 178 mm), each box containing five individuals of the same species. The boxes were filled with 3.5 cm of a substrate composed of 1/3 horticultural compost

3 3 3 (Fertiligène ®; N: 320 g/m , P2O5: 120 g/m , K2O: 140 g/m ) and 2/3 washed river sand and supplemented with tap water. The boxes were then placed in a climate chamber with a temperature of 18 °C at night and up to 25 °C during the day, a relative air humidity of 50% and a 12 hours photoperiod at averaged 75 μmol.m-2.s-1 PPFD. The boxes were randomly assigned to two treatments: control (C), and dewatering (D). After 28 days of acclimatization

(t1), the dewatering treatment was applied, i.e. the boxes were emptied of water. The substrate was kept moist throughout the experiment in the dewatering treatment, while the control boxes were kept full of water throughout the experiment. At the beginning of the experiment, plants in the control treatment were between 1 and 3.5 cm under the water surface depending on the species.

Harvest

After 28 days of acclimatization (t1), 20 individuals of each species were collected from the control treatment. Following the onset of treatment, 18-20 individuals of each treatment and species were collected on three sampling dates (t2, t3 and t4). The sampling dates were separated by 20-27 days. On each collection date, the individuals were collected evenly from all boxes to

RESULTATS Mise en place des réponses plastiques uniformly decrease the plant density in each box over time. Because the individuals collected at t2 and t3 provided more information about the kinetics of the plastic response than about the plants’ performance, we did not include them in the present study but considered only individuals collected at t4.

The collected plants were divided into roots, stems, and leaves, and the different parts were weighed to obtain their fresh and dry mass (measured after drying for 48 hours at 70 °C;

±0.0001 g). The total mass was calculated as the sum of the masses of the transplanted individual and all associated juveniles produced during the experiment.

Performance measurement

To measure plant performance, the survival and relative growth rates were determined.

Individuals were considered dead when no more tissue (underground and aerial parts) were alive, and no living meristem remained in the sediment. The relative growth rate was calculated from the individuals collected at the end of the experiment (t4) as RGR=[ln(final dry mass) – ln(initial dry mass)]/culture time (Poorter and Garnier 1996). To establish the relationship between fresh and dry mass at transplantation (t0), we measured the fresh and dry masses

(measured after drying for 48 hours at 70 °C; ±0.0001 g) of 20 individuals collected from the same locations as the transplanted individuals at the beginning of the experiment. The initial dry masses of the transplanted individuals were obtained using the linear relationship between fresh and dry mass (the regressions were significant for all species; p < 0.001, R2 > 0.8).

Traits of aboveground organs

In the course of dewatering, several traits were expected to change:

RESULTATS Mise en place des réponses plastiques

The production of smaller leaves for reducing evapotranspiration was expected because water becomes limiting. The average leaf size (cm2) was calculated as the total leaf area divided by the number of leaves. To measure the total leaf area, all plant leaves were scanned (150 dpi,

Epson Perfection 4180 Photo), and the images were analyzed using the WinFolia 2006 image- analysis software (Regent Instrument Inc., Quebec, Canada).

We also expected the development of self-supporting above ground organs with higher dry matter content enabling plants to withstand gravity. The dry matter content of above-ground organs was represented by the leaf dry-matter content, calculated as leaf dry mass/leaf fresh mass (LDMC), and the stem dry-matter content, calculated as stem dry mass/stem fresh mass

(SDMC). These two traits provide a different function according to plant growth form because organs interacting with terrestrial conditions differ. For caulescent species SDMC relates to the production of a self-supporting growth form and LDMC relates to efficient leaf positioning.

For rosette species LDMC relates to the production of a self-supporting growth form and efficient leaf positioning while SDMC may relate to other anatomical adjustments to terrestrial conditions.

Leaves were expected to have a higher dry matter content and a smaller area, leading to a reduction in specific leaf area (SLA). SLA (cm2/g) was calculated as total leaf area/leaf dry mass.

Due to the need to produce more dry matter content from a reduced photosynthetic area, the leaf construction cost was supposed to be higher in dewatered conditions. In such situation, a longer leaf life span, which increases nutrient retention, may partially compensate for a higher construction cost. The life span of newly produced leaves under aquatic conditions (control) or after dewatering was measured on 20 additional individuals of each species (10 under aquatic conditions and 10 under dewatering conditions). These individuals were collected at the same locations and cultivated under the same conditions as the transplanted individuals. After the

RESULTATS Mise en place des réponses plastiques acclimation period, each new leaf produced was marked and its life span measured (time between production and decay). Only the life span of leaves produced and dying during the experimental period was taken into account.

Data analysis

All of the statistical analyses were performed using the R.2.10.1 software (R Development Core

Team 2009). The potential effect of initial dry mass on survival was tested by a logistic model with species, initial dry mass, and their interaction as main effects. The initial dry mass had no

2 2 significant effect on survival (initial dry mass: χ 1 = 0.50, p = 0.48; species: χ 7 = 68.74, p <

2 0.001; initial dry mass × species, χ 7 = 6.30, p = 0.51). Consequently it was not considered further in the survival analysis: the effect of treatment, species and their interaction on survival was tested with a logistic model followed by multiple comparisons to test the effect of treatment on survival for each species. To test the effect of 1) growth form (rosette vs caulescent species) and 2) phylogeny (monocotyledon vs eudicotyledon species) on survival for control and dewatering treatments planned comparisons (contrast tests) were then used.

The effect of treatment, species, and their interaction on RGR was tested by a two-way

ANOVA followed by multiple comparisons to test the effect of treatment on RGR for each species. As only one monocotyledon species survived to dewatering treatment, only the effect of growth form on RGR was tested. To test the effect of growth form (rosette vs caulescent species) on the change of growth between treatments, the planned comparison were carried out on the difference of RGR between control and dewatering treatment.

Student’s t-test was used to test treatment effect on leaf life span. Analyses of covariance

(ANCOVA) were used to test the effect of treatment on all plant traits except leaf life span.

Each trait was either dependent on plant size or expressed relative to another trait. Therefore,

RESULTATS Mise en place des réponses plastiques we used ANCOVA to test how the treatments affected the relationships between traits (e.g., allometric relationships) rather than comparing trait combinations (e.g., ratios), which may be biased by non-linear relationships between traits and by ontogenetic drift (Evans 1972;

Jasienski and Bazzaz 1999). The data were log-transformed to improve the normality and homogeneity of the variances. For leaf size, ANCOVA was performed using individual leaf area as the dependent variable and total plant dry mass as a covariate. For specific leaf area,

ANCOVA was performed using total leaf area as the dependent variable and leaf dry mass as a covariate. For the aboveground organ dry matter content, ANCOVA was performed using the organ dry mass (leaves or stem) as the dependent variable and the organ fresh mass as a covariate. Non-significant interaction terms were removed to obtain the final model, and post- hoc tests were performed. When the covariate was correlated with the treatment factor, the

ANCOVA was performed on a subset of individuals to ensure to have the same range of covariate value for the two treatments (between 1 and 4 individuals removed).

For Baldellia ranunculoides, the treatment resulted in great differences between individual sizes (0.01g was the largest dry mass in dewatering treatment and 0.02 g the lowest in control treatment). Consequently it was not possible to have a common range of covariate values for the different traits in the 2 treatments. As a consequence, we did not test the effect of treatment (control vs. dewatering) on the variation of plant traits.

Results

Survival

Survival was significantly affected by the treatment, species and their interaction (Table 1). All species showed a high survival rate in control conditions (Table 2) and survival did not differ significantly according to phylogeny or growth form (Table 3). For dewatering treatment,

RESULTATS Mise en place des réponses plastiques survival did not differ significantly between growth forms, but was significantly lower for the monocots compared to eudicots (Table 3). Among the monocots, the survival of B. ranunculoides did not differ between treatments (Table 2). In contrast, the survival of the three other monocot species (S. emersum, P. coloratus and E. canadensis) was significantly lower in the dewatering treatment compared to the control one (Table 2). For the four eudicot species, survival did not differ significantly between treatments (Table 2).

Table 1. Factors affecting survival tested by logistic regression.

Factor df Wald statistic (χ2) P Treatment 1 86.89 *** Species 7 100.86 *** Treatment × species 7 22.13 ** The Wald statistic used to test the significance of the parameters, degrees of freedom (df) and P-values (P) are indicated. Significant differences: *** p < 0.001; ** p < 0.01.

Table 2. Phylogeny, growth form and survival rate (%) under control and dewatering treatments for each species. The difference between the survival rate under aquatic (control) and dewatered conditions was tested for significance using multiple comparisons following logistic regression (*** p < 0.001; ns, not significant).

Species Phylogeny Growth form Control Dewatering Tukey’s HSD B. ranunculoides Monocot Rosette 100 90 ns S. emersum Monocot Rosette 100 0 *** E. canadensis Monocot Caulescent 100 0 *** P. coloratus Monocot Caulescent 80 0 *** S. valerandi Eudicot Rosette 95 70 ns B. erecta Eudicot Rosette 95 90 ns M. aquatica Eudicot Caulescent 100 80 ns V. anagallis-aquatica Eudicot Caulescent 95 100 ns

RESULTATS Mise en place des réponses plastiques

Table 3. Effect of growth form and phylogeny on survival for control and dewatering treatments tested by planned comparisons (contrast tests) following logistic regression.

Control Dewatering

Growth form effect z = -1.33 ns z = -1.69 ns

Phylogeny effect z = -0.20 ns z = 5.72 ***

Z values and significance levels of the effects are presented. Significant differences: *** p < 0.001; ns, not significant.

Growth

RGR was significantly affected by the treatment, species and their interaction (Table 4). The effect of treatment on RGR differed significantly between growth forms (t = -3.84, p < 0.001).

However among the three rosette species (B. ranunculoides, B. erecta and S. valerandi), only

B. ranunculoides had a significantly lower RGR in the dewatering treatment compared to the control one (Fig. 2). For the two caulescent species (M. aquatica and V. anagallis-aquatica)

RGR did not differ between treatments (Fig. 2).

Table 4. Factors affecting relative growth rate tested by a two-way ANOVA.

Factor Fdf P 1 Treatment F 169= 6.90 ** 4 Species F 169= 85.8 *** Treatment × species F4169= 22.0 ***

Fdf values and significance levels of the effects (treatment, covariate and their interaction) are presented. Significant differences: *** p < 0.001; ** p < 0.01

RESULTATS Mise en place des réponses plastiques

Figure 2. Effect of treatment on relative growth rate (±se) for the species that survived dewatering (C = control, D = dewatering). Significance: *** p < 0.001; ns, not significant (multiple comparisons following two-way ANOVA).

RESULTATS Mise en place des réponses plastiques

Responses of aboveground traits to dewatering

For all traits (stem and leaf dry matter content, leaf size and specific leaf area) and species, the interaction between treatment and covariates (stem and leaf fresh mass, total plant dry mass and leaf dry mass respectively) was not significant (Table 5).

When we investigated how far the dewatering treatment affected species traits, we demonstrated that the specific leaf area and the stem and leaf dry matter content of the four eudicot species differed significantly between treatments. In the dewatering treatment, stem and leaf dry matter content were higher (between 15 and 41% higher and 14 and 28%, respectively;

Table 5, Fig.3) and specific leaf area was lower (between 2 and 6% lower, respectively). Leaf size only differed significantly between treatments for V. anagallis-aquatica, and was lower in dewatering treatment (20% lower, Table 5, Fig.3).

RESULTATS Mise en place des réponses plastiques

Table 5. Effects of treatment on aboveground traits (leaf and stem dry-matter content, leaf size, specific leaf area) in species that survived dewatering (except B. ranunculoides). The significance of each effect was tested using ANCOVA.

Effect Stem dry-matter content Leaf dry-matter content Leaf size Specific leaf area Dependent variable log(stem dry mass) log(leaf dry mass) log(individual leaf area) log(total leaf area) Species Covariate log(stem fresh mass) log(leaf fresh mass) log(total dry mass) log(leaf dry mass) 1 1 1 1 B. erecta Covariate (C) F 32= 166.4 *** F 34= 272.3 *** F 32= 158.8 *** F 33= 143.0 *** 1 1 1 1 Treatment (T) F 32= 15.7 *** F 34= 19.0 *** F 30= 2.10 ns F 33= 5.42 * 1 1 1 1 C × T F 31=2.19 ns F 33=0.002 ns F 30= 3.30 ns F 32= 0.44 ns 1 1 1 1 M. aquatica Covariate (C) F 28= 844.5 *** F 26= 1229.7 *** F 29= 178.2 *** F 24= 244.2 *** 1 1 1 1 Treatment (T) F 28= 58.7 *** F 26= 13.2 ** F 27= 0.12 ns F 24= 5.14 * 1 1 1 2 C × T F 27= 3.78 ns F 25= 0.25 ns F 27= 2.77 ns F 23= 2.59 ns 1 1 1 1 S. valerandi Covariate (C) F 26= 232.1 *** F 26= 411.4 *** F 27= 11.7 ** F 25= 492.2 *** 1 1 1 1 Treatment (T) F 26= 17.4 *** F 26= 10.5 ** F 25= 0.45 ns F 25= 7.84 ** 1 1 1 1 C × T F 25= 3.24 ns F 25= 0.30 ns F 25= 0.18 ns F 24= 1.17 ns 1 1 1 1 V. anagallis-aquatica Covariate (C) F 35= 923.8 *** F 36= 902.5 *** F 34= 10.1 *** F 32= 229.3 *** 1 1 1 1 Treatment (T) F 35= 10.6 ** F 36= 13.6 *** F 34= 9.38 *** F 32= 4.38 * 1 1 1 1 C × T F 34= 2.58 ns F 35= 2.10 ns F 33= 0.45 ns F 31= 1.62 ns

The Fdf values and significance levels of the effects (treatment, covariate and their interaction) are presented. For stem and leaf dry-matter content, stem and leaf dry mass were used as dependent variables, and stem and leaf fresh mass as covariate respectively. For leaf size, individual leaf area was used as dependent variable and total dry mass as covariate. For specific leaf area, total leaf area was used as dependent variable and leaf dry mass as covariate. Non-significant interaction terms were excluded from the final model and are indicated in italics; the F values and significance levels for these terms correspond to the full model in which all terms were present. Significance: *** p < 0.001; ** p < 0.01; * p < 0.05; ns, not significant.

RESULTATS Mise en place des réponses plastiques

Figure 3. Effect of treatment on plastic variations of morphological traits (stem and leaf dry-matter content, leaf size, and specific leaf area) for the four eudicot species (M. aquatica, V. anagallis-aquatica, S. valerandi, B. erecta; C = control, D = dewatering). Points are least-squares means (±se) predicted from the models. Solid lines indicated significant differences and dashed lines non-significant differences between the control and dewatering treatments (see Table 5 for the statistical-test results).

RESULTATS Mise en place des réponses plastiques

Leaf life span differed significantly between treatments only for two species (B. ranunculoides and V. anagallis-aquatica, Table 6). In the dewatering treatment, leaf life span was significantly lower for the monocot species (B. ranunculoides) and higher for the eudicot species (V. anagallis-aquatica).

Table 6. Leaf life span (days, means ±sd) under control and dewatering treatments for species that survived dewatering. The difference between the leaf life span under aquatic (control) and dewatered conditions was tested for significance using Student’s t-test (*** p < 0.001; * p < 0.05; ns, not significant). Species Control Dewatering Student’s t-test B. erecta 31.3±4.5 33.8±11.2 ns B. ranunculoides 62.7±9.4 52.1±6 *** M. aquatica 74±12.4 81.4±9.1 ns S. valerandi 64.8±9.4 62.1±12 ns V. anagallis-aquatica 57.5±10.2 73.1±3.9 *

Discussion

In agreement with our hypothesis, the aquatic plants’ ability to tolerate dewatering differed according to the phylogeny of the species. The four eudicot species exhibited high survival rates and similar growth rates under dewatered conditions compared to aquatic conditions. In contrast, monocots generally did not survive dewatering, as only B. ranunculoides survived with a lower growth rate under dewatered compared to aquatic conditions. This low capacity of Alismatales to withstand dewatered conditions is surprising because members of this group are frequently considered to have high phenotypic plasticity (e.g. Sagittaria sagittifolia;

Sculthorpe 1967; Cook 1990). Because of their earlier shift to aquatic habitats (Cook 1999;

Chambers et al. 2008), monocots possess strong morphological, anatomical and physiological adaptations to aquatic life (e.g., superficial rooting, substantial aerenchyma development, and dissected and/or thin leaves; Rascio 2002), possibly diminishing their survival under dewatering conditions. Furthermore, due to high maintenance costs (De Witt et al. 1998), these lineages may have lost key functions that would enable them to develop terrestrial-adapted

RESULTATS Mise en place des réponses plastiques phenotypes (e.g., ancestral terrestrial traits such as vascular and structural tissues, cuticle and stomata). The production of terrestrial-adapted phenotypes in monocots may also require more progressive stimulation than in the present study. Finally, the developmental stage of the individuals used in the present experiment may have been too early; even if the early stages of development can be more plastic (McConnaughay and Coleman 1998; Wright and

McConnaughay 2002), the individuals possessed no rhizomes or storage organs, which might have enabled them to construct new leaves and stems after dewatering.

Contrary to our hypothesis, we found that growth form did not determine the plants’ responses to dewatering in terms of plasticity or performance. All eudicot species, whatever their growth forms, survived and maintained their growth rate. All these species have higher stem dry matter content, but also higher leaf dry matter content, which is, as expected associated with a lower specific leaf area under dewatering conditions. For rosette species, higher stem dry matter content may reflect strengthening vascular tissues for decreasing the risk of embolism and increasing the efficiency of xylem-sap transport (Pedersen and Sand-Jensen

1993; Vasellati et al. 2001), floral-axis development (reorganization of vegetative meristems to form floral meristems), or increased resource allocation to storage organs (the stem in the case of rosette species; Puijalon et al. 2008). The higher leaf dry matter content observed in all studied species after dewatering indicated that these species produced either self-supporting leaves, allowing a more efficient photosynthetic light harvesting and gas exchange, and/or leaves with a thicker cuticle reducing water loss. The greater investment in leaf tissue was not accompanied by an increase in leaf life span, but the potentially higher leaf-construction cost may be compensated for by more favorable conditions for photosynthesis under terrestrial conditions (e.g., higher light and CO2 availability; Sand-Jensen and Frost-Christensen 1999).

The reduction of leaf size reducing water loss in dewatered conditions was observed only for

V. anagallis aquatica. It corresponds to a reduction of their photosynthetic surface, but the

RESULTATS Mise en place des réponses plastiques potential decrease of photosynthetic performance may be compensated for by an increased leaf life span for this species. All eudicot species performed similarly under submerged and emergent conditions, probably due to accommodation to both environments, through plasticity.

This work also highlights the difficulty to study phenotypic plasticity especially in the case where the condition change causes a difference in size. For B. ranunculoides, dewatering led to reduced size and therefore made difficult to study allometric relationships. In this case, it was impossible to determine if the variation in morphological traits was the result of ontogenetic drift or resulted from a functional adjustment (Evans 1972). It would be interesting to examine the variation in morphological traits for individuals of the same size range, even if not of the same age, which would require to make measurements over the different stages of growth

(Coleman et al. 1994).

- Considering certain morphological and physiological traits (the inability to use HCO3 , homophylly, thick leaves and leaf cuticles; Sand-Jensen and Frost-Christensen 1999), B. erecta,

M. aquatica and V. anagallis-aquatica appeared poorly adapted to aquatic life. Several studies have reported that these species have higher photosynthetic rates in terrestrial than in aquatic habitats due to their low affinity for CO2 in water, despite other physiological adaptations to aquatic life (Sand-Jensen et al. 1992; Sand-Jensen and Frost-Christensen 1999). The relatively strong plasticity and performance of eudicot species in terrestrial habitats, suggests that their optimal niche is the interface between aquatic and terrestrial ecosystems.

Conclusion

This study is the first attempt to build a predictive model of aquatic plant response to dewatering. Indeed, we demonstrated that phylogeny is the key element for predicted plant tolerance to dewatering, whereas growth form seems of minor importance. The relatively strong

RESULTATS Mise en place des réponses plastiques plasticity and performance of eudicot species in terrestrial habitats, suggests that their optimal niche is the interface between aquatic and terrestrial ecosystems.

The adaptive value of the responses observed has still to be quantified. Indeed, the establishment of a plastic response allowing species to maintain similar growth rates in dewatered and aquatic conditions may have repercussions on energy allocated to reproduction, because of the contrasting construction costs of submerged vs. emerged phenotypes.

Acknowledgments

M.R. Viricel is gratefully acknowledged for technical assistance. This research was performed under the aegis of the LTER “Zone Atelier Bassin du Rhône” and was funded by the Wetchange

Program (ANR-09-CEP-006-01) of the French National Research Agency (Agence Nationale de la Recherche-ANR).

RESULTATS Mise en place des réponses plastiques

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RESULTATS Comparaison des phénotypes inondés et exondés

2.2 Comparaison des phénotypes inondés et exondés

Mélissa De Wilde, Julien Clavel, Jehanne Oudot-Canaff, Quitterie Hugues, Evelyne Martel,

Gilles Escarguel, Sara Puijalon, Gudrun Bornette. A comparative study of the phenotypic plasticity in aquatic angiosperms to dewatering. (en préparation)

Problématique

L’objectif de cette article est d’identifier les ajustements plastiques que les espèces d’Angiosperme aquatiques mettent en place lorsqu’elles sont exondées et le déterminisme des ces variations phénotypiques.

Les différentes hypothèses testées dans ce travail sont :

1) Les variations phénotypiques sont similaires pour les espèces proches

phylogénétiquement, indépendamment de leur niche écologique, car le conservatisme

phylogénétique peut gouverner les variations phénotypiques.

2) La forme de croissance affecte les traits morphologiques impliqués dans la variation

phénotypique et la performance des plantes

3) La variation phénotypique entraine une meilleure performance.

Afin de tester ces hypothèses, pour 19 espèces de plantes aquatiques ayant des origines phylogénétiques, des niches écologiques (hydrophyte vs amphiphyte) et des formes de croissance (rosette vs caulescente) différentes, des mesures in situ de traits morphologiques ont

été réalisées afin d’obtenir des informations sur leur phénotype et leur performance en conditions inondée et exondée. Les reconstructions de l’utilisation des habitats ancestraux et de l’acquisition des forme de croissance déduis de phylogénies calibrées dans le temps ont été utilisées pour ajuster des modèles d’évolution de la performance et la variation phénotypique des espèces.

RESULTATS Comparaison des phénotypes inondés et exondés

A comparative study of the phenotypic plasticity in aquatic angiosperms to dewatering

Mélissa De Wilde*1, Julien Clavel2, Jehanne Oudot-Canaff1, Quitterie Hugues1, Evelyne Martel1, Gilles

Escarguel2, Sara Puijalon1, Gudrun Bornette3

1 UMR CNRS 5023 « Laboratoire d’Ecologie des Hydrosystèmes Naturels et Anthropisés »; Université de Lyon;

Université Lyon 1; ENTPE; 43 boulevard du 11 novembre 1918, 69622 Villeurbanne Cedex, France; 2 UMR

CNRS 5276 « Laboratoire de Géologie de Lyon, Terre, Planètes, Environnement », 3 UMR CNRS 6259

« Chronoenvironnement »,Université de Franche Comté, Campus de la Bouloie, 16, route de Gray, 25000

Besançon Cedex

* e-mail: [email protected]; phone: +33 (0)4 72 43 12 54

En préparation

RESULTATS Comparaison des phénotypes inondés et exondés

Introduction

The colonization of aquatic habitats by angiosperms result from many terrestrial lineages, and occurred 2, sometimes 3 times in most plant families, and several hundred times among angiosperms (Cook 1999). These repeated events of colonization of aquatic habitats from terrestrial lineages were sometimes only involve one genus, or sometimes containing only one species, growing in aquatic habitats (Cook 1990). In some other situations, some orders or family lineages are strictly and fully aquatic and phylogenetically basal (e.g Nympheales,

Alismataceae, Potamogetonaceae, Hydatellaceae, Acorales, Ceratophyllales). Despite this high diversity of evolutionary histories, aquatic angiosperms show convergent evolution of traits related to the aquatic life (Les et al. 1991 ; Cook 1999 ; Rascio 2002). Among morpho- anatomical traits the presence of aerenchyma, the reduction of support tissues, thin leaves, with reduced or absent cuticle are quite systematically observed.

Due to global change, the water deficiency may increase in temperate areas, leading to the fact that aquatic plants may be more frequently and durably submitted to dewatering (Winter

2000 ; Erwin 2009 ; Junk et al. 2013). Dewatering induces drastic changes for aquatic plants: water becomes limiting, there is no more limitation of gas and light, the access to nutrients is modified because of the disappearance of the water column and the changes in biogeochemical processes occurring at the sediment, and plants are exposed to higher gravitational forces

(Maberly & Spence 1989 ; Madsen & Sand-Jensen 1991 ; Rattray et al. 1991; Niklas 1998 ;

Baldwin & Mitchell 2000 ; Rascio 2002). For aquatic plants, a way to survive in this new environment is to develop a new growth form through phenotypic plasticity (Bradshaw 1965 ;

Sculthorpe 1967 ; Wells & Pigliucci 2000).

Depite the significant literature dealing with the phenotypic plasticity of aquatic plants in response to dewatering (Millington & Schmidt 1968 ; Bruni et al. 1996 ; Loreti & Oesterheld

1996 ; Niklas 1998 ; Vasellati et al. 2001; Jackson & Colmer 2005 ; Mommer & Visser 2005 ;

RESULTATS Comparaison des phénotypes inondés et exondés

Geng et al. 2006 ; Touchette et al. 2008 ; Li et al. 2011 ; Hamann & Puijalon 2013), most of these studies focused on a small set of species, and used contrasting experimental conditions, making difficult to establish universal patterns of responses to dewatering. However, as aquatic angiosperms may belong to contrasting phylogenetic lineages, and have various ecological niches and growth forms, they may have potentially diverse plastic responses to dewatering and related performance.

The variation in traits across species are influenced by the convergent adaptations of species to their present niche but also by phylogenetic conservatism i.e. conservation over time of traits inherited from ancestors (Prinzing et al. 2001 ; Hansen & Orzack 2005). The groups being returned early to aquatic life may have lost most of the traits allowing life on terrestrial conditions (ancestral traits) that groups with more recently evolved into an aquatic environment

(Les & Sheridan 1990 ; Germ & Gaberščik 2003). In this case the older aquatic groups may show, when dewatered, lower plasticity and performance than the groups that colonized more recently aquatic environment.

Aquatic angiosperm species are commonly classified into three main ecological niches for which they should present convergent adaptations: the hydrophytes (strictly aquatic in their vegetative phase); the amphiphytes (able to tolerate emersion if the sediment remain wet), and the helophytes (emerged plants tolerating partial sediment drying). In this work, only hydrophytes and amphiphytes were considered; both are distributed in the various phylogenetic groups comprising aquatic species. Some amphiphyte species show few traits associated with aquatic life and relatively low leaf plasticity (homophyllous species). These species may be less tolerant to aquatic conditions than to terrestrial ones (among them, Myosotis scorpioides, Germ

& Gaberscik 2003 ; Berula erecta, Mentha aquatica, Nasturtium officinale and Veronica anagallis-aquatica, Sand-Jensen Frost & Christensen 1999 ; Madsen & Breinholt 1995 ;

Nielsen 1993 , Nielsen & Sand-Jensen 1993 ; Nielsen & Sand-Jensen 1997; Pedersen & Sand-

RESULTATS Comparaison des phénotypes inondés et exondés

Jensen 1992). However other amphiphyte species, among them many Alismatales species, have high leaf plasticity, suggesting a better ability to tolerate phase shifts.Similarly, some hydrophyte species have high plasticity when dewatered (e.g. some Ranunculus species

Gaberscik & Germ 2003 ; Bruni et al. 1996); but no information about the related plant performance is available. Therefore relationships between morpho-anatomical plasticity and performance of dewatered species appear consequently difficult to establish.

Aquatic plants have different growth forms, as rosette and caulescente species. For aquatic species with rosette growth form, the ability to produce a terrestrial phenotype is mainly based on the production of self-supporting leaves (Hamann & Puijalon 2013). The very short stem bearing the apical meristem undergoes little or no gravity and because of its partial burial in the sediment remains protected from desiccation, which should allow the plant to produce quickly leaves adapted to terrestrial conditions (Wells & Pigliucci 2000 ; Grime & Mackey

2002). For caulescent species, producing a self-supporting stem is required to maintain an erect posture and allow effective leaf positioning relative to light (Hamann & Puijalon 2013). When dewatered the stems fall on the sediment, and must issue secondarily a new self-supporting stem from any terminal or axillary bud (Shimizu-Sato & Mori 2001), the efficiency and kinetics of the response are likely to be lower than in rosette species.

The present work intended to identify plastic adjustments to aerial life among aquatic plant species and their determinism. We expected that this phenotypic variations might be similar for closer taxonomic groups, independently from their ecological niche, because phylogenetic conservatism may rule the phenotypic variation. We also expected that species growth form may affect morphological traits involved in the phenotypic variation, and plant performance. Finally, we expected that the trait variation may lead to a better plant performance in dewatered environment compared to the submeregd one.

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For testing these hypotheses, for nineteen aquatic plant species of contrasting phylogenetic origins, ecological niche (hydrophyte vs amphiphyte) and growth form (rosette vs caulescent), in situ measurements of a set of morphological traits were assessed in dewatered and submerged conditions allowing to bring information about plant phenotypes and performance in both conditions. Ancestral habitat use reconstructions and growth form acquisition inferences on reconstructed time-calibrated phylogenies were used to fit evolutionary models of species performance and morphological traits variability.

Materiel and method

1. Plant sampling

The study was conducted on 19 species of aquatic Angiosperms selected in order to maximize the variability of phylogenetic membership (14 families over 11 orders are represented), ecological niche (amphiphytes vs. hydrophytes) and growth form (rosette vs. caulescent). (Table

1). For each species, at least 10 fully developed individuals were collected in July 2010 under submerged and dewatered conditions in cut-off channels of the Ain River and shallow lakes of the Dombes, in France. For a given species, submerged and dewatered individuals were collected in close locations in the same sites, on the same date to minimize the risk of taking different populations.

RESULTATS Comparaison des phénotypes inondés et exondés

Table 1. Sampling site, order, family, ecological niche and growth form of the species.

Sampling size Species (submerged/dewatered Sampling site Order Familly Ecological niche Growth form individual) Alisma plantago-aquatica L. 11/15 Dombes shallow lake Alismatales Alismataceae Amphipyte Rosette Berula erecta (Huds.) Coville 14/12 CFO Apiales Apiaceae Amphipyte Rosette Elatine alsinastrum L. 15/15 Dombes shallow lake Malpighiales Elatinaceae Amphipyte Caulescent Hottonia palustris L. 12/12 CFO Ericales Primulaceae Hydrophyte Caulescent Hippuris vulgaris L. 11/12 ALB Lamiales Plantaginaceae Hydrophyte Caulescent Juncus articulatus L. 15/15 ALB Poales Juncaceae Amphipyte Caulescent Luronium natans (L.) Raf. 15/15 SBR Alismatales Alismataceae Hydrophyte Rosette Ludwigia palustris (L.) Elliott 15/15 Dombes shallow lake Myrtales Onagraceae Amphipyte Caulescent Mentha aquatica L. 13/12 ALB Lamiales Lamiaceae Amphipyte Caulescent Myriophyllum verticillatum L. 15/15 VIL Saxifragales Haloragaceae Hydrophyte Caulescent Nuphar lutea (L.) Sm. 10/10 SBR Nympheales Nympheaceae Hydrophyte Rosette Oenanthe aquatica (L.) Poir. 10/10 Dombes shallow lake Apiales Apiaceae Amphipyte Rosette Polygonum amphibium(L.) Delarbre 15/13 Dombes shallow lake Caryophyllales Polygonaceae Amphipyte Caulescent Potamogeton gramineus L. 15/15 Dombes shallow lake Alismatales Potamogetonaceae Hydrophyte Caulescent Potamogeton coloratus Hornem. 15/15 ALB Alismatales Potamogetonaceae Hydrophyte Caulescent Ranunculus flammula L. 10/15 Dombes shallow lake Ranunculales Ranunculaceae Amphipyte Rosette Ranunculus peltatus Schrank 12/15 Dombes shallow lake Ranunculales Ranunculaceae Hydrophyte Caulescent Sparganium emersum Rehmann 14/15 CFO Poales Typhaceae Amphipyte Rosette Sagittaria sagittifolia L 15/15 Dombes shallow lake Alismatales Alismataceae Amphipyte Rosette Sampling sites were riverine wetlands of the Ain River (ALB : 05°15’40’’E, 45°58’21’’N ; CFO : 05°13’52”E 45°50’08”N; SBR : 05°14’43”E 45°49’24”N; VIL : 05°16’57”E 45°59’08”N) and Dombes shallow lakes (05°06’01’’E 46°05’45’’N and 05°07’18’’E 46°03’18’’N).

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2. Measurements of phenotypic traits

Phenotypic traits were measured on the submerged and dewatered individuals immediately after plant sampling. Each sampled plant was measured (plant height), then divided into 3 parts: leaves, stem and roots. Leaves were scanned (150 dpi, Epson Perfection 4180 Photo). The different parts were weighed to obtain their fresh and dry mass (measured after drying for 48 hours at 70 °C; ±0.0001 g).

The contrast between the phenotypes in submerged and dewatered conditions was assessed through the measurement of a set of traits representative of plant performance and morphology at both plant and leaf scale.

Plant performance was assessed by the measure of plant total dry mass and plant height.

Water becomes limiting and light non-limiting when the plant is dewatered, which may lead to a lower photosynthetic area (leaf area) for dewatered individuals (Bruni et al. 1996 ;

Loreti & Oesterheld 1996 ; Li et al. 2011). The total leaf area (cm2) was assessed through the analyse of images of scanned leaves using WinFolia 2006 image-analysis software (Regent

Instrument Inc., Quebec, Canada). As leaf area may vary with plant size, leaf area was corrected by the total dry mass.

Plants were likely to invest more in the underground organs when dewatered, for optimizing water uptake or resource storage, which was supposed to result in an higher root/shoot ratio for dewatered plants (Vasellati et al. 2001 ; Hussner et al. 2008 ; Touchette et al. 2008).

Due to the higher gravitational constraints, the development of self-supporting above ground organs with higher dry matter content enabling plants to withstand gravity is expected

(Hamman & Puijalon 2013). The leaf dry-matter content was calculated as leaf dry mass/leaf fresh mass (LDMC), and the stem dry-matter content was calculated as stem dry mass/stem fresh mass (SDMC). For caulescent species, SDMC relates to the production of a self-

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RESULTATS Comparaison des phénotypes inondés et exondés supporting growth form and LDMC relates to efficient leaf positioning. For rosette species,

LDMC relates to the production of a self-supporting growth form and efficient leaf positioning while SDMC may relate to other anatomical adjustments not simply related to gravity.

Leaves were expected to have a higher dry matter content and a smaller area in dewatered situations, leading to a reduction in specific leaf area (SLA ; Iida et al. 2007). SLA (cm2/g) was calculated as total leaf area/leaf dry mass.

Concerning leaf morphology, three indices were used for describing variations of leaf shape: the blade length/ blade width ratio (leaf shape), the petiole length/ blade length ratio

(petiole-blade ratio), and the dissection index,(perimeter/ blade length). In dewatered conditions, plants are expected to have a lower blade length/ blade width ratio, because of less elongated leaves, a lower petiole-blade ratio, allowing them to better support leaves, and a lower dissection index, as they should have less dissected leaves (Bruni et al. 1996 ; Millington &

Schmidt in 1968 Wells & Pigliucci 2000 ; Iida et al. 2007 ; Wanke 2011). A sample of 5 fully developed upper leaves was selected for each individual, scanned, and analyzed with WinFolia

2006 image-analysis software (Regent Instrument Inc., Quebec, Canada). All these measurements were done on leaves, except Berula erecta, whose measurements were done on leaflets (one leaflet per leaf). The petiole/blade ratio was not considered for sessile species:

Elatine alsinastrum, Hippuris vulgaris, Juncus articulatus and Sparganium emersum.

3. Genetic data acquisition and phylogenetic reconstruction

We constructed a dated phylogeny by combining sequence data of the rbcL plastid gene and matK mitochondrial gene (e.g., Hilu et al. 2003 ; Soltis et al. 2011). DNA sequences from 49 angiosperms species belonging to the sampled families were obtained from GenBank

(accession number given in Table S1). We used Amborella trichopoda, proposed as the single

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RESULTATS Comparaison des phénotypes inondés et exondés sister species to all other extant angiosperms (Amborella Genome Project, 2013) as well as

Cycas revoluta, Ginkgo biloba and Picea abies to root the tree. Because of uncertainty regarding the basal branches of angiosperms (e.g., Clarke et al. 2011) Amborella trichopoda position was not a priori forced to be in the outgroup. When multiple hits were found in the

GenBank repository, the longest and least degenerated sequences were chosen. In addition, sequences for the matK and rbcL genes representing Ranuculus peltatus, Potamogeton gramineus, Oenanthe aquatica, Ludwigia palustris and Elatine alsinastrum species for which no data were available were sequenced. Total DNA of these six species (Elatine alsinastrum,

Ludwigia palustris, Oenanthe Aquatica, Potamogeton gramineus and Ranunculus peltatus) were extracted from leaves and buds tissues using the CTAB protocol (Doyle, 1991). The rbcL gene from E. alsinastrum, O. aquatica and R. peltatus were amplified using the polymerase chain reaction (PCR) with the primers: rbcL1F (5’-ATGTCACCACAAACAGAGACT-3’) and rbcL1369R (5’-TTCCATACTTCACAAGCAGC-3’). The matK gene from E. alsinastrum, L. palustris and P. gramineus were amplified using the polymerase chain reaction (PCR) with the primers: TrnKmatKF (5’-GGTAGAGTACTCGGCTTTTA-3’), trnKmatKR (5’-

GGGTTGCCCGGGACTCGAAC-3'), ElaT-MKF (5’-CTGTATCGCACTATGTATC-3’),

PoM-MKF (5’-GACCATATCGCACTATGTATC- 3’) and LuP-MKF (5’-.

GGCTGTATCGCACTATGTATC-3’). Amplifications were performed in 25μl reactions containing: 10ng DNA solution, 1x PCR buffer (QIAGEN, France), 0.8mM dNTPs, 0.4μM of each primer, 1 U of Taq polymerase (QIAGEN, France). PCR was carried out in a PTC-200 thermocycler (MJ Research) with the following settings: an initial denaturation at 94 °C for

3 min, followed by 35 cycles of 1min at 94 C, 1min at 55°C, and 90 s at 72 C; followed by one step of 5 min at 72 C. The PCR products were purified and bidirectionally sequenced by

Biofidal (Lyon, France). Sequences were edited and assembled using Sequencher (Gene Codes

Corporation, United States). All sequences have been deposited in GenBank (for accession

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RESULTATS Comparaison des phénotypes inondés et exondés numbers see Table S1). All sequences were aligned with Muscle implemented in SeaView

(version 4.4.0; Gouy et al. 2010) and the minor adjustments for ambiguous sites were done manually.

A Bayesian inference analysis was performed with BEAST (1.7.5) to simultaneously estimate the tree topology and branching time (Drummond & Rambaut 2007). We used lognormal uncorrelated relaxed clocks (Drummond et al. 2006) with exponential prior distributions to accommodate for the uncertainty in the calibration of the molecular clock.

Minimum and maximum age constraints for 17 nodes in the phylogeny were used for time calibrating the tree (See Appendix 1). The maximum bound was made soft by allowing that 5% of the probability distribution exceed the specified limit (e.g., Clarke et al. 2011 ; Warnock et al. 2011). This was achieved by accommodating the rate parameter of the exponential distribution so that the 95 percentile of the upper tail correspond to the maximum fossil-based age estimate. We used a Yule model for the speciation model (conditioning the initial rate parameter on the mean root age and the number of species – See Appendix 2). The matK and rbcL sequences were, respectively, best fitted by a GTR+Γ and a TVM+I+Γ substitution model according to information criteria AIC, AICc and BIC in jModelTest (version 0.1.1; Posada

2008). The TVM model of substitution from jModelTest is not provided in default settings of the BEAST software, and was consequently incorporated by hand-edited XML files. Two independent analyses with Markov chains of 50,000,000 generations were run simultaneously and sampled every 1000 generations. Convergence of the Markov chains and the number of generation to be discarded as burn-in (the first 5,000,000 generations) were assessed using the program Tracer (Drummond et al. 2006). This resulted in effective sample sizes for the parameters estimates with acceptable values (ESS>200) indicating that the chains has been run for an adequate length (Drummond et al. 2006). We finally retained 1000 topologies randomly

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RESULTATS Comparaison des phénotypes inondés et exondés sampled in the posterior stationary distribution for use in the subsequent analyses. This collection of trees was used to account for uncertainty in tree topology and branch lengths.

4. Data analysis

Species trait variation

For each trait, the significance of the variation between submerged and dewatered conditions was tested. Student’s t-tests were used for testing the significance of variation between submerged and dewatered conditions on total dry mass and plant height. For testing the significance of variation between the two conditions on size-related traits (leaf area) and on the relation between the traits rather than by directly comparing the ratio, analyses of covariance

(ANCOVA) were done. For total leaf area, ANCOVA was performed using total leaf area as the dependent variable and total plant dry mass as a covariate. For the aboveground organ dry matter content, an ANCOVA was performed using the organ dry mass (leaves or stem) as the dependent variable and the organ fresh mass as a covariate. For the belowground dry mass/aboveground dry mass ratio, ANCOVA was performed using the belowground dry mass as the dependent variable and the aboveground dry mass as a covariate. For the leaf shape index,

ANCOVA was performed using the blade width as the dependent variable and the blade length as a covariate. For the petiole-blade ratio, ANCOVA was performed using the petiole length as the dependent variable and the blade length as a covariate. For the dissection index, ANCOVA was performed using the blade perimeter as the dependent variable and the blade length as a covariate. The data were log-transformed to improve the normality and homogeneity of the variances. Non-significant interaction terms were removed to obtain the final model. For all analyses, sequential Bonferroni corrections were used (Dunn–Sidak method; Sokal & Rohlf,

1995). Statistical analyses were performed using the R.2.10.1 software (R Development Core

Team 2009).

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Comparison of model fitting of trait plasticity evolution

Standardization of phenotypic traits and calculation of the variation norms

To suppress the size contrast between species and compare the species variation norms in the same reference framework, phenotypic traits were normalized and centered by species. The average phenotypic variation norms (difference between submerged and dewatered phenotype; here ratio were used) and error standard were estimated by bootstrap iterations (1000 reiterations).

Ancestral states estimations

Stochastic character mapping (Huelsenbeck & Bollback 2001 ; Huelsenbeck et al. 2003 ;

Bollback 2006) was used to reconstruct the history of the acquisition of the two contrasted growth form (caulescent and rosette) as well as ecological niche (amphiphytes and hydrophytes) for the 19 species in the trait dataset. These two grouping schemes represents different selective regimes in which condition (submerged vs dewatered) and phenotypic traits are associated with a particular growth form or ecological niche (Baum & Larson 1991). Ancestral states reconstructions with stochastic mapping was done on the 1000 subsampled trees using the

“make.simmap” function in the R package phytools (Revell 2012). This is achieved in order to take into account uncertainty in ancestral state reconstructions as well as phylogeny estimation in the subsequent analysis (e.g., Collar et al. 2011; Price et al. 2013), but it should be noted that such an approach could lead to a decrease in statistical power (prone to more type II errors;

Revell 2013).

Model fitting

Phenotypic data were analyzed using a model comparison approach for testing six evolutionary hypotheses regarding phenotypic responses of aquatic plants to dewatering while accounting for inter-species statistical dependences. The evolutionary hypothesis that phenotypic

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RESULTATS Comparaison des phénotypes inondés et exondés variations depend on growth form and ecological niche (caulescent or rosette – amphiphytes or hydrophytes) for species related by a phylogenetic tree could be represented by different models of phenotypic traits evolution (e.g., Collar et al. 2011). Competing models of continuous traits evolution could be the Brownian motion (BM) and the Ornstein-Uhlenbeck (OU) process

(Felsenstein 1985, 1988; Hansen & Martins 1996; Hansen 1997). Brownian motion is a random drift model were phenotypic similarities are on average proportional to shared evolutionary history (Felsenstein 1985, 1988). In this model, the expected phenotypic variance at each instant in time (i.e. through the phylogeny) is related to the rate parameter σ². The Ornstein-Uhlenbeck process is a model that assumes stabilizing selection across the phylogeny with a common phenotypic optimum for all species (Felsenstein 1988 ; Hansen 1997; Butler & King 2004).

Under OU process, the phenotypic variance of the Brownian process represented by the rate parameter σ² is no longer related to time and shared ancestry between species but is constrained around an optimum value by the selection parameter α. Note that under such a process, when the strength of the selection parameter α increases, the interspecies autocorrelation decreases exponentially (e.g., Felsenstein 1988 ; Ho & Ané 2013).

Phenotypic variation for which species with contrasting niches or growth forms are allowed to evolve toward different phenotypic optima could be represented by multiple-peak

Ornstein-Uhlenbeck process (with the same strength of selection and Brownian rate for all selective regimes; e.g. Hansen 1997 ; Butler & King 2004). These evolutionary hypotheses of phenotypic variations depending on ecological niches and growth forms are represented by the

OUMamp and OUMros models. The ecological niches and growth forms could also affect differently the phenotypic variations. For example, within a given lineage, the phenotypic variation may be greater for species colonizing highly variable niches (e.g. amphiphytes). This could be approximated using BM models (BMMros and BMMamp), which allow evolutionary rate to vary across lineages depending on the ecological niches or growth forms of the species

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(e.g., O’Meara et al. 2006 ; Thomas et al. 2006). Finally, we also fitted a simple Brownian motion (BM1) and Ornstein-Uhlenbeck model with a unique optimum (OU1) to the data. While they represent two distinct evolutionary hypotheses and way of phenotypic variance accumulation through time, these two models assume a common evolutionary process for the whole species, and as such could be seen as null models with respect to the hypothesis of a phenotypic variation related to growth form or ecological niche (i.e. the selective regime) .

The models were fitted to the trait data and the 1000 subsampled phylogenies using maximum likelihood, as implemented in the R package mvMORPH 1.0.2 (Clavel et al. submitted). Measurement error or intra-specific variance (sensu Ives et al. 2007) was included as the estimated variance of the mean value of a given trait and estimated by bootstrap (see section Standardization of phenotypic traits and calculation of the variation norms). Model fitting comparisons were done by computing Akaike weights with the corrected AICc criterion

(Hurvich and Tsai 1989; Burnham and Anderson 2002). Convergence in the likelihood search and reliability of the maximum likelihood parameters estimates were assessed using the outputs of the mvMORPH package. All computations were performed on a SGI Altix Xe 320 Cluster with a Linux platform and using a 64-bit version of the open access R-3.0.1 (Ihaka & Gentleman

1996 ; R Development Core Team 2005). All of the cited R packages are freely available online on the CRAN repository (Comprehensive R Archive Network; http://www.r-project.org).

Phenotypic traits and performances

How plant phenotypic variations when dewatered led to an increase of plant fitness was assessed using generalized least square models (GLS), in which the phylogenetic covariance between species in the error terms was represented by an Ornstein-Uhlenbeck model. As stated before, interspecific covariations under an OU model were expected to decrease when the selective parameter α increased. Thus, the OU model was considered as a measure of phylogenetic signal or effect (sensu Hansen et al. 2008) which could be jointly estimated with

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RESULTATS Comparaison des phénotypes inondés et exondés the parameters of the linear model. The phenotypic traits were used as predictor variables to test whether trait variations changes significantly affected performance (expressed by the total dry mass). Linear models fitting was done using the R package phylolm (Ho and Ané 2014).

Results

Phylogenetic tree

Bayesian phylogenetic analysis in BEAST resulted in a phylogenetic tree (Fig 1) that was in a broad agreement with current knowledge about angiosperm phylogenic relationships (APGIII,

2009).

Rosette

Figure 1.Time calibrated phylogeny of the 19 Angiosperm aquatic species studied. Colored boxes next to species names indicate species ecological niche and growth form.

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Species traits variation

Performance traits

Total dry mass differed significantly between submerged and dewatered conditions for

14 species (9 species after sequential Bonferroni correction), it was higher in dewatered conditions for 6 species (5 species after sequential Bonferroni correction) and lower for 7 species (4 species after sequential Bonferroni correction; Table 2). Plant height differed between both conditions for 14 species (11 species after sequential Bonferroni correction), it was higher in dewatered conditions for 4 species (2 species after sequential Bonferroni correction) and lower for 10 species (9 species after sequential Bonferroni correction, Table 2).

Morphological traits

Total leaf area/ total dry mass was lower in dewatered conditions for 17 among the 19 species studied (13 species remained significant after sequential Bonferroni correction; Table

3). Dry matter content of aboveground organs showed a common trend with higher leaf dry matter content in dewatered conditions for 10 species (5 species after sequential Bonferroni correction) and higher stem dry matter content in dewatered conditions for 6 species (6 after sequential Bonferroni correction; Table 3). The root/shoot ratio varied significantly for 9 species (5 after sequential Bonferroni correction), it was higher in dewatered conditions for 7 species (3 species after sequential Bonferroni correction) and lower for 2 species (Table 3). The petiole-blade ratio varied significantly for 11 among the 15 petiolated species, it was lower in dewatered conditions for 8, and higher for 3 species. Leaf shape varied significantly for 10 species (5 after sequential Bonferroni correction). It was higher in dewatered conditions for 7 species (3 after sequential Bonferroni correction) and lower for 3 (2 after sequential Bonferroni correction; Table 3). Finally, the dissection index varied significantly for 6 species (4 after sequential Bonferroni correction), it was higher in dewatered conditions for 1 species and lower for 5 species (3 after sequential Bonferroni correction ; Table 3). 110

RESULTATS Comparaison des phénotypes inondés et exondés

Table 2. Contrasts in total dry mass and plant height between submerged and dewatered conditions tested for each species using Student’s t-test (*** p < 0.001; ** p < 0.01; * p < 0.05; ns, not significant).

Species log(total dry mass) log(plant height) Alisma plantago-aquatica t = 4.49, df = 23.7 *** (+) t = -5.38, df = 23.4 *** (-) Berula erecta t = 6.52, df = 13.4 *** (+) t = 7.66, df = 23.5 *** (+) Elatine alsinastrum t = 3.73, df = 28.9 *** (+) t = 4.74, df = 20.8 *** (+) Hippuris vulgaris t = -5.20, df = 15.9 *** (-) t = 0.11, df = 19.9 ns Hotonia palustris t = -4.07, df = 10.3 *** (-) t = -4.66, df = 17.2 *** (-) Juncus articulatus t = -1.67, df = 26.3 ns t = -0.91, df = 26.1 ns Ludwigia palustris t = -0.26, df = 24.2 ns t = 1.04, df = 26.6 ns Luronium natans t = -5.32, df = 14.5 *** (-) t = -13.07, df = 24.6 *** (-) Mentha aquatica t = 0.39, df = 17.5 ns t = 3.71, df = 20.0 ** (+) Myriophyllum verticillatum t = -2.73, df = 23.6 * (-) t = -6.85, df = 23.5 *** (-) Nuphar lutea t = -3.31, df = 16.2 ** (-) t = -7.93, df = 18.0 *** (-) Oenanthe aquatica t = -0.50, df = 15.8 ns t = -3.40, df = 13.8 ** (-) Polygonum amphibium t = 4.37, df = 17.0 *** (+) t = 0.43, df = 23.9 ns Potamogeton coloratus t = -4.78, df = 16.6 *** (-) t = -12.25, df = 27.7 *** (-) Potamogeton gramineus t = -6.37, df = 17.7 *** (-) t = -14.79, df = 29.5 *** (-) Ranunculus flammula t = 2.02, df = 15.9 ns t = 5.21, df = 4.909 ** (+) Ranunculus peltatus t = -3.47, df = 16.8 ** (-) t = -6.36, df = 24.3 *** (-) Sagittaria sagittifolia t = 6.55, df = 15.5 *** (+) t = -0.74, df = 25.5 ns Sparganium emersum t = 4.51, df = 19.5 *** (+) t = -3.99, df = 27.1 *** (-)

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Table 3. Results of Ancova’s testing the significance of the differences in morphological traits between submerged and dewatered conditions for each species. Leaf dry matter Stem dry matter Petiole-blade Specific leaf Trait Leaf area Root/shoot ratio Leaf shape Dissection index content content ratio area Dependent log(total leaf log(root dry log(petiole log(blade log(total leaf log(leaf dry mass) log(stem dry mass) log(blade width) variable area) mass) length) perimeter) area) log(total dry log(stem fresh log(shoot dry log(leaf dry Species Covariate log(leaf fresh mass) log(blade length) log(blade length) log(blade length) mass) mass) mass) mass) Alisma plantago- Covariate (Cv) F1 = 378.6 *** F1 = 434.5 *** F1 = 2482.2 *** F1 = 85.0 *** F1 = 625.0 *** F1 = 6.27 * F1 = 1331.5 *** F1 = 258.8 *** aquatica 23 23 24 24 24 23 24 23 1 1 1 1 1 1 1 1 Condition (Cd) F 23= 49.5 *** (-) F 23= 11.3 ** (+) F 22= 0.72 ns F 22= 0.04 ns F 22= 1.77 ns F 23= 43.1 *** (-) F 22= 0.04 ns F 23= 9.90 ** (-) 1 1 1 1 1 1 1 1 Cv × Cd F 22= 1.46 ns F 22= 0.43 ns F 22= 1.16 ns F 22= 0.38 ns F 22= 0.15 ns F 22= 0.79 ns F 22= 0.99 ns F 22= 0.22 ns 1 1 1 1 1 1 1 1 Berula erecta Covariate (Cv) F 23= 753.6 *** F 23= 1441.4 *** F 23= 737.5 *** F 23= 359.0 *** F 23= 371.7 *** F 23= 4.93 * F 23= 664.7 *** F 23= 345.6 *** F1 = 102.8 *** (- F1 = 28.1 *** (- Condition (Cd) F1 = 28.5 *** (-) F1 = 52.2 *** (+) F1 = 27.9 *** (+) F1 = 45.5 *** (-) F1 = 14.9 *** (+) 23 F1 = 13.4 ** (-) 23 23 23 23 23 23 ) 23 ) 1 1 1 1 1 1 1 1 Cv × Cd F 22= 2.37 ns F 22= 3.02 ns F 22= 3.38 ns F 22= 0.51 ns F 22= 1.09 ns F 22= 1.25 ns F 22= 0.02 ns F 22= 2.79 ns 1 1 1 1 1 1 1 Elatine alsinastrum Covariate (Cv) F 28= 8.58 ** F 29= 66.1 *** F 28= 393.4 *** F 29= 18.2 *** F 28= 7.18. * F 29= 3285.7 *** F 28= 17.5 *** F1 = 20.6 *** (- F1 = 34.5 *** (- Condition (Cd) 28 F1 = 0.83 ns F1 = 38. *** (+) F1 = 2.12 ns F1 = 547.6 *** (+) NA F1 = 3.27 ns 28 ) 27 28 27 28 27 ) 1 1 1 1 1 1 1 Cv × Cd F 27= 0.59 ns F 27= 0.03 ns F 27= 0.32 ns F 27= 0.40 ns F 27= 0.44 ns F 27= 1.07 ns F 27= 3.48 ns 1 1 1 1 1 1 1 Huppuris vulgaris Covariate (Cv) F 21= 370.8 *** F 22= 235.9 *** F 22= 1124.4 *** F 21= 97.8 *** F 21= 32.4 *** F 22= 47689 *** F 21= 310.2 *** F1 = 34.3 *** (- Condition (Cd) F1 = 64.9 *** (-) F1 = 3.30 ns F1 = 3.86 ns F1 = 25.9 *** (+) F1 = 9.53 ** (+) NA F1 = 0.08 ns 21 21 20 20 21 21 20 ) 1 1 1 1 1 1 1 Cv × Cd F 20= 0.83 ns F 20= 3.21 ns F 20= 0.0002 ns F 20= 3.82 ns F 20= 0.13 ns F 20= 1.38 ns F 20= 0.07 ns 1 1 1 1 1 1 1 1 Hottonia palustris Covariate (Cv) F 20= 330.5 *** F 21= 428.3 *** F 21= 264.7 *** F 21= 38.6 *** F 20= 454 *** F 21= 21.2 *** F 20= 725.8 *** F 20= 1280.9 *** F1 = 77.6 *** (- Condition (Cd) F1 = 51.3 *** (-) F1 = 0.13 ns F1 = 0.03 ns F1 = 0.63 ns F1 = 12.6 ** (-) F1 = 3.03 ns F1 = 25.5 *** (-) 20 20 19 19 19 20 19 20 ) 1 1 1 1 1 1 1 1 Cv × Cd F 19= 3.58 ns F 19= 0.10 ns F 19= 1.39 ns F 19= 0.02 ns F 19= 0.30 ns F 19= 0.12 ns F 19= 0.43 ns F 19= 4.25 ns 1 1 1 1 1 1 1 Juncus articulatus Covariate (Cv) F 27= 114.2 *** F 27= 582.2 *** F 27= 177.1 *** F 28= 5.91 * F 28= 29.3 *** F 28= 3447.9*** F 27= 494.1*** F1 = 29.6 *** (- Condition (Cd) F1 = 33.6 *** (-) F1 = 19.8 *** (+) F1 = 19.2 *** (+) F1 = 0.01 ns F1 = 2.28 ns NA F1 = 0.02 ns 27 27 27 27 26 26 26 ) 1 1 1 1 1 1 1 Cv × Cd F 26= 0.01 ns F 26= 0.39 ns F 26= 0.17 ns F 26= 0.04 ns F 26= 0.15 ns F 26= 0.19 ns F 26= 1.34 ns 1 1 1 1 1 1 1 1 Ludwigia palustris Covariate (Cv) F 27= 292.5 *** F 27= 188.3 *** F 27= 326.2 *** F 28= 59.4 *** F 27= 148.2 *** F 27= 282.3 *** F 28= 685.7 *** F 27= 211.0 *** 1 1 1 1 1 1 1 1 Condition (Cd) F 27= 18.0 *** (-) F 27= 9.0 ** (+) F 27= 81.5 *** (+) F 26= 0.04 ns F 27= 43.4 *** (-) F 27= 69.1 *** (+) F 26= 4.05 ns F 27= 10.0 ** (-) 1 1 1 1 1 1 1 1 Cv × Cd F 26= 0.01 ns F 26= 0.20 ns F 26= 0.89 ns F 26= 0.69 ns F 26= 0.74 ns F 26= 0.0003 ns F 26= 0.36 ns F 26= 0.22 ns 1 1 1 1 1 1 1 1 Luronium natans Covariate (Cv) F 28= 864.8 *** F 28= 306.3 *** F 29= 331.6 *** F 28= 238.1 *** F 28= 22.9 *** F 28= 423.9 *** F 28= 206.2 *** F 28= 3499.3 *** F1 = 249.2 *** F1 = 704.6 *** Condition (Cd) 28 F1 = 9.53 ** (+) F1 = 0.68 ns F1 = 25.3 *** (+) F1 = 36.6 *** (+) F1 = 82.7 *** (-) F1 = 0.41 ns 28 (-) 28 27 28 28 28 27 (-) 1 1 1 1 1 1 1 1 Cv × Cd F 27= 1.36 ns F 27= 0.87 ns F 27= 0.33 ns F 27= 0.003 ns F 27= 0.03 ns F 27= 0.27 ns F 27= 2.36 ns F 27= 2.44 ns 1 1 1 1 1 1 1 1 Mentha aquatica Covariate (Cv) F 23= 42.8 *** F 23= 161.1 *** F 22= 702.2 *** F 23= 112.8 *** F 23= 101.6 *** F 22= 101. *** F 23= 1157.2 *** F 23= 100.8 *** 1 1 1 1 1 1 1 1 Condition (Cd) F 21= 1.18 ns F 21= 1.0 ns F 22= 22.8 *** (+) F 21= 2.25 ns F 21= 0.01 ns F 22= 22.0 *** (+) F 21= 3.34 ns F 21= 3.55 ns 1 1 1 1 1 1 1 1 Cv × Cd F 21= 0.01 ns F 21= 0.01 ns F 21= 0.20 ns F 21= 0.004 ns F 21= 0.002 ns F 21= 0.59 ns F 21= 0.61 ns F 21= 0.23 ns M.yriophyllum Covariate (Cv) F1 = 406.6 *** F1 = 243.1 *** F1 = 222.8 *** F1 = 64.1 *** F1 = 337.1 *** F1 = 21.4 *** F1 = 461.4 *** F1 = 106.1 *** verticillatum 27 28 27 27 27 27 27 27 F1 = 125.2 *** Condition (Cd) F1 = 37.0 *** (-) F1 = 3.39 ns F1 = 26.0 *** (+) F1 = 58.7 *** (+) F1 = 28.2 *** (-) F1 = 0.02 ns F1 = 18.8 *** (-) 27 27 26 27 27 27 26 27 (-)

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1 1 1 1 1 1 1 1 Cv × Cd F 26= 0.08 ns F 26= 0.24 ns F 26= 0.51 ns F 26= 1.44 ns F 26= 0.72 ns F 26= 0.03 ns F 26= 0.06 ns F 26= 0.21 ns 1 1 1 1 1 1 1 1 Nuphar lutea Covariate (Cv) F 17= 384.8 *** F 18= 1055.1 *** F 18= 781.1 *** F 17= 47.8 *** F 18= 530.1 *** F 17= 168.8 *** F 18= 237.3 *** F 17= 618.6 *** 1 1 1 1 1 1 1 1 Condition (Cd) F 17= 13.8 ** (-) F 16= 3.85 ns F 16= 0.39 ns F 17= 15.4 ** (+) F 16= 0.35 ns F 17= 21.8 *** (-) F 16= 0.51 ns F 17= 5.55 ** (-) 1 1 1 1 1 1 1 1 Cv × Cd F 16= 0.45 ns F 16= 0.14 ns F 16= 0.11 ns F 16= 0.02 ns F 16= 0.63 ns F 16= 1.52 ns F 16= 1.04 ns F 16= 0.19 ns 1 1 1 1 1 1 1 1 Oenanthe aquatica Covariate (Cv) F 15= 148.3 *** F 16= 239.4 *** F 16= 227.6 *** F 15= 52.1 *** F 16= 55.3 *** F 16= 128 *** F 15= 93.1 *** F 15= 50.8 *** 1 1 1 1 1 1 1 1 Condition (Cd) F 15= 16.5 ** (-) F 14= 3.89 ns F 14= 0.14 ns F 15= 13.8 ** (+) F 14= 0.003 ns F 14= 0.32 ns F 15= 9.53 ** (-) F 15= 7.06 * (-) 1 1 1 1 1 1 1 1 Cv × Cd F 14= 1.78 ns F 14= 0.07 ns F 14= 3.60 ns F 14= 0.09 ns F 14= 0.0002 ns F 14= 0.12 ns F 14= 0.94 ns F 14= 0.36 ns Polygonum Covariate (Cv) F1 = 386 *** F1 = 925.9 *** F1 = 320.1 *** F1 = 48.2 *** F1 = 146.9 *** F1 = 5.76 * F1 = 974.3 *** F1 = 534.7 *** amphibium 27 26 27 27 27 26 27 26 F1 = 1345.1 *** F1 = 24.3 *** (- Condition (Cd) F1 = 0.003 ns F1 = 30.8 *** (+) F1 = 2.33 ns F1 = 2.45 ns F1 = 2.38 ns 26 F1 = 2.05 ns 26 25 26 25 25 25 (-) 25 ) 1 1 1 1 1 1 1 1 Cv × Cd F 25= 1.77 ns F 25= 2.45 ns F 25= 1.16 ns F 25= 1.33 ns F 25= 0.11 ns F 25= 3.12 ns F 25= 0.66 ns F 25= 3.36 ns Potamogeton Covariate (Cv) F1 = 521.1 *** F1 = 1074.7 *** F1 = 147.8 *** F1 = 7.37 *** F1 = 48.1 *** F1 = 200.1 *** F1 = 11983 *** F1 = 622.9 *** coloratus 27 28 28 27 27 28 28 27 1 1 1 1 1 1 1 1 Condition (Cd) F 27= 8.63 ** (-) F 26= 0.06 ns F 26= 0.61 ns F 27= 9.27 ** (+) F 27= 15.5 *** (+) F 26= 4.0 ns F 26= 4.21 ns F 27= 10.9 ** (-) 1 1 1 1 1 1 1 1 Cv × Cd F 26= 0.01 ns F 26= 2.53 ns F 26= 0.001 ns F 26= 0.04 ns F 26= 0.005 ns F 26= 0.0006 ns F 26= 0.90 ns F 26= 0.70 ns Potamogeton Covariate (Cv) F1 = 1534.7 *** F1 = 360.4 *** F1 = 194.9 *** F1 = 41.6 *** F1 = 229.5 *** F1 = 8.54 ** F1 = 343.3 *** F1 = 1508.0 *** gramineus 29 29 30 30 29 29 30 29 F1 = 526.5 *** (- F1 = 125.6 *** Condition (Cd) F1 = 49.7 *** (-) F1 = 6.0 * (+) F1 = 0.82 ns F1 = 0.45 ns F1 = 55.7 *** (+) 29 F1 = 4.0 ns 29 29 29 28 28 29 ) 28 (-) 1 1 1 1 1 1 1 1 Cv × Cd F 28= 1.62 ns F 28= 0.96 ns F 28= 1.08 ns F 28= 0.06 ns F 28= 0.005 ns F 28= 0.92 ns F 28= 0.04 ns F 28= 0.42 ns Ranunculus Covariate (Cv) F1 = 83.8 *** F1 = 151.6 *** F1 = 171.3 *** F1 = 5.56 * F1 = 11.5 ** F1 = 5.74 * F1 = 1890.9 *** F1 = 115.8 *** flammula 15 15 16 15 16 15 16 15 F1 = 104.8 *** (- Condition (Cd) F1 = 31.2 *** (-) F1 = 9.1 ** (+) F1 = 4.19 ns F1 = 29.8 *** (-) F1 = 0.30 ns 15 F1 = 3.25 ns F1 = 12.1 ** (-) 15 15 14 15 14 ) 14 15 1 1 1 1 1 1 1 1 Cv × Cd F 14= 0.14 ns F 14= 0.001 ns F 14= 0.001 ns F 14= 0.15 ns F 14= 0.10 ns F 14= 0.85 ns F 14= 0.13 ns F 14= 1.99 ns Ranunculus Covariate (Cv) F1 = 145.8 *** F1 = 292.6 *** F1 = 1208.9 *** F1 = 30.1 *** F1 = 202.8 *** F1 = 18.9 *** F1 = 364.9 *** F1 = 297.3 *** peltatus 24 25 25 24 25 25 25 25 F1 = 206.9 *** F1 = 29.4 *** (- Condition (Cd) F1 = 15.5 *** (-) F1 = 2.65 ns F1 = 3.65 ns F1 = 11.3 ** (+) F1 = 0.02 ns 24 F1 = 34.4 *** (-) 24 24 23 23 24 23 (+) 24 ) 1 1 1 1 1 1 1 1 Cv × Cd F 23= 0.5 ns F 23= 0.69 ns F 23= 1.67 ns F 23= 0.02 ns F 23= 0.75 ns F 23= 0.42 ns F 23= 1.76 ns F 23= 3.59 ns Sagittaria F1 = 11991.9 Covariate (Cv) F1 = 597.4 *** F1 = 2572.2 *** F1 = 2767.1 *** F1 = 343.9 *** F1 = 111 *** F1 = 104.8 *** 26 F1 = 183.3 *** sagittifolia 26 26 27 27 27 26 *** 26 F1 = 160.2 *** F1 = 74.5 *** (- Condition (Cd) 26 F1 = 19.9 *** (+) F1 = 1.99 ns F1 = 0.67 ns F1 = 3.0 ns F1 = 98.5 *** (-) F1 = 41.6 *** (+) 26 (-) 26 25 25 25 26 26 ) 1 1 1 1 1 1 1 1 Cv × Cd F 25= 4.18 ns F 25= 3.74 ns F 25= 3.84 ns F 25= 3.55 ns F 25= 0.02 ns F 25= 0.08 ns F 25= 2.36 ns F 25= 3.39 ns Sparganium Covariate (Cv) F1 = 227.8 *** F1 = 543.2 *** F1 = 333.4 *** F1 = 61.1 *** F1 = 28.8 *** F1 = 1143.7 *** F1 = 180.8 *** emersum 27 27 28 28 27 28 27 1 1 1 1 1 1 1 Condition (Cd) F 27= 48.3 *** (-) F 27= 54.7 *** (+) F 26= 0.23 ns F 26= 1.90 ns F 27= 70.3 *** (+) NA F 26= 0.70 ns F 27= 57 *** (-) 1 1 1 1 1 1 1 Cv × Cd F 26= 2.79 ns F 26= 2.04 ns F 26= 2.46 ns F 26= 0.005 ns F 26= 0.004 ns F 26= 0.86 ns F 26= 0.06 ns The Fdf values and significance levels of the effects (growing habitat-submerged vs dewatered, covariate and their interaction) are presented. For leaf area, total leaf area was used as dependent variable and total dry mass as covariate. For leaf and stem dry-matter content, leaf and stem dry mass were used as dependent variables, and leaf and stem fresh mass as covariate respectively. For root/shoot ratio, root dry mass was used as dependent variable and shoot dry mass as covariate. For leaf shape blade width was used as dependent variable and blade length as covariate. For petiole blade ratio, petiole length was used as dependent variable and blade lengt as covariate. For dissection index, blade perimeter was used as dependent variable and blade length as covariate. For specific leaf area, total leaf area was used as dependent variable and leaf dry mass as covariate. Non-significant interaction terms were excluded from the final model and are indicated in italics; the F values and significance levels for these terms correspond to the full model in which all terms were present. Significance: *** p < 0.001; ** p < 0.01; * p < 0.05; ns, not significant. NA, data not available. 113

RESULTATS Comparaison des phénotypes inondés et exondés

Model selection

The best fitting models for the variation of performance (total dry mass and plant height) and root/shoot ratio between submerged and dewatered conditions, are the two-peak OUMamp

(Table 4) which infer selection toward an optimum in amphiphyte and hydrophyte species.

These models both conclude that hydrophyte total dry mass and plant height are significantly lower in dewatered conditions compared to submerged ones. Total dry mass and plant height were both higher in dewatered conditions compared to submerged ones for amphiphyte species.

For the root/shoot ratio, the model concludes respectively to a higher root/shoot for hydrophytes, and a lower for amphiphytes, in dewatered conditions compared to submerged ones.

The trait variations between dewatered and submerged conditions for LDMC, petiole-blade ratio and leaf area/dry mass were best described by the two-peak OUros model (Table 4). For caulescent species, the model concluded to a higher LDMC and petiole-blade ratio, and a lower leaf area/total dry mass in dewatered conditions compared to submerged ones. For rosette species, the model concluded to a higher LDMC and a lower petiole-blade ratio and leaf area/total dry mass in dewatered conditions compared to submerged ones.

The best supported model for the dissection index, SDMC and SLA trait variation is the single- peak OU model (OU1), which concluded a lower value of dissection index, a higher SDMC and a lower SLA in dewatered conditions compared to submerged ones (Table 4).

Leaf shape variation was the best described by the single-rate Brownian model (BM1), corresponding to a phylogenetic signal i.e. related species show relatively similar values (Table

4).

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Table 4. Parameter estimates for model fit to phenotypic traits.

Traits Model Ancestral Brownian OU strenght hydrophytes Brownian amphipytes Brownian rosette Brownian caulescent Brownian AICc state rate of selection rate or optimum rate or optimum rate or optimum rate or optimum Dissection BM1 0.111 0.022 0.002±0 index BMMamp 0.149 0.032 0.013 0.002±0 BMMros 0.158 0.018 0.027 0.001±0 OU1 -0.064 1.637 0.741 0.379±0.003 OUMamp 12.810 6.407 -0.557 0.286 0.253±0.004 OUMros 5.616 2.969 0.468 -0.441 0.363±0.003 LDMC BM1 0.773 0.016 0±0 BMMamp 0.840 0.026 0.007 0.002±0.001 BMMros 0.939 0.005 0.025 0.002±0 OU1 0.598 1.411 0.904 0.12±0.001 OUMamp 12.719 9.799 0.091 0.959 0.207±0.005 OUMros 6.741 6.102 1.169 0.175 0.669±0.04 SDMC BM1 0.313 0.009 0.072±0.001 BMMamp 0.309 0.07 0.011 0.035±0.001 BMMros 0.307 0.009 0.010 0.022±0 OU1 0.289 0.034 0.021 0.362±0.004 OUMamp 0.049 0.034 -0.037 0.516 0.163±0.005 OUMros 2.254 1766 -0.251 0.617 0.346±0.005 Leaf shape BM1 0.400 0.014 0.298±0.003 BMMamp 0.399 0.015 0.016 0.208±0.006 BMMros 0.434 0.012 0.017 0.117±0.004 OU1 0.244 0.046 0.014 0.216±0.003 OUMamp 4.591 1.461 0.754 -0.087 0.074±0.003 OUMros 6.306 2.045 0.826 -0.114 0.088±0.033 Petiole- BM1 -0.553 0.048 0±0 blade ratio BMMamp -0.574 0.066 0.032 0±0 BMMros -0.616 0.039 0.058 0±0 OU1 -0.475 3.706 1.037 0.079±0.001 OUMamp 14.817 4.201 -0.247 -0.633 0.021±0.001 OUMros 13.326 5.686 -1.417 0.217 0.9±0.001 Root/shoot BM1 0.506 0.035 0±0 ratio

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BMMamp 0.525 0.028 0.040 0±0 BMMros 0.495 0.059 0.020 0±0 OU1 0.338 2.149 0.910 0.24±0.003 OUMamp 26.884 14.972 1.020 -0.157 0.707±0.003 OUMros 6.679 2.857 0.219 0.426 0.053±0.001 Leaf area/ BM1 -1.283 0.016 0±0 total dry mass BMMamp -1.343 0.026 0.001 0.005±0.001 BMMros -1.541 0.000 0.028 0.051±0.005 OU1 -1.226 0.699 0.981 0.066±0 OUMamp 4.721 6.915 -1.055 -1.351 0.024±0.001 OUMros 2.793 6.075 -1.646 -0.918 0.854±0.004 Total dry BM1 -0.275 0.022 0±0 mass BMM -0.400 0.010 0.031 0±0 BMMros -0.393 0.042 0.007 0±0 OU1 -0.124 0.235 0.101 0.001±0 OUMamp 7.382 12.135 -1.260 0.676 0.999±0.001 OUMros 6.340 3.294 0.417 -0.543 0.001±0 Plant BM1 -0.720 0.033 0±0 height BMMamp -0.896 0.007 0.048 0.001±0 BMMros -0.637 0.067 0.009 0.001±0 OU1 -0.478 0.665 0.229 0.027±0.001 OUMamp 16.478 9.951 -1.481 0.226 0.964±0.001 OUMros 0.996 0.344 -0.697 -0.313 0.007±0 Specific BM1 -0.891 0.019 0±0 leaf area BMMamp -0.891 0.025 0.002 0.002±0 BMMros -0.911 0.001 0.026 0.002±0 OU1 -0.972 0.727 1.030 0.68±0.001 OUMamp 0.491 0.698 -0.962 -0.979 0.142±0.001 OUMros 2.701 3.918 -1.089 -0.885 0.175±0.001

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Phenotypic traits and performances

GLS regression between morphological traits and performance (total dry mass) variations was only significant for root/shoot ratio (t-value = -2.342; p= 0,034).

Discussion

Several morpho-anatomical traits showed common variation in response to dewatering. The most commonly observed changes were a higher above-ground dry matter content (leaf and stem) and a lower leaf area/ total dry mass and specific leaf area in dewatered conditions compared to submerged ones. These observations were consistent with what has been observed in different studies focused on a small set of species (Bruni et al. 1996 ; Vasellati et al. 2001;

Geng et al. 2006 ; Iida et al. 2007 ; Li et al. 2011; Hamann & Puijalon 2013). The lower leaf area permits to reduce water loss through evapotranspiration in terrestrial conditions. The higher dry matter content of above-ground organs suggests a lower proportion of aerenchyma and a greater investment in strengthening tissues, leading to self-supporting aboveground organs able to withstand gravity (Niklas 1998 ; Vasellati et al. 2001; Jackson & Colmer 2005 ;

Mommer &Visser 2005 ; Li et al. 2011 ; Hamann & Puijalon 2013). Self-supporting permits a better light harvesting for photosynthesis. The lower leaf area and higher leaf dry matter content may lead to a decrease of specific leaf area (Iida et al. 2007), and a increasing construction cost for dewatered leaves.

Contrary to what we expected, morphological changes were not more similar for closer taxonomic groups excepted for leaf shape. Leaf shape showed a phylogenetic signal, which corresponds to an association between phylogeny and pattern of trait changes. Phylogenetic signal may be due to either phylogenetic inertia (i.e. the tendency to resist a current adaptive force) or convergent adaptation in related taxa (Hansen & Orzack 2005). As the best models

117

RESULTATS Comparaison des phénotypes inondés et exondés where not those that retained the adaptive hypothesis (BMMros and BMMamp models), leaf shape variation between habitats was probably constrained by species evolutionary history

(Orzack & Sober 2001).

In terms of performance (total dry mass and plant height), the selective pressure differed according to species eecological niches. Amphiphytes produced a performant phenotype when dewatered, whereas hydrophytes size and dry mass decreased in dewatered conditions compared to submerged ones. These results seem intuitive however, in view of the evolutionary history of aquatic angiosperms, phylogenetic conservatism was expected which could have resulted in phylogenetic signal (BM1 model) or phylogenetic signal within each niche

(BMMamp model) because species ecological niche depends on the species adaptations to its present habitat, but also on the legacy from its ancestors (Prinzing et al. 2001).

The Root/shoot ratio variation also evolved as a consequence of selection associated with different ecological niches: the ratio was higher in dewatered conditions for hydrophytes, while it was lower for amphiphytes. A high investment in roots compared to shoots is usually considered as an adaptation to stressful conditions (Grime 2002). In the present case, it may relate to water or nutriments deficiency, because of the inability for aquatic plants to collect nutrients in water. In such situation, plants may invest more in belowground organs for optimizing water and nutrient uptake. Such increasing investment in the belowground organs may also result from higher storage, possibly due to increasing translocation of resources from decaying leaves and stems, and may in this case favors the regeneration of above ground organs when the water level increases again. Amphiphytes may collect nutrients mostly in the substrate under submerged conditions, and they organs may better tolerate dewatering, with the consequence of few supplementary investment in the below ground organs when dewatered.

As expected, the variation of some morpho-anatomical traits was constrained by growth form. The dewatered phenotype of rosette species was, in accordance with our hypothesis,

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RESULTATS Comparaison des phénotypes inondés et exondés mainly based on the production of self-supporting leaves with greater investment in strengthening tissues, shorter petiole and smaller leaf are to withstand gravity (Golibert 1989 ;

Usherwood et al. 1997 ; Iida et al. 2007; Hamann & Puijalon 2013). Higher stem dry matter content in terrestrial conditions was expected for caulescent species, but the best model (OU1 model) suggested few contrasts between growth forms for SDMC variations. Contrary to what we expected, the variation of performance traits was not related to species growth form.

Finally, contrary to expectations, no relationship between morphological trait and performance variation was observed.

Interestingly, SLA variation remained poorly differed among species. This may result from stabilizing selection toward a strategy of resource economy in response to dewatering

(Wright et al. 2004).

Conclusion

This study is a first attempt to investigate the morphological variation in response to dewatering of a large number of aquatic Angiosperm species in an evolutionary context. This work showed that growth form drives the changes in morphological traits, and the niche was related to the variation of performance.

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Supporting information

Table S1. Species sampled, additional species and species of the outgroup (phylogeny and GenBank accession number) used for construction of the phylogenetic tree. Indicated phylogeny is based on the Angiosperm Phylogeny Group III (APG III, 2009). Bold type GenBank number indicated newly sequenced genes. NA: data not available. Species Order Familly GenBank number rbcL gene matK gene Species sampled Alisma plantago-aquatica L. Alismatales Alismataceae L08759 JF781065 Berula erecta (Huds.) Coville Apiales Apiaceae AM234813 JN895702 Elatine alsinastrum L. Malpighiales Elatinaceae will be published will be published Hottonia palustris L. Ericales Primulaceae AF395002 AY647534 Hippuris vulgaris L. Lamiales Plantaginaceae AF248028 JN895316 Juncus articulatus L. Poales Juncaceae AY395543 JN894759 Luronium natans (L.) Raf. Alismatales Alismataceae U80680 JN894192 Ludwigia palustris (L.) Elliott Myrtales Onagraceae AM235670 will be published Mentha aquatica L. Lamiales Lamiaceae GU344680 HM850796 Myriophyllum verticillatum L. Saxifragales Haloragaceae GU344678 EF178983 Nuphar lutea (L.) Sm. Nympheales Nympheaceae DQ182338 AF117100 Oenanthe aquatica (L.) Poir. Apiales Apiaceae will be published JN893874 Polygonum amphibium (L.) Delarbre Caryophyllales Polygonaceae FM883621 KC342453 Potamogeton gramineus L. Alismatales Potamogetonaceae AB196943 will be published Potamogeton coloratus Hornem. Alismatales Potamogetonaceae GU344673 JN895690 Ranunculus flammula L. Ranunculales Ranunculaceae HM850295 AY954204 Ranunculus peltatus Schrank Ranunculales Ranunculaceae will be published JN894996 Sparganium emersum Rehmann Alismatales Alismataceae GU344672 JN894086 Sagittaria sagittifolia L. Poales Typhaceae GU344676 JN895790 Outgroup Cycas revolute Thunb. JQ512537 JQ512413 Ginkgo biloba L. JQ512540 AF456370 Picea abies (L.) H. Karst AJ001004 AY289610 Additional species Alisma lanceolatum With. Alismatales Alismataceae JN890989 HM850585 Alisma gramineum Lej. Alismatales Alismataceae JF781041 JF781067 Amborella trichopoda Baill Amborellales Amborellaceae L12628 DQ185522 Apium nodiflorum (L.) Lag. Apiales Apiaceae JN893767 JN895216 Baldellia ranunculoides (L.) Parl. Alismatales Alismataceae DQ859163 HQ456458 Callitriche platycarpa Kutz. Lamiales Plantaginaceae AF248022 JN896071

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RESULTATS Comparaison des phénotypes inondés et exondés

Elatine hexandra (Lapierre) DC. Malpighiales Elatinaceae JN892810 HM850928 Elatine triandra Schkuhr. Malpighiales Elatinaceae AY380349 EF135532 Elodea Canadensis Michx. Malpighiales Elatinaceae DQ859167 NA Elodea nuttallii (Planch.) H. St. John. Alismatales Hydrocharitaceae AB004888 AB002568 Galium palustre L. Gentianales Rubiaceae X81101 JN893881 Glyceria fluitans (L.) R. Br. Poales Poaceae AJ746290 HM850538 Groenlandia densa (L.) Fourr. Alismatales Potamogetonaceae AB196954 JN894619 Hydrocotyle vulgaris L. Apiales Araliaceae DQ133813 JN895453 Juncus effuses L. Poales Juncaceae AY216612 HQ180871 Juncus subnodulosus Schrank Poales Juncaceae AY216630 JN896173 Myosotis scorpioides L. Boraginaceae GU344681 JN895929 Myriophyllum spicatum L. Saxifragales Haloragaceae GU344679 EF178976 Nasturtium officinale R.Br. Brassicales Brassicaceae HM850197 JN894326 Nymphea alba L. Nympheales Nympheaceae JN892012 JN894938 Polygonum hydropiper (L.) Spach Caryophyllales Polygonaceae AB008781 HM357924 Potamogeton natans L. Alismatales Potamogetonaceae AB196946 JN895956 Potamogeton nodosus Poir. Alismatales Potamogetonaceae GU344674 HM851047 Potamogeton pusillus L. Alismatales Potamogetonaceae HM850283 HM851049 Ranunculus fluitans Lam. Ranunculales Ranunculaceae JN891886 AY954129 Ranunculus ophioglossifolius Vill. Ranunculales Ranunculaceae NA AY954207 Ranunculus trichophyllus Chaix. Ranunculales Ranunculaceae L08766 AY954133 Samolus valerandi L. Ericales Primulaceae U96659 JN895196 Sparganium natans L. Poales Typhaceae JN966012 JN966687 Veronica anagallis-aquatica L. Lamiales Plantaginaceae AY034021 JN894700 Veronica beccabunga L. Lamiales Plantaginaceae JN891845 JN894414

References

Angiosperm Phylogeny Group III (2009) An update of the angiosperm phylogeny group classification for the orders and families of flowering plants: APG III. Bot J Linn Soc 161:105–121

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RESULTATS Comparaison des phénotypes inondés et exondés

Appendix: 1) Defining the diversification rate for the Yule process Defining the expected diversification rate ߣ prior distribution range conditioned on tree height (e.g., between 306 and 366 Ma – Clarke et al. 2011) and species number under a Yule process: σ௡ିଵ ݊Ȁ݅ ߣൌ ௜ୀଶ ൈݐ݊

Where n is the number of species in the phylogeny, and t is the expected age of the root node.

2) Fossil calibrations for the phylogenetic tree.

Stem/crown Age min* Age max* Clade References Exp. mean Crown 306,2 366 root (1) 16 Crown 306,2 366 outgroup (1) 16 Crown 160 306,2 Gingko-Picea (1) 55,5 Crown 98,7 248 Nympheales + Amborella (2, 3) 40,5 Crown 51 115 Nymphaea-Nuphar (4) 53,5 Crown 124 248 Mesioangiospermae (1) 34 Stem 112 248 Poales (5, 6) 37 Stem 54,5 248 Hydrocharitales (7) 52,5 Crown 124 248 Angiospermae (1) 34 Crown 22,5 68,1 Alismatales** (8) 12 Crown 92,7 248 Monocotyledon (1, 9) 42 Crown 124 248 Eudicotyledon (1) 34 Stem 89,3 248 Saxifragales (10) 43 Crown 48 248 Malpighiales (11) 54 Crown 45,15 92 Apiales (12) 12 Stem 44,3 92 Lamiales (13) 13 Stem 70 248 Sparganium (14) 48 *Age in Ma **clade Alisma-APL-Baldelia/LPA

Starting value for the birth rate parameter of the Yule process: 0.0112 (for a mean root age of 315 ma, see Appendix A) The stem/crown defines whether the minimum bound represent the age of a stem member for the basal most branch of the clade. References : 1. J. T. Clarke, R. C. M. Warnock, P. C. J. Donoghue, Establishing a time-scale for plant evolution. New Phytol. 192, 266–301 (2011). 2. D. Winship Taylor, G. J. Brenner, S. H. Basha, Scutifolium jordanicum gen. et sp. nov. (Carbombaceae), an aquatic fossil plant from the Lower Cretaceous of Jordan, and the relationships of related leaf fossils to living genera. Am. J. Bot. 95, 340–352 (2008). 3. E. M. Friis, K. R. Pedersen, P. R. Crane, Fossil evidence of water lilies (Nymphaeales) in the Early Cretaceous. Nature. 410, 357–360.

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RESULTATS Comparaison des phénotypes inondés et exondés

4. I. Chen, S. R. Manchester, Z. Chen, Anatomically preserved seeds of Nuphar (Nymphaeaceae) from the Early Eocene of Wutu, Shandong Province, China. Am. J. Bot. 91, 1265–1272 (2004). 5. J. Muller, Fossil pollen records of extant angiosperms. Bot. Rev. 47, 1–142 (1981). 6. C. P. Daghlian, A review of the fossil record of Monocotyledons. Bot. Rev. 47, 517–555 (1981). 7. N. P. Sille, M. E. Collinson, M. Kucera, J. J. Hooker, Morphological Evolution of Stratiotes through the Paleogene in England: An Example of Microevolution in Flowering Plants. PALAIOS. 21, 272–288 (2006). 8. L-Y Chen, J-M Chen,R.W Gituru, T D Temame,Q-F Wanga, Generic phylogeny and historical biogeography of Alismataceae, inferred from multiple DNA sequences Mol Phylogenet Evol 63, 407–416 (2012) 9. E. M. Friis, K. R. Pedersen, P. R. Crane, Diversity in obscurity: fossil flowers and the early history of angiosperms. Philos. Trans. R. Soc. B. 365, 369–382 (2010). 10. E. J. Hermsen, M. A. Gandolfo, K. C. Nixon, W. L. Crepet, Divisestylus gen. nov. (aff. Iteaceae), a fossil saxifrage from the Late Cretaceous of New Jersey, USA. Am. J. Bot. 90, 1373–1388 (2003). 11. L. D. Boucher, S. R. Manchester, W. S. Judd, An extinct genus of Salicaceae based on twigs with attached flowers, fruits, and foliage from the Eocene Green River Formation of Utah and Colorado, USA. Am. J. Bot. 90, 1389–1399 (2003). 12. S Magallón, M.J Sanderson, Absolute diversification rates in angiosperm clades. Evolution 55, 1762– 1780 (2001) 13. V. B. Call, D. L. Dilcher, Investigations of angiosperms from the Eocene of southwestern North America: samaras of Fraxinus wilcoxiona berry. Rev. Palaeobot. Palynol. 74, 249–266 (1992). 14. J. D. Sulman, B. T. Drew, C. Drummond, E. Hayasaka, K. J. Sytsma, Systematics, biogeography, and character evolution of Sparganium (Typhaceae): Diversification of a widespread, aquatic lineage. Am. J. Bot. 100, 2023–2039 (2013).

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RESULTATS Comparaison des phénotypes inondés et exondés

Supplementary material

Fig S1 Maximum Clade Credibilility (MCC) tree from the BEAST analysis for the 49-taxa of aquatic angiosperm species and related possible root species , combined 2-gene data set. 95% highest posterior density (HPD) intervals for node ages represented by error bars. Node values represent the posterior probability. In red : species considered in this study, APL Alisma plantago-aquatica BER Berula erecta EAL Elatine alsinastrum HVU Hippuris vulgaris HPA Hottonia palustris JAR Juncus articulatus LPA Ludwigia palustris LNA Luronium natans MAQ Mentha aquatica MVE Myriophyllum verticillatum NLU Nuphar lutea OAQ Oenanthe aquatic PAM Polygonum amphibium PCO Potamogeton coloratus PGR Potamogeton gramineus RFL Ranunculus falmmula RPE Ranunculus peltatus SSA Sagattaria sagittifolia SEM Sparganium emersum.

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RESULTATS Réponses à court terme des communautés végétales aquatiques

3. Réponses à court terme des communautés végétales aquatiques.

Mélissa De Wilde, Sara Puijalon, et Gudrun Bornette. Resistance and resilience of aquatic plant communities to dewatering: the role of sediment type (soumis dans Journal of Vegetation

Science).

Problématique

L’objectif de cet article est de mesurer comment différents types de sédiment, via leur capacité différentielle à retenir l’eau, gouvernent la réponse des communautés végétales aquatiques à l’exondation.

Les hypothèses testées sont :

1) Plus les sédiments sont grossiers et pauvres en matière organique et moins ils retiennent

l’eau, plus l’intensité de l’exondation pour les plantes aquatiques est intense menant à

une diminution de la résistance et la résilience des communautés,

2) La réponse des plantes le long de ce gradient d’intensité diffère selon leur niche

écologique. La résistance des hydrophytes établies est la plus élevée sur les sédiments

organiques, la plus faible sur les sédiments grossier, et intermédiaire sur les sédiments

limoneux. Durant la phase d’exondation, les zones où les hydrophytes ne survivent pas

sont colonisées par les amphiphytes et les hélophytes, conduisant à une abondance plus

élevée des amphiphytes et hélophytes sur les sédiments grossiers durant la phase

d’exondation et une abondance intermédiaire sur les sédiments limoneux. Après la

remise en eau, la résilience des hydrophytes est la plus faible sur les sédiments grossiers

du fait des condition défavorables pour la survie des propagules végétatives, et leur

résilience est la plus élevée sur les sédiments organiques, les sédiments limoneux

montrent un patron intermédiaire.

3) La relation entre richesse spécifique et intensité de l’exondation prend la forme d’une

courbe en cloche en conformité avec la théorie des perturbations intermédiaires.

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RESULTATS Réponses à court terme des communautés végétales aquatiques

Pour tester ces hypothèses, des communautés végétales colonisant 3 types de sédiment

(graviers, limons, vase organique), en eau permanente et subissant simultanément une exondation estivale, puis une remise en eau, ont été étudiées.

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RESULTATS Réponses à court terme des communautés végétales aquatiques

Resistance and resilience of aquatic plant communities to dewatering: the role of sediment type.

Mélissa De Wilde*1, Sara Puijalon1, Gudrun Bornette2

1 UMR CNRS 5023 « Laboratoire d’Ecologie des Hydrosystèmes Naturels et Anthropisés »; Université de Lyon;

Université Lyon 1; ENTPE; 43 boulevard du 11 novembre 1918, 69622 Villeurbanne Cedex, France; 2 UMR

CNRS 6259 Chronoenvironnement,Université de Franche Comté, Campus de la Bouloie, 16, route de Gray, 25000

Besançon Cedex

Abstract

Questions: How different types of sediment, through their differential ability to retain water, govern the short term response of aquatic plant communities to dewatering?

Location: Riverine wetlands located in the floodplain of the Ain River, France

Methods: 18 sampling units were delimited, dispatched on 3 sediment types: gravel-dominated coarse sediment, silt, and organic-matter dominated sediment. For each sediment type, 3 sampling units were permanently aquatic (reference units), and 3 underwent a summer dewatering. Vegetation was surveyed in each sampling unit at 4 dates: before the summer dewatering, at the beginning of the event, at the end of the event, and 2 months after rewetting.

The resistance and resilience of communities were assessed, as well as the changes through time in the proportion of the different ecological groups (helophytes, hydrophytes, and amphiphytes), and the effect of dewatering on species renewal and richness.

Results: For the same dewatering event, the nature of sediment (from organic, to silty and then coarse sediment) was significantly related to a decrease of aquatic plant community resistance and resilience, suggesting increasing disturbance intensity. Organic sediment may retain more efficiently water during dewatering, allowing the maintenance of amphiphytes and of the more tolerant hydrophytes, and consequently high community resistance. On silty sediment, the

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RESULTATS Réponses à court terme des communautés végétales aquatiques disturbance was high enough to lead to the disappearance of vegetative parts of aquatic plants, but seeds and propagules may be preserved in the substrate and rapidly sprout when rewetting, allowing a good resilience of communities. On coarse sediment, vegetation changes corresponded to a decrease of the relative abundance of resident amphiphyte species, together with helophyte colonization, and their maintenance after rewetting. Coarse sediment is probably poorly favorable to seed and propagule survival in case of complete dewatering, explaining the low resilience of communities. After dewatering, a linear positive relationship between dewatering intensity and species richness was observed, whereas an hump-shaped relationship was expected according to the intermediate disturbance hypothesis.

Conclusion: Our study demonstrated that a rather simple description of sediment type allows predicting the impact of dewatering on plant communities.

Keywords: Aquatic plant community, Dewatering, Disturbance intensity, Resistance,

Resilience, Sediment type

Nomenclature: Lambinon et al. 2004; Lauber & Wagner 2007

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Introduction

Wetlands have been recognized as key ecosystems for biodiversity and ecosystem services, such as water purification, carbon storage, flood control and regulation of nutrient cycles

(RAMSAR 2002; Millennium Ecosystem Assessment 2005; Keddy et al. 2009). Apart from their contribution to wetland biodiversity, plant communities are strongly involved in several services, as CO2 capture, food and habitat providing for other organisms, and resource supplies, and they influence hydrology, sediment dynamics and biogeochemical cycles (Costanza et al.

1997; Engelhardt & Ritchie 2001, 2002).

For several decades, wetlands are threatened by hydrological and chemical alterations induced by human activities. Among them, the increasing use of water resources, together with climate change, lead to periodic or chronic water deficit, and consequently large seasonal variations in water level (Arnell 1999; Lehner et al. 2006). In this context, many temperate wetlands undergo more frequent and more intense summer dewatering (Winter 2000; Erwin

2009; Junk et al. 2013).

Hydrology is a major determinant of species composition and successional dynamics of aquatic plant communities because the hydroperiod directly affects recruitment, growth and survival of aquatic plant species (Van der Valk 2005; Van Geest et al. 2005 a, b). Dewatering, which corresponds to the disappearance of the water column, places sediment and organisms in new environmental conditions. Dewatering has marked effects on aquatic plant communities

(Van der Valk, 2005, Lake 2003) and consequently ecosystem functioning (Cavanaugh et al.

2006), especially in temperate, previously permanently aquatic wetlands.

The resistance of a plant community represents its ability to tolerate dewatering or other disturbing factors without showing significant changes in its structure and composition (Lake

2000). The short-term resistance of the community may rely on species ability to withstand dewatering through phenotypic adjustments (Sculthorpe 1967; Robe & Griffiths 1998). When

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RESULTATS Réponses à court terme des communautés végétales aquatiques the disturbance intensity exceeds a certain threshold, the resistance of established submerged aquatic species may be null, and in this case, the community maintenance relies on its resilience

(i.e. the capacity to recover from disturbance, Lake 2000), enabled by the short-term recolonization from propagule banks or less impacted areas (Barrat-Segretain & Amoros 1996;

Brock et al. 2003; Combroux & Bornette 2004).

Responses of aquatic plant communities to dewatering/rewetting episodes may differ depending on habitat conditions, and more particularly on sediment properties. Indeed, particle size and sediment content in organic matter both influence sediment water retention (Walczak et al. 2002; Rawls et al. 2003). The sediment rich in organic matter has a high ability to retain water during dewatering, as organic matter acts like a sponge. Because of their high bulk density, silty and clay sediments have also high water retention capacity compared to coarse sediments with large gravel-size particles and low organic matter content (Walczak et al. 2002;

Saxton & Rawls 2006). Consequently, for the same duration of dewatering, the intensity of dewatering may differ according to sediment type, therefore differently influencing plant survival, and propagule conservation and germination.

Several studies described vegetation changes caused by water-level oscillations and some of them took into account the range of water level change, or the duration of dewatering period (Van Geest et al. 2005 a, b; Wilcox & Nichols 2008). However, only few have considered habitat conditions as sediment type. The aim of the present study was consequently to assess how sediment type rules the response of aquatic plant communities to dewatering. For this purpose, we compared in situ the response to summer dewatering of natural wetland plant communities growing on 3 different sediment types (coarse, silty and organic sediment).

We hypothesized firstly that the coarser and poorer in organic matter is the sediment, the higher is the intensity of dewatering for aquatic plants, leading to decreasing resistance and

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RESULTATS Réponses à court terme des communautés végétales aquatiques resilience of plant communities. In this framework, we expected a positive relationship between resilience and resistance.

As the niche of plant species is known to differ in terms of water availability, different responses were expected for hydrophytes (strictly aquatic plants in their vegetative phase), amphiphytes (plants able to tolerate emersion if the sediment remain wet) and helophytes

(emerged plants tolerating partial sediment drying). Consequently, we hypothesized that the resistance of established hydrophyte species should be the highest on organic sediment, the lowest on the coarse one, and intermediate on the silty sediment. We then expected that during the dewatering phase, the areas where hydrophytes decayed would be colonized by amphiphytes and helophytes, leading to a higher abundance of amphiphytes and helophytes on the coarser sediment during the dewatering phase, and an intermediate abundance on the silty sediment. After rewetting, we expected the lowest resilience of hydrophyte species on coarse sediment, because of unfavorable conditions for the survival of vegetative propagules, and the highest on organic one, the silty sediment exhibiting an intermediate pattern.

According to the intermediate disturbance hypothesis (Connell & Slatyer 1977; Huston

1979), species richness may peak for intermediate intensity of dewatering, which may be reached in silty habitats, because of the successful coexistence of hydrophytes, helophytes and amphiphytes. Coarse habitats may correspond to high disturbance intensity, and a low richness was expected because of too high disturbance intensity, only allowing helophytes and some amphiphytes to survive. A low richness was also expected on organic sediments, corresponding to the lowest disturbance intensity, allowing hydrophytes and amphiphytes to dominate the communities.

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Materials and methods

Study area

The study was conducted on 6 riverine wetlands located in the floodplain of the Ain River, which is a tributary of the Rhône River (France). The riverine wetlands were naturally created by the river dynamics, and consequently diverse in terms of sediment characteristics, depending on the local pattern of erosion-deposition and the energy of spates (Bornette et al. 1998;

Bornette et al. 2008).

Among the 6 wetlands, 18 sampling units were delimited and were divided into 3 sediment types: 6 gravel-dominated coarse sampling units (C), 6 silt-dominated sampling units (S), and

6 organic-matter dominated sampling units (O). For each sediment type, 3 sampling units were permanently aquatic (CP, SP and OP; Table 1) and the 3 others underwent a summer dewatering

(CT, ST and OT; Table 1).

In January 2011 (high water level period), sediment cores were collected in each sampling unit for measuring sediment organic matter content, mass water content and bulk density.

Organic matter content was analyzed following the Dumas method (Dumas 1831, organic matter content corresponding to organic carbon content × 1.72; Table 1). Wet and dry sediments

(after drying 48 hours at 105 °C) were weighted to calculate: bulk density (dry sediment mass/volume; Table 1) and mass water content ([wet sediment mass – dry sediment mass]/ dry sediment mass; Table 1).

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Table 1. Organic matter content, mass water content and bulk density of the sampling units studied.

Sampling unit Sediment type Water condition Wetland name Organic matter content Mass water content Bulk density (g.kg-1) (g.g-1) (g.cm-3) CP1 coarse permanent Bellegarde 1.47 0.1 1.9 CP2 coarse permanent Brotteaux 0.44 0.07 1.9 CP3 coarse permanent Villette 3.03 0.1 1.7 CT1 coarse temporary Bellegarde 0 0.1 1.4 CT2 coarse temporary Brotteaux 1.52 0.2 1.9 CT3 coarse temporary Villette 1.08 0.05 1.7 SP1 silty permanent Bellegarde 37.7 0.7 1.9 SP2 silty permanent Carronière 22.7 1.1 1.8 SP3 silty permanent Brotteaux 38.6 0.4 1.5 ST1 silty temporary Bellegarde 36.5 1.1 1.8 ST2 silty temporary Carronière 27.9 0.9 1.8 ST3 silty temporary Brotteaux 34.9 0.7 1.9 OP1 Organic permanent Villette 91.4 2 0.5 OP2 Organic permanent Albarine 105 1.2 0.7 OP3 Organic permanent Vers la Borne Ouest 63.5 2.3 0.4 OT1 Organic temporary Villette 59.1 1.2 0.7 OT2 Organic temporary Albarine 66 2 0.8 OT3 Organic temporary Vers la Borne Ouest 54.3 1.5 0.6

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Vegetation survey

Vegetation survey in the sampling units was done at the same 4 dates, successively before a summer dewatering event (T1; December 2010), at the beginning of the event (T2; April 2011), at the end of the event, before sediment rewetting (T3; September 2011), and 2 months after rewetting (T4; November 2011). The survey dates T1 corresponded to the pre-disturbance stage. As dewatering is a ramp disturbance, its intensity increases with its duration (Lake 2000).

For this reason, the T2 and T3 surveys were done for documenting the short (just after dewatering establishment) and long term (at the end of dewatering) resistance of plant communities. The T4 survey informed about community resilience. The permanently aquatic sampling units were used as reference undisturbed units. In each sampling unit, 5 quadrats of 1 square m were randomly selected. The plant coverage was measured at the four dates of survey in each quadrat (using Braun-Blanquet cover-abundance scale (1 = 0-5%; 2 = 5-25%; 3 = 25-

30%; 4 = 5-75%; 5 = 75-100%). The Braun-Blanquet scores of each species were converted to mean values of percentage covers (2.5; 15; 37.5; 62.5 and 87.5%) for statistical analyses.

For investigating how far the floristic changes through time related to plant biological types, each species was classified into one of 3 biological types: hydrophytes (strictly aquatic), amphiphytes (able to support short term dewatering) and helophytes (partly emerged) according to the flora of Lambinon et al. (2004) and Lauber & Wagner (2007).

For assessing how far the species content changed during dewatering/rewetting episode, species were classified into 4 response groups: resident species that were present both before and after dewatering, disappeared species that disappeared after dewatering, ephemeral species that appeared ephemerally during dewatering, appeared species that appeared during dewatering and were still present after rewetting.

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Water levels

The 6 wetlands belonged to the same river system and therefore were synchronically dewatered.

To quantify the changes in water levels, Mini-Diver® DI501 water-level data loggers were placed in each wetland. The compensation with the ambient atmospheric pressure was realized using a Baro-Diver DI500 placed at a locality close to the Ain River, thus providing the recovery of the study area (max 15 km radius). Measurements were made every hour and were averaged by month (Fig. 1). Water levels were also recorded manually with a graduated rad in each quadrat at each date of vegetation survey, and matched to month-averaged measurements to ensure the continuity of each condition.

Fig. 1. Chronicles of water level for each wetland, and position of sampling units in permanent water conditions (solid lines) and temporary ones (broken lines) relative to the recording spot.

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Data analysis

One-way ANOVAs followed by post-hoc tests (Tukey HSD) were used to test the effect of sediment type on organic matter content, mass water content and bulk density. The values of organic matter and mass water contents were log-transformed to improve the homogeneity of variances.

Vegetation data were analyzed using a within-class centered principal component analysis

(PCA); a class grouping the 4 dates of survey of each sampling unit. In this analysis, data are averaged per class, and averaged values are centered at the origin of the factorial axes. The analyze maximizes then the variance within the classes (i.e. among dates of survey), and each date of survey is then positioned on each factorial axis in function of its contrast with the mean value of the class (Benzécri 1983; Bornette & Amoros 1991). This analysis aimed at discriminating the temporal variance within a given sampling unit after having eliminated the average floristic contrast between sampling units. The factorial distances between the barycenter of dates of survey were then calculated for each quadrat for assessing short and long- term resistance and resilience. Short and long term resistance corresponded to factorial distances between factorial coordinates of barycenters at dates T1 and T2 and T1 and T3, respectively. Large distances reflect high changes in community structure and composition, i.e. low community resistance. Resilience was calculated as factorial distances between factorial coordinates of barycenters at dates T4 and T1. Large distances reflected low return of the community structure and composition to the stage that prevailed before the disturbance, i.e. low resilience. The data were log transformed to improve normality and homoscedasticity. Values were converted into negative to ensure that graphic reading is more intuitive.

Nested full-factorial two-way ANOVA followed by post-hoc test (Tukey HSD), taking into account the possible sampling unit effect, was used to test the effect of sediment type and water conditions on short and long term resistance and resilience of communities.

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The relationship between resilience and resistance was tested using a linear mixed-effect model with sampling units as random effect.

For each sediment type, linear mixed-effects models were used to test the effect of water condition and date of survey on the relative cover of the different biological types (hydrophytes, amphiphytes and helophytes) and response groups (residents, disappeared, ephemerals and appeared species), with date of survey, water condition (dewatered or permanently aquatic) and their interactions as fixed effects, and sampling units as random effect. When not significant, the interaction term was removed for obtaining the final models. To determine differences between dates of survey for both water conditions, post-hoc tests were performed.

The disturbance intensity on each sampling unit was considered as directly correlated to the disturbing effect of the dewatering event on the plant community, i.e. negatively correlated to its resistance to dewatering. Consequently, we used the community resistance as a proxy of the disturbance intensity to asses the relationship between species richness after disturbance (T4) and disturbance intensity. Quadratic mixed-effect models were fitted on the relation between species richness after disturbance (T4) and disturbance intensity (T1-T3 factorial distance), with sampling units as random effect.

All of the statistical analyses were performed using the R.3.0.2 software (R-

Development-Core-Team 2014).

Results

Sediment properties

Mass water content, organic matter content and bulk density differed significantly between

2 2 2 sediment types (F 15= 62.9, p < 0.001; F 15= 183.9, p < 0.001 and F 15= 85.8, p < 0.001 respectively). The mass water and organic matter contents were significantly higher in organic sediment than in the two other situations, and higher in silty sediment than in coarse one (p <

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0.01). The bulk density did not differ between silt and coarse sediments and was significantly higher than the one observed in organic sediment (p < 0.001).

Community resistance and resilience

Short and long term resistance in permanently aquatic sampling units did not differ between sediment types, and were significantly higher than the ones observed in temporary situations during the dewatering event (Table 2; Fig. 2a, b). Within temporary aquatic sampling units, both short and long-term resistance of plant communities significantly differed according to the sediment type (Table 2; Fig. 2a, b). The short-term resistance of plant communities did not differ between silty and organic sediments, but was significantly higher than the one observed on coarse sediment (Fig. 2a). The long-term resistance of plant communities was significantly higher on organic sediment than on the two others, and higher on silty sediment than on the coarse one (Fig. 2b).

The resilience in permanently aquatic sampling units did not differ among sediment types (Table 2; Fig. 2c). Within temporary aquatic sampling units, the resilience of plant communities significantly differed according to the sediment type (Table 2; Fig. 2c). For organic sediment, the resilience of plant communities did not differ between temporary and permanently aquatic sampling units (Fig; 2c). The resilience of plant communities was significantly higher on organic sediment than on the two other sediment types, and higher on silty sediment than on the coarse one (Fig. 2c).

Table 2. Effect of sediment type and water condition on short and long term resistance and resilience of communities tested by nested two-way ANOVA. Sediment type Water condition Sediment type × water condition 2 1 2 Short term resistance F 12= 87.31 *** F 12= 427.62 *** F 12= 59.95 *** 2 1 2 Long term resistance F 12= 23.55 *** F 12= 119.05 *** F 12= 11.71 ** 2 1 2 Resilience F 12= 86.81 *** F 12= 156.41 *** F 12= 63.75 *** The Fdf values and significance levels of the effects are presented. Significance levels: ***P < 0.001; **P < 0.01.

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RESULTATSESULTATS RéRéponsesponses à courcourtt termeterme ddeses communaucommunautéstés végétalesvégétales aquaaquatiquestiques

Fig. 2. (a) Short-term resistance, (b) long-term resistance and (c) resilience of plant communities according to water condition and sediment type (CP = coarse permanently aquatic; CT = coarse temporary aquatic; SP = silty permanently aquatic; ST = silty temporary aquatic; OP = organic permanently aquatic; OT = organic temporary aquatic). Different letters indicate significant differences between community short-term resistance, long-term resistance and resilience observed for each water condition and sediment type (Tukey HSD following nested full factorial two-way ANOVA).

1 The resilience of plant community increased significantly when resistance increased (F 35 =

38.59, p < 0.001; Fig. 4a).

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Changes in biological types

In permanently aquatic conditions, the abundance of the three biological groups (hydrophytes, helophytes and amphiphytes) remained unchanged through time whatever the sediment type

(Table 3, Fig. 3).

On coarse sediment, before the dewatering event, hydrophytes were absent of the sampling units in temporary conditions. The relative abundance of amphiphytes was higher in temporary than in permanently aquatic conditions, and the helophyte abundance did not differ according to the water permanency (Table 3, Fig. 3a). During the dewatering event, amphiphyte and helophyte covers respectively decreased and increased rapidly (from the beginning of disturbance) and continuously (during the whole dewatering event) and remained stable after rewetting (Table 3, Fig. 3a). Hydrophytes were absent during and after the dewatering event.

On silty sediment, before the dewatering event, hydrophyte and amphiphyte covers were respectively lower and higher in the temporary aquatic sampling units than in the permanently aquatic ones, and helophyte cover did not differ according to the water permanency (Table 3,

Fig. 3b). During the dewatering event, hydrophyte and amphiphyte covers respectively decreased and increased rapidly while the relative helophyte cover increased more slowly in the dewatered sampling units (Table 3, Fig. 3b). After rewetting, hydrophyte and amphiphyte covers increased and decreased respectively without fully recover from disturbance (Table 3,

Fig. 3b).

On organic sediment, before the dewatering event, hydrophyte and helophyte covers were respectively lower and did not differ in the temporary aquatic sampling units as compared with the permanently aquatic ones (Table 3, Fig. 3c). The amphiphyte cover was higher in temporary than in permanently aquatic sampling units and remained stable through the whole dewatering event and after rewetting (Table 3, Fig. 3c). Hydrophyte and helophyte covers

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RESULTATS Réponses à court terme des communautés végétales aquatiques respectively decreased and increased rapidly during the dewatering event in the dewatered units

(Table 3, Fig. 3c). After rewetting, they recover completely from disturbance (Table 3, Fig. 3c).

Changes in species response groups

In permanently aquatic conditions, the part of resident species remained stable through time, whatever the sediment type (Table 4, Fig. 3).

On coarse sediment, during the dewatering event, the part of resident species decreased and the part of new species increased rapidly and continuously (Table 4, Fig. 3a). After rewetting, the proportion of resident species did not recover its pre-disturbance stage and the proportion of new species that appeared during the dewatering event remained unchanged

(Table 4, Fig. 3a).

On silty sediment, during the dewatering event, the part of resident species decreased and the part of ephemeral species increased (Table 4, Fig. 3b). After rewetting, the proportion of resident species recovered its pre-disturbance stage, and the proportion of new species that appeared during the dewatering event remained unchanged (Table 4, Fig. 3b).

On organic sediment, during the dewatering event, the part of resident species decreased and few ephemeral species appeared and increased rapidly in abundance (Table 4, Fig. 3c). No new species colonized the sites during dewatering (Table 4, Fig. 3c). After rewetting, the proportion of resident species recovered its pre-disturbance stage (Table 4, Fig. 3c).

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Table 3. Effect of the date of survey, water conditions, and their interaction on the relative abundance of the biological types for each sediment type, tested by linear mixed-effects models. Date of survey Water condition Date of survey × water condition Coarse 3 1 3 Hydrophytes F 84= 0.14 ns F 28= 47.02 *** F 84= 0.14 ns 3 1 3 Amphiphytes F 84= 17.61 *** F 28= 6.34 * F 84= 19.25 *** 3 1 3 Helophytes F 84= 26.06 *** F 28= 45.17 *** F 84= 28.69 *** Silty 3 1 3 Hydrophytes F 84= 10.24 *** F 28= 61.26 *** F 84= 5.80 ** 3 1 3 Amphiphytes F 84= 5.40 ** F 28= 51.06 *** F 84= 3.03 * 3 1 3 Helophytes F 84= 6.11 *** F 28= 16.30 *** F 84= 5.51 ** Organic 3 1 3 Hydrophytes F 84= 6.71 *** F 28= 22.61 *** F 84= 5.12 ** 3 1 3 Amphiphytes F 84= 1.57 ns F 28= 12.59 ** F 84= 1.17 ns 3 1 3 Helophytes F 84= 3.64 * F 28= 7.3 * F 84= 3.64 * The Fdf values and significance levels of the effects are presented. Significance levels: ***P < 0.001; **P < 0.01; *P < 0.05; ns, not significant. The non-significant interaction terms that were excluded from the analysis are indicated in italics.

Table 4. Effect of the date of survey, water conditions, and their interaction, on the relative abundance of the response groups for each sediment type, tested by linear mixed-effects models. Date survey Water condition Date survey × water condition Coarse 3 1 3 Residents F 84= 24.03 *** F 28= 53.67 *** F 84= 24.03 *** Disappeared NA NA NA 3 1 3 Appeared F 84= 24.05 *** F 28= 60.09 *** F 84= 24.05 *** 3 1 3 Ephemerals F 84= 1.76 ns F 28= 1.79 ns F 84= 1.76 ns

Silty 3 1 3 Residents F 84= 8.17 *** F 28= 30.67 *** F 84= 7.78 *** 3 1 3 Disappeared F 84= 5.37 ** F 28= 7.66 ** F 84= 5.37 ** 3 1 3 Appeared F 84= 5.03 ** F 28= 14.15 *** F 84= 5.03 ** 3 1 3 Ephemerals F 84= 10.66 *** F 28= 11.62 ** F 84= 9.93 **

Organic 3 1 3 Residents F 84= 10.38 *** F 28= 10.47 ** F 84= 8.84 *** Disappeared NA NA NA 3 1 3 Appeared F 84= 2.17 ns F 28= 2.01 ns F 84= 1.16 ns 3 1 3 Ephemerals F 84= 10.35 *** F 28= 11.69 ** F 84= 9.26 ***

The Fdf values and significance levels of the effects are presented. Significance levels: ***P < 0.001; **P < 0.01; *P < 0.05; ns, not significant, NA: data not available.

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Fig. 3. Changeshanges in the relative abundance of biological types and response groups throthrough time for each sediment type: (a) coarse; (b) silty and (c) organic, in permanent and temporary water conditions respectively.

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Species richness

The quadratic model used for identifying the humped shaped relationships between community resistance, as a proxy of disturbance intensity, and species richness after disturbance

1 (F 34 = 17.33, p < 0.001) was not significant (quadratic term not significant). The relationship between species richness after disturbance and disturbance intensity was consequently assessed using a linear mixed-effect model with sampling unit as random effect. The species richness

1 after rewetting was significantly and negatively related with community resistance (F 35 =

16.17, p < 0.001, Fig. 4b), meaning that the higher the disturbance intensity was (i.e. the lower community resistance), the higher the species richness was after disturbance.

FiFig. 44. RRelationship l ti hi between b t (a)() resilienceili and d resistance i t ((expressed d as ththe ffactorial t i l didistances t between factorial coordinates of sampling unit barycenters at survey dates T1 and T4 and T1 and T3 respectively), and (b) species richness and resistance (the resistance of sampling units facing dewatering was used as a proxy of dewatering intensity). Resistance and resilience were converted into negative values to ensure that graphic reading is more intuitive. Each point represents a quadrat; the 5 quadrats sampled in a given sampling unit are represented by the same symbol.

Discussion

The high organic matter content of organic sediments explained their low bulk density and high water content, corresponding to sediments with a high capacity of water retention (Avnimelech et al. 2001). Silty and coarse sediments had high and similar bulk density but for different reasons. Silty sediments are composed of many particles of small size, which corresponds to

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RESULTATS Réponses à court terme des communautés végétales aquatiques high bulk density but low porosity. Coarse sediments are composed of fewer heavier large particles with large spaces between them and consequently both high bulk density and porosity.

The combination of a relatively low organic matter content and a high bulk density, allowed the silty sediments to have a relatively high water content (Rawls et al. 2003), while the high porosity and low organic matter content of coarse sediments explain their low mass water content, and reflect a low capacity to retain water (Saxton & Rawls 2006).

As hypothesized, the resistance and resilience of plant communities to the dewatering event differed according to the sediment type. At the beginning of the dewatering period, community responses poorly differed according to sediment type, but the expected gradient of increased disturbance intensity from organic, silty, and then coarse sediment was observed at the end of the dewatering period. This gradient of disturbance also related to resilience processes, as communities recovered more when they grew on organic sediment, and less when they colonized coarse sediment. The linear and positive relationship between resilience and long-term resistance suggests that a high resistance allows a low alteration of communities, and low community resistance was associated with low community ability to recover after the event.

These three sediment types can consequently be considered as leading to a gradient of disturbance intensity for similar dewatering periods.

Only few vegetation changes were observed for organic sediment. This may be related to its ability to retain efficiently water, limiting by this way plant dehydration, and allowing the maintenance of amphiphytes and of the more tolerant hydrophytes. Due to this high community resistance, few niches were available, and therefore only few helophytes could colonize the sampling units during the dewatering event. Vegetation changes in silty sediment were important during the dewatering/rewetting episode. Silty sediment may allow a great conservation of propagules which may potentially develop in both conditions (Liu et al. 2006;

Capers 2003). The disturbance was high enough to lead to the disappearance of vegetative parts

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RESULTATS Réponses à court terme des communautés végétales aquatiques of aquatic plants, but seeds and propagules may have been preserved in the substrate, and rapidly sprout when rewetting, allowing a good resilience of communities. During the dewatered period, some ephemeral species established when a gap was created in the vegetation

(Seabloom et al, 1998). In this case, the establishment of optional ruderal amphiphyte and helophyte species may have occurred from the propagule bank or from dispersed propagules

(Coops & van der Velde 1995; Abernethy & Willby 1999; Havens et al. 2005). On coarse sediment, vegetation changes due to dewatering corresponded to the decreasing relative abundance of resident amphiphyte species, together with the colonization by helophytes, and their maintenance after rewetting. Coarse sediment is poorly favorable to seed and propagule survival in case of complete dewatering, potentially explaining that no ephemeral species colonized the sites during the dewatering phase, and that the resilience of communities was low.

Van der Valk (1981, 2005) describes two kinds of vegetation change during wet-dry cycles of wetlands: 1) fluctuations, defined as changes in the relative abundance of species and

2) succession, defined as changes in species composition. The changes in species composition may be cyclic with the alternation of terrestrial and aquatic communities. Fluctuations and cyclic successions may allow the stability of communities through time. Patterns observed in our study during one dewatering/rewetting event suggest that 1) dewatering events on organic sediment result on low vegetation changes, i.e. fluctuation; 2) marked floristic changes and strong community resilience that occur on silty sediments may correspond to cyclic succession and 3) low resistance and resilience of plant communities on coarse sediment may result in succession, and long-term community changes. However the latter pattern remains to be confirmed by long-term survey, because resilience may be achieved over a longer time scale because the disturbance is more intense on coarse sediment.

In temporary habitats, we observed in both organic and silty habitats a lower hydrophyte, and higher amphiphyte relative abundance than in permanent water conditions, and

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RESULTATS Réponses à court terme des communautés végétales aquatiques the absence of hydrophytes in coarse habitats, suggesting that species tolerating dewatering were selected in these habitats by previous dewatering events.

Contrary to what we expected, we did not observe a humped-shaped relationship between species richness and dewatering intensity but a linear one. The lowest species richness was recorded for organic sediment, and may be related to low disturbance intensity. In such situation, the disturbance intensity may be sufficiently low to allow competitive species to maintain in the habitat, impeding the recruitment of species that may be more tolerant to water stress. The highest species richness was recorded for coarse sediment, corresponding to high disturbance intensity. High species richness was partly related to the recruitment of new helophyte species in the disturbed areas during the dewatering phase. If there is long-term resilience of communities on coarse substrate, species richness may then decrease over time, due to the progressive failure of these helophyte species in the submerged conditions that followed dewatering. Intermediate situations occurred on silty habitats, in which species richness was relatively high due to the simultaneous presence of hydrophyte, amphiphyte and helophyte species after the disturbance.

Only few studies reported the effect of dewatering on the survival and viability of aquatic plants (Doyle & Smart 2009). However, the thresholds of species tolerance to dewatering and subsequent substrate drying may not only relate to their biological type

(helophytes, amphiphytes, helophytes), but, within a given group, may also differ according to their growth form. As an example, as sediment drying may decrease with sediment depth, the species with a deep anchoring system may better survive that the shallower anchored ones. It is consequently necessary, for determining how species may tolerate such events, to link species growth form to their resistance and resilience to dewatering.

Our study demonstrated that a rather simple description of sediment types allows predicting the impact of dewatering on plant communities. However, within a given sediment

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RESULTATS Réponses à court terme des communautés végétales aquatiques category, the variance in the plant response may vary according to the texture and grain-size of sediment. For example, the proportion of clay, or the organic matter concentration and fiber content of sediments may greatly alter its drying kinetic, and may be quantified for improving the model. Propagule bank is known for its significant role in the establishment of plant communities in wetlands but the abundance, location and survival of seeds in the sediment probably differ according to the sediment type, even in permanently aquatic conditions. The occurrence, density and composition of propagule banks in the different sediment types, have still to be assessed, as they can greatly affect the pattern of vegetation changes during and after the dewatering events, either by controlling directly the species composition, or by allowing the recruitment of species that may act as nurses, favoring the recruitment or survival of other species by impeding the sediment desiccation (through the shade they provide, for example).

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RESULTATS Réponses à court terme des communautés végétales aquatiques

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Engelhardt, K.A.M., & Ritchie, M.E. 2001. Effects of macrophyte species richness on wetland ecosystem functioning and services. Nature 411: 687-689. Engelhardt, K.A.M., & Ritchie, M.E. 2002. The effect of aquatic plant species richness on wetland ecosystem processes. Ecology 83: 2911-2924. Erwin, K.L. 2009. Wetlands and global climate change: the role of wetland restoration in a changing world. Wetlands Ecology and Management 7:71-84. Havens, K.E., Fox, D., Gornak, S. & Hanlon, C. 2005. Aquatic vegetation and largemouth bass population responses to water-level variations in Lake Okeechobee, Florida (USA). Hydrobiologia 539:225-237. Huston, M. 1979. A general hypothesis of species diversity. The American Naturalist 113: 81- 101. Junk, W.J., An, S., Finlayson, C.M., Gopal, B., Květ, J., Mitchell, S.A., Mitsch, W.J. Robarts, R.D. 2013. Current state of knowledge regarding the world’s wetlands and their future under global climate change: a synthesis. Aquatic Sciences 75:151-167. Keddy, P.A., Fraser, L.H., Solomeshch, A.I., Junk, W.J., Campbell, D.R., Arroy, M.T.K. & Alho, C.J.R. 2009. Wet and wonderful: the world’s largest wetlands are conservation priorities. BioScience 59: 39-51. Lake, P.S. 2000. Disturbance, Patchiness, and Diversity in Streams. Journal of the North American Benthological Society 19: 573-592. Lake, P.S. 2003. Ecological effects of perturbation by drought in flowing waters. Freshwater Biology 48: 1161-1172. Lambinon, J., De Langhe, J.E., Delvosalle, L. & Duvigneaud J. 2004. Nouvelle flore de la Belgique, du grand-duché de Luxembourg, du nord de la France et des régions voisines. 5th edition. Jardin botanique national de Belgique, Meise. Lauber, K. and Wagner, G. 2007. Flora Helvetica. 4th edition, Haupt, Bern.. Lehner, B., Döll, P., Alcamo, J., Henrichs, T. & Kaspar, F. 2006. Estimating the impact of global change on flood and drought risks in Europe: a continental, integrated analysis. Climatic Change 75: 273-299. Liu, G., Li, W., Zhou, J., Liu, W., Yang, D., & Davy, A.J. 2006. How does the propagule bank contribute to cyclic vegetation change in a lakeshore marsh with seasonal drawdown? Aquatic Botany 84: 137-143. Millennium Ecosystem Assessment. 2005. Ecosystems and human well-being: wetlands and water. Synthesis. World resource institute, Washington DC. RAMSAR. 2002. Les zones humides, valeurs et fonctions. Gland, Suisse. Rawls, W.J., Pachepsky, Y.A., Ritchie, J.C., Sobecki, T.M. & Bloodworth, H. 2003. Effect of soil organic carbon on soil water retention. Geoderma 116: 61-76. Robe, W.E. & Griffiths H. 1998. Adaptations for an amphibious life: changes in leaf morphology, growth rate, C and N investment, and reproduction during adjustment to emersion by the freshwater macrophyte Littorella uniflora. New Phytologist 140:9-23. Saxton, K.E. & Rawls W.J. 2006. Soil water characteristic estimates by texture and organic matter for hydrologic solutions. Soil Science Society of America Journal 70:1569-1578. Sculthorpe, C.D. 1967. The biology of aquatic vascular plants. Edward Arnold Ltd, London

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Seabloom, E.W., van der Valk, A.G. & Moloney K.A. 1998. The role of water depth and soil temperature in determining initial composition of prairie wetland coenoclines. Plant Ecology 138: 203-216. van der Valk, A.G. 1981. Succession in wetlands: a Gleasonian approach. Ecology 62: 688- 696. van der Valk, A.G. 2005. Water-level fluctuations in North American prairie wetlands. Hydrobiologia 539: 171-188. van Geest, G.J.V., Coops, H., Roijackers, R.M.M., Buijse, A.D., & Scheffer, M. 2005a. Succession of aquatic vegetation driven by reduced water-level fluctuations in floodplain lakes. Journal of Applied Ecology 42: 251-260. van Geest, G.J.V., Wolters, H., Roozen, F.C.J.M., Coops, H., Roijackers, R.M.M., Buijse, A.D., & Scheffer, M. 2005b. Water-level fluctuations affect macrophyte richness in floodplain lakes. Hydrobiologia 539: 239-248. Walczak, R., Rovdan, E. & Witkowska-Walczak, B. 2002 Water retention characteristics of peat and sand mixtures. International Agrophysics 16: 161-165. Wilcox, D.A & Nichols, J. 2008. The effects of water-level fluctuations on vegetation in a lake Huron wetland. Wetlands 28: 487-501. Winter, T.C. 2000. The vulnerability of wetlands to climate change: a hydrologic landscape perspective. Journal of the American Water Resources Association 36: 305-311.

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DISCUSSION

Mon travail de doctorat avait pour objectif de comprendre comment les exondations régissent le fonctionnement et les communautés végétales aquatiques des zones humides en climat tempéré, en mesurant leurs effets à 3 niveaux emboités (espèce, communauté végétale,

écosystème) et à 3 échelles de temps (pluriannuelle, saisonnière, et à l’échelle de quelques jours).

1. Aspects fonctionnels

1.1 Fonctionnement hydrologique et biogéochimique

Les conséquences des exondations sur les caractéristiques physico-chimiques des zones humides étaient jusque-là principalement abordées à travers des études à court terme en laboratoire, sans prendre en compte le fonctionnement hydrogéologique des écosystèmes (De

Groot & Van Wijck 1993 ; Olila et al. 1997 ; Olde Venterink et al. 2002 ; Gómez et al. 2012).

Ce travail a permis de démontrer que l’augmentation de l’intensité des exondations sur une période de 15 ans ne se traduit pas forcément par une augmentation de la teneur en nutriments

(eutrophisation interne) des eaux de surface liée à la minéralisation des sédiments exondés. Les zones humides alimentées par des eaux souterraines peuvent ne pas montrer de phénomène d’eutrophisation, grâce probablement au lessivage des nutriments libérés lors de la remise en eau suivant l’exondation.

Les résultats obtenus montrent également que le contexte hydrogéologique et géomorphologique des zones humides est à prendre en compte comme facteur explicatif des changements physico-chimiques observés. Les patrons suggèrent soit une augmentation de l’alimentation en eau des zones humides par la nappe de versant (augmentation de la conductivité et température constante), liée à l’augmentation du drainage de la zone humide par la rivière (phénomène d’incision), soit une augmentation de l’alimentation de la zone humide par la nappe d’accompagnement (augmentation de la température et du pH, diminution de la

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DISCUSSION conductivité), du fait de la migration latérale de la rivière, et de son rapprochement de la zone humide.

Dans mon travail de thèse, des suivis in situ des paramètres physico-chimiques (pH,

- + 3- oxygène, conductivité, température et teneur en nutriments : NO3 , NH4 et PO4 ) des sédiments et des eaux interstitielle et de surface ont été réalisés au niveau de différents types de sédiments ayant subi un épisode d’exondation ou non. Les données témoignent d’une grande variabilité des patrons. Ceci peut s’expliquer par le fait que la remise en eau, entrainant le relargage des nutriments, correspond à une période relativement courte (« hot moment » ; McClain et al.

2003) qu’il est difficile de détecter sans une mesure en continu de ces paramètres lors de l’épisode d’exondation et de remise en eau. Les zones humides étudiées, de surcroit, peuvent

être soumises à des apports d’eau d’origine variée (nappe d’accompagnement de la rivière, nappe karstique, eau de surface de la rivière) dans des proportions elles aussi variées, ce qui peut conduire, au moment de la remise en eau, à augmenter la variabilité inter-sites des patrons observés, et à masquer les variations physico-chimiques relatives aux processus biogéochimiques de relargage de nutriments dans la lame d’eau. (Corstanje & Reddy 2004).

Dans mon travail, j’ai constaté que la qualité du sédiment, et en particulier sa teneur en matière organique, régissait sa capacité de rétention de l’eau et par conséquent la tolérance des communautés végétales aux exondations. Je n’ai cependant pas abordé la question de l’intensité de l’assèchement sur les processus biogéochimiques. Pourtant, selon l’intensité de la déshydratation, la disponibilité des différentes formes d’azote peut différer dans le sédiment

(James et al. 2004 ; Austin & Strauss 2011), et les processus en jeu peuvent également dépendre de la matière organique et la granulométrie du sédiment. En effet, en influençant la taille des pores, ces deux paramètres jouent un rôle prépondérant dans la capacité du sol à retenir l’eau, et par conséquent dans la quantité d’air qui y circule pendant la phase de ressuyage (Walczack et al. 2002 ; Rawls et al. 2003). De plus, la matière organique fournit le carbone organique

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DISCUSSION nécessaire aux bactéries hétérotrophes (ammonification et dénitrification). La teneur en matière organique des sédiments a donc un rôle clef dans la réponse du fonctionnement biogéochimique des zones humides aux exondations dont la mesure de la complexité n’était pas l’objet du présent travail.

1.2 Structuration des communautés végétales aquatiques

Dans ce travail, la réponse des communautés à l’exondation a été étudiée sur un temps court (une année), et à l’échelle de la tache de végétation. Mes résultats montrent qu’une description simple du type de sédiment peut permettre de prédire la réponse à court terme des communautés végétales aquatiques à l’exondation. En effet la réponse des communautés à un

épisode d’exondation diffère selon le type de sédiment, en particulier en lien avec sa capacité à retenir l’eau.

La structuration des communautés à plus long terme pourrait dépendre de cette réponse

à court terme mais également du potentiel de modification de la trophie du milieu liée aux assecs. La trophie des habitats peut régir la capacité des plantes à tolérer l’exondation, en augmentant la résistance des communautés (augmentation des ressources disponibles, favorisant la croissance des plantes, James et al. 2004 ; et la production de réponses plastique), ou leur résilience (augmentation du succès reproducteur). Elle peut également agir négativement sur la résilience des communautés, en défavorisant le stockage des réserves chez les plantes (Puijalon et al. 2008), diminuant ainsi leur capacité à produire un phénotype adapté

à l’exondation.

Si les changements floristiques observés à court terme se maintiennent à plus long terme, notamment la dominance d’espèces émergentes sur les substrats grossiers soumis aux exondations, des changements en terme de quantité et de qualité de la matière organique devraient être observés au niveau du sédiment. Les espèces émergentes ayant une production primaire élevée (Westlake et al. 1998), une accumulation de matière organique dans le sédiment

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DISCUSSION pourrait être observée, qui entrainerait potentiellement, au moins pendant les phases inondées, une anaérobiose plus marquée, et des changements des processus biogéochimiques (par exemple augmentation de la dénitrification, Bastviken et al. 2009). Cette activité devrait dépendre de la qualité de la matière organique, et donc de la qualité des végétaux et de leur nécro-masse (en particulier, en termes de tissus ligneux et d’azote), qualité qui pourrait conduire

à une limitation de l’activité des bactéries hétérotrophes dénitrifiantes car moins facilement assimilable (Bastviken et al. 2005).

2. Aspects évolutifs : ajustements plastiques des espèces

La prévision des réponses des communautés végétales aux perturbations est souvent faite à l’aide de traits fonctionnels censés traduire les fonctions écologiques d’intérêt pour la survie face aux perturbations, mais ces prévisions posent rarement la question de la variation des traits individuels induite par la perturbation (norme de réaction) et de la performance de l’ajustement observé (en termes de productivité, de régénération végétative post perturbation, ou de succès reproducteur).

Les plantes aquatiques étudiées peuvent pour la plupart d’entre elles mettre en place un phénotype permettant de survivre pendant l’exondation (stratégie de résistance : capacité à minimiser l’impact négatif des conditions environnementales ; Levitt 1972). Deux types d’ajustements plastiques ont été observés dans mon étude : un phénotype autoportant (stratégie de tolérance, par exemple liée à la production d’une tige verticale, ligneuse, autoportante) et un phénotype d’évitement (réduction de la partie aérienne, port prostré, et augmentation de la biomasse souterraine). La résilience potentielle ou avérée des espèces via la production de semences résistantes à l’exondation et leur recrutement lors de la remise en eau n’a pas été quantifiée. Pourtant, on a observé à plusieurs reprises lors des mesures de terrain que les plantes fleurissaient parfois abondamment pendant la mise en place de l’exondation suggérant que celle-ci pourrait stimuler la reproduction sexuée, car elle s’accompagne d’une augmentation de

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DISCUSSION la température de l’eau, du carbone bio-disponible (via la mise en contact des tissus avec le milieu aérien) et de l’éclairement (via la diminution de l’épaisseur de la lame d’eau), et donc potentiellement de l’activité photosynthétique du végétal (Nielsen & Sand-Jensen 1989; Sand-

Jensen & Frost-Christensen 1999). De plus, la faible profondeur permet à la plante de produire des fleurs aériennes pour un cout énergétique plus modeste (taille nécessaire pour atteindre la surface plus faible ; (Barrat-Segretain 1996 ; Boeger & Poulson 2003). Les espèces aquatiques en réponse aux exondations pourraient donc constituer une banque de graines abondante favorisant la résilience des communautés en cas de mort des individus (Combroux et al. 2001 ;

Brock et al. 2003 ; Bonis et al. 1995). La part relative de reproduction sexuée vs. asexuée ayant potentiellement une influence sur la diversité génétique au sein et entre les populations (Honnay

& Bossuyt 2005 ; Pollux et al. 2007), un régime d’exondations marquée pourrait favoriser le brassage génétique et la diminution de la taille des clones dans les populations, et même, chez certains groupes (eg. Potamogetonacées, Ranunculacées) favoriser la genèse et le recrutement d’hybrides interspécifiques (Iida et al. 2007).

Le choix d’étudier la plasticité en condition expérimentale et in situ s’est avéré pertinent car certaines espèces communes aux 2 protocoles montrent des réponses différentes dans les deux études. Certaines espèces n’ayant pas survécu à l’exondation en condition expérimentale y survivent in situ, et même montrent des ajustements plastiques non observés en laboratoire

(Potamogeton coloratus), et parfois une biomasse comparable (Sparganium emersum). Le caractère abrupt de la mise en place de l’exondation en laboratoire peut avoir diminuer la capacité des plantes à produire un phénotype adapté. De surcroit, l’expérimentation en laboratoire a dissocié l’exondation des autres variables environnementales covariant en général in situ (augmentation de la température, de l’éclairement, par exemple). Hors, ces paramètres peuvent stimuler l’activité métabolique des plantes, favorisant la production de phénotypes

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DISCUSSION adaptés. Il est donc probable que les modalités de mise en place de l’exondation et les conditions d’habitat expliquent les contrastes observés.

La capacité des plantes aquatiques à faire face à l’exondation semble dépendre, selon mes premiers résultats, de leur niche écologique et de leur position phylogénétique. La forme de croissance ne semble pas influencer la capacité des espèces à faire face à l’exondation et semble uniquement régir, en partie, certains traits morpho-anatomiques impliqués dans la réponse plastique. Enfin, il reste difficile d’établir un lien entre la variation des traits morpho- anatomiques mesurés et la performance (croissance, taille) des espèces.

Les modèles de stratégies adaptatives des végétaux stipulent que les traits biologiques des espèces permettent de déterminer leur tolérance aux perturbations et aux facteurs de stress

(Grime 1979, 2002). La forte plasticité de nombreuses espèces de plantes aquatiques leur permet d’ajuster leur phénotype aux conditions de l’habitat, leur permettant d’être performantes dans différentes conditions de stress et de perturbation (e.g. Sparganium emersum ; Santamaria et al. 2003). Ces observations peuvent rendre difficile l’agencement des différentes espèces dans les différentes stratégies et donc la prédiction de la structure et la composition des communautés végétales soumises aux exondations.

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PERSPECTIVES

1. Aspects fonctionnels

1.1 Fonctionnement hydrologique et biogéochimique

L’étude que j’ai menée in situ sur la réponse des écosystèmes aux exondations ne considérait que 3 situations bien tranchées (en termes de qualité de sédiment) et ne prenait pas en compte les processus s’y déroulant. En particulier, les cinétiques de ressuyage, de dessèchement, de compaction, et de libération de nutriments au moment de la remise en eau, n’ont pas été mesurées. Une perspective intéressante consisterai à étudier en laboratoire des sédiments naturels de composition variable, représentant de larges gammes de situations (en termes de teneur en matière organique, de granulométrie) afin de mieux comprendre comment ces sédiments se comportent pendant la mise en place de l’assèchement et au moment de la remise en eau. Des sédiments sablo-graveleux, pauvres en matière organique, pourraient s’oxyder et se réchauffer plus rapidement lors de l’épisode d’exondation, car ils présentent une faible capacité à retenir l’eau. Ces conditions pourraient engendrer une stimulation de la nitrification au moins les premiers jours de l’exondation, mais une inhibition de la dénitrification (Gómez et al. 2012). La dessiccation rapide de ces sédiments pourrait entrainer une forte mortalité microbienne et le relargage d’azote et de phosphore du fait de la lyse cellulaire (De Groot &

Van Wijck 1993 ; Baldwinn & Mitchell 2000). Les sédiments à granulométrie plus fine

(limono-argileux) sont potentiellement plus riches en matière organique, et pourraient se réchauffer et s’oxyder plus lentement lors de l’exondation, permettant ainsi la réalisation simultanée de phénomènes de nitrification et de dénitrification et donc une perte potentielle d’azote (Koschorreck 2005). De plus, la matière organique fournit le carbone organique nécessaire aux bactéries hétérotrophes (ammonification et dénitrification). Ainsi dans un sédiment ayant une teneur en matière organique élevée, la disponibilité en carbone organique ne sera pas limitante et permettra une activité microbienne élevée résultant en une libération

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PERSPECTIVES importante de nutriments (McIntyre et al. 2009). En contrepartie dans les sols pauvres en matière organique, le carbone organique peut être rapidement épuisé dès les premiers jours de l’exondation par une minéralisation importante, ce qui aura pour conséquence de limiter la relargage de nutriments (Gómez et al. 2012).

1.2 Structuration des communautés végétales aquatiques

Dans cette thèse, seule la réponse à court terme des communautés aux exondations a été étudiée.

L’efficacité des processus de dispersion et l’intensité des interactions biotiques n’ont pas été abordées. Cependant, les deux phénomènes agissent probablement fortement sur la dynamique

à long terme des communautés soumises à des exondations récurrentes (Keddy 1992 ; Lortie et al. 2004). La dispersion devrait fortement dépendre du type de zones humides et de leur connectivité (White & Jentsch 2001). Les interactions biotiques peuvent être modifiées par les exondations car celles ci modifient les conditions abiotiques après la remise en eau (compétition pour les ressources), et parce qu’elles détruisent une partie des organismes végétaux et animaux

(White & Pickett 1985 ; White & Jentsch 2001), limitant les interactions biotiques entre les organismes (compétition pour l’espace, broutage, parasitisme, par exemple).

Les espèces tolérant l’exondation par la mise en place d’un port autoportant pourraient montrer une meilleure aptitude compétitive durant la phase d’exondation. La forte teneur en

éléments structuraux pourrait leur conférer une palatabilité moindre et donc une plus forte résistance à l’herbivorie. Les espèces évitant l’exondation sous forme chétive avec une augmentation de leur ratio racine/tige, pourraient constituer un phénotype peu compétitif pendant la phase d’exondation, mais pourraient présenter des capacités élevées de prélèvements des ressources (eau, nutriments) du fait de leur développement racinaire important. Dans le cas d’un milieu riche en nutriments, les espèces compétitives à phénotype autoportant pourraient

être favorisées, alors que dans les milieux oligotrophes, les espèces à port prostré domineraient.

Une analyse de la représentation de ces différentes formes de croissance dans des écosystèmes

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PERSPECTIVES temporaires présentant des niveaux de ressources différents, mais des sédiments de qualité comparable pourraient permettre d’établir les règles d’assemblages des communautés végétales aquatiques faisant face aux exondations.

La banque de propagules, composante importante dans la dynamique des communautés végétales aquatiques soumises aux exondations, n’a pas été abordée dans cette thèse. Le type de sédiment devrait régir la constitution (au travers des ressources disponibles) et la résistance de la banque de propagules aux assèchements (au travers de sa cinétique de dessèchement).

L’étude de composition de cette banque et de la viabilité des différents types de propagules au cours de l’exondation, en fonction du type de sédiment, reste à être réalisée.

2. Aspects évolutifs : ajustements plastiques des espèces

Mon travail repose sur l’étude d’une seule population par espèce, l’histoire d’exondation peut avoir fortement sélectionné les populations résistantes, engendrant de fortes différences populationnelles au sein d’une même espèce en terme de performance et de réponse plastique.

Ainsi, on peut faire l’hypothèse que la réponse plastique et la performance des espèces face aux exondations devraient être supérieures pour les populations se développant dans des

écosystèmes soumis à cette contrainte (adaptation locale ; Van Kleunen & Fischer 2011). Ainsi des expérimentations de transplantations réciproques in situ et des expérimentations en jardin commun de différentes populations génétiquement distinctes permettraient de mettre en

évidence l’existence d’écotypes et leur diversité de réponse en fonction des groupes taxonomiques et de leur appartenance phylogénétique.

Les plantes aquatiques sont potentiellement confrontées à des changements du niveau trophique de leur habitat du fait des exondations. En conditions eutrophes, les plantes produisent des tissus de faible densité, avec de faibles teneurs en composés structuraux du fait de leur taux de croissance élevé (Ryser 1996 ; Craine et al. 2001 ; Puijalon et al. 2007), et certaines études soulignent qu’elles tendent à investir une plus grande proportion de leurs tissus

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à des aérenchymes du fait des conditions anoxiques induites par l’eutrophisation (Hussner et al. 2009 ; Ryser et al. 2011). Ces réponses anatomiques peuvent avoir des répercussions sur la capacité des plantes à supporter les contraintes mécaniques telles que la gravité lors des exondations (Lamberti-Raverot & Puijalon 2012 ; Puijalon et al. 2011 ; Hamann&t Puijalon

2013). On pourrait alors observer un effet synergique de ces deux contraintes (eutrophisation et exondation) qui limiterait de manière accentuée la capacité des espèces à tolérer l’exondation.

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Comme dit mon voisin, une bonne thèse est une thèse finie !!! Conséquences des exondations pour les communautés végétales aquatiques et le fonctionnement des zones humides fluviales

Résumé :

Mots-clés : Zones humides, Exondation, Ecologie des communautés, Végétation aquatique, Plasticité, Hydrologie, Caractéristiques physico-chimiques

Consequences of dewatering for aquatic plant communities and the functioning of riverine wetlands

Abstract :

The objective of this thesis was to measure how changes in hydrological regimes, particularly dewatering govern 1) aspects of the functioning of wetlands, 2) the organization and short-term dynamics of aquatic plant communities and 3 ) survival and plastic response of aquatic plants. This thesis addresses issues at different spatial and temporal scales. First, at the decade scale, I measured the effect of water-level decreases in riverine wetlands on their physico-chemistry characteristics. Second, at the season scale, I measured the influence of sedimentary characteristics of wetlands on short-term response of plant communities to dewatering. Finally, at the scale of a few weeks, I was interested in the ability of aquatic angiosperm species to develop a plastic adjustment to dewatering, in experimental laboratory conditions and in situ, and I looked determinism of this response (ecological, morphological, phylogenetic). In terms of physico-chemical characteristics of surface waters, the 15- year study of the dynamics of riverine wetlands undergoing dewatering, not reached, as is usually described in the literature, with an increase of water body nutrient contents, but rather changes suggesting variations of the hydrogeological functioning of wetlands in favor of a greater influence of the hillslope groundwater table in their water supply. In situ response of plant communities to dewatering differs according to sediment type. Both, resistance and resilience of communities decrease with the sediment water retention capacity. The ability of aquatic plants to tolerate dewatering, in experimental conditions, seems to differ according to their phylogenetic position, but not according to their growth form (rosettes or caulescentes). Species tolerating dewatering show phenotypic adjustments such as denser aerial organs and high plasticity of the leaves, which may explain the maintenance of a similar growth rate in terrestrial and aquatic conditions in these species. Comparison of emerged and submerged phenotypes in situ, also suggests that the phylogenetic origin and the ecological niche of the species (amphiphytes or hydrophytes) govern the plastic adjustments and performance of species to dewatering.

Keywords: Wetlands, Dewatering, Community ecology, Aquatic vegetation, Plasticity, Hydrology, Physico- chemical characteristics