Effets Des Coupes De Récupération Sur Les Successions

Effets des coupes de récupération sur les successions naturelles de coléoptères saproxyliques le long d’une chronoséquence de 15 ans après feu en forêt boréale commerciale

Mémoire

Olivier Jeffrey

Maîtrise en sciences forestières avec mémoire Maître ès sciences (M.Sc.)

Québec, Canada

Résumé

Cette étude vise à décrire l’état naturel des communautés de huit familles de coléoptères saproxyliques, reconnues pour être associées aux brûlis, à partir d’un échantillonnage fait au niveau des souches et à l’intérieur de peuplements commerciaux d’épinettes noires (Picea mariana (Mill)) résiduels à la coupe de récupération (état naturel) provenant de brûlis de 1995, 2003, 2005, 2007, 2009 et 2010. L’étude vise également à comparer ces communautés à celles retrouvées dans des peuplements d’épinettes noires récupérés et répartis dans les brûlis de 1995, 2003, 2005 et 2007. Les coléoptères ont été inventoriés durant l’été 2010 et plus de 6 000 spécimens répartis dans les huit familles étudiées ont été capturés. À l’état naturel, quatre phases de colonisation se sont succédées pendant les cinq premières années après feu. Les assemblages des communautés de coléoptères saproxyliques sont quant à eux altérés par la récupération et ce, pour une période d’au moins sept ans.

iii

Abstract

This study aims to describe the natural state of communities of eight saproxylic beetle families known to be associated with burned forest from a sample done at stump level and in commercial black spruce (Picea mariana (Mill)) stands residual to salvage logging (natural state) from burns of 1995, 2003, 2005, 2007, 2009 and 2010. The study also aims to compare these communities to those found in salvaged black spruce stands distributed in burns of 1995, 2003, 2005 and 2007. Beetles were sampled during summer 2010 and over 6 000 specimens among the eight studied families were captured. In its natural state, four colonization phases have succeeded along the first five postfire years. Saproxylic beetle assemblages were affected by salvage logging for a period of seven years.

Table des matières

Résumé ...... iii Abstract ...... v Table des matières ...... vii Liste des tableaux ...... ix Liste des figures ...... xi Remerciements ...... xiii Avant-Propos ...... xv

INTRODUCTION GÉNÉRALE ...... 1 Les feux de forêt ...... 1 Le bois mort : un élément clé de la diversité biologique ...... 2 Importance des feux pour les coléoptères saproxyliques ...... 3 Problématique ...... 5 Impacts des coupes de récupération sur la faune et la flore ...... 6 Développement durable et aménagement écosystémique ...... 7 Contexte de l’étude ...... 8

CHAPITRE 1 - SAPROXYLIC BEETLE SUCCESSION ALONG A 15 YEARS POSTFIRE CHRONOSEQUENCE IN THE COMMERCIAL BOREAL FOREST.. 13 Résumé ...... 14 Abstract ...... 15 1 Introduction ...... 16 2 Methods ...... 18 2.1 Study area ...... 18 2.2 Stand description ...... 18 2.3 Beetle sampling ...... 20 2.4 Statistical analysis ...... 20 2.4.1 Habitat attributes along the chronosequence ...... 21 2.4.2 Saproxylic beetle abundance and species richness along the chronosequence .. 21 2.4.3 Beetle assemblages along the chronosequence ...... 22 2.4.4 Association between saproxylic beetles and habitat attributes along the chronosequence ...... 22 3 Results ...... 23 3.1 Overview ...... 23 3.2 Habitat attributes along the chronosequence ...... 24 3.3 Saproxylic beetle abundance and species richness along the chronosequence ...... 25 3.4 Beetle assemblages along the chronosequence ...... 25 3.5 Saproxylic beetle communities and habitat attributes along the chronosequence .. 25 4 Discussion ...... 26 Acknowledgments ...... 31 References cited ...... 42 Appendix A1 Species list and number of beetles captured for each postfire year in black spruce boreal forest ...... 49

vii CHAPITRTE 2 - EFFECTS OF SALVAGE LOGGING ON SAPROXYLIC BEETLE SUCCESSION ALONG A 12 YEARS POSTFIRE CHRONOSEQUENCE IN THE COMMERCIAL BOREAL FOREST ...... 51 Résumé ...... 52 Abstract ...... 53 1 Introduction ...... 54 2 Methods ...... 56 2.1 Study area ...... 56 2.2 Stand description ...... 57 2.2 Beetle sampling ...... 58 2.4 Statistical analysis ...... 59 2.4.1 Habitat attributes ...... 59 2.4.2 Beetle abundance and species richness ...... 59 2.4.3 Beetle assemblages in residual versus salvaged stands over time ...... 59 2.4.4 Residual versus salvage stands associated beetles ...... 60 3 Results ...... 61 3.1 Overview ...... 61 3.2 Effects of salvage logging on habitat attributes structure along the chronosequence ...... 61 3.3 Effects of salvage logging on beetle abundance and species richness along the chronosequence ...... 61 3.4 Effects of salvage logging on beetle assemblages along the chronosequence ...... 62 3.5 Beetles associated with residual or salvaged stands along the chronosequence .... 62 4 Discussion ...... 63 4.1 Effects of salvage logging on saproxylic beetle communities along the chronosequence ...... 63 4.2 Saproxylic beetles associated exclusively with salvaged stands ...... 66 4.3 Effect of salvage logging on elaterids’ association along the chronosequence ..... 66 5 Conclusion ...... 67 Acknowledgments ...... 67 References cited ...... 75 Appendix A2 Species list and number of beetles captured for each treatment of each postfire year in black spruce boreal forest ...... 80

CONCLUSION GÉNÉRALE ...... 83 Dynamique temporelle après feu ...... 83 Aménagement écosystémique des brûlis ...... 85 Investigations futures ...... 88 Références citées ...... 90

viii Liste des tableaux

Table 1. 1 Habitat attributes (average ± standard error) measured in postfire black spruce stands along a 15 years chronosequence...... 33 Table 1. 2 PERMANOVA and PERMDISP results showing the effect of time since wildfire on beetle assemblages. Pairwise comparisons between years, using t-tests, are shown for PERMANOVA and PERMDISP...... 34 Table 1. 3 Beetle species significantly associated (p < 0.05) with a postfire year or a combination of postfire years from point-biserial group-equalized phi coefficient analysis (Pearson correlation). Number of individuals in bold refers to the total number of specimens captured for each year with which a species was positively associated. Code species column refers to the abbreviations of the full name species and those codes were used in the RDA instead of the full name species...... 35

Table 2. 1 PERMANOVA results showing the effect of salvage logging, time since disturbance and their interaction on the beetles’ assemblages. Significant effects are outlined in bold (α = 0.05)...... 69 Table 2. 2 PERMANOVA results showing the by-year effects of salvage logging on postfire beetle assemblages. Significant effects are outlined in bold (α = 0.05)...... 70 Table 2. 3 Beetle species significantly associated (p <0.05) with residual or salvaged stands of a postfire year or a combination of postfire years from point-biserial group- equalized phi coefficient analysis (Pearson correlation). Number of individuals in bold refers to the total number of specimens captured for each treatment and year with which a species was positively associated...... 71

Liste des figures

Figure 1. 1 Study area showing selected burns with their respective time (year) since wildfire. X represent burns with a surface area less than 275 hectares...... 36 Figure 1. 2 ANOVA’s results comparing % bark covering on trunks of snags along the chronosequence (average ± standard error). Letters represent statistical differences at α < 0.05...... 37 Figure 1. 3 Average (± se) water content (%) along a 15 years postfire chronosequence in black spruce boreal forest for A) snags, B) DDW and C) stumps. Different letters indicate significant differences at α < 0.05...... 38 Figure 1. 4 Comparison of the water content (average ± se in %) between black spruce DDW, stumps and snags for each year, separately, along a 15 years postfire chronosequence in the boreal forest. Letters indicate significant differences at α < 0.05...... 39 Figure 1. 5 Average (± se) abundance (A) and species richness (B) of beetles along a 15 years postfire chronosequence in black spruce boreal forest. Letters indicate signigicant differences at α < 0.05. Anova’s were done on fourth root-transformed values for the abundance...... 40 Figure 1. 6 RDA ordination of beetles caught in postfire black spruce stands of the boreal forest. Centroids of stands of each postfire time since disturbance are illustrated by the number of years in bold. Species are marked with a cross while habitat attributes are illustrated with arrows in bold. All species were used in the analysis but only those which were significantly associated with specific postfire years (or a combination of several ones) in the species-habitat associations analysis (Pearson’s point-biserial correlation; see table 3) were illustrated on the ordination. Also, only significant habitat attributes are shown: basal area of black spruce snags and volume of black spruce DDW of each decay class. The code species, presented in table 3, were used instead of the full name species. Habitat attribute codes were: R. sna. = recent snag, M. sna. = medium snag, O. sna. = old snag, R. DDW = recent down deadwood, M. DDW = medium down deadwood and O. DDW = old down deadwood...... 41

Figure 2.1 Study area showing selected burns with their respective time (year) since wildfire. The X mark represents a 15 year burn that was smaller than 275 hectares. ... 72 Figure 2.2 Student paired t-test’s results comparing A) the basal area (m²/ha) of snags (average ± standard error), B) the volume (m³/ha) of DDW (average ± standard error) and C) the basal area of dead saplings (average ± standard error) between residual and salvaged stands for each postfire years. Values in bold indicate significant results (p < 0.05)...... 73 Figure 2.3 Student paired t-test’s results comparing A) the abundance (average ± standard error) and B) the species richness (average ± standard error), between residual and salvaged stands for each number of years after fire. Values in bold indicate significant results (p < 0.05)...... 74

Remerciements

Tout d’abord, je tiens à remercier mon directeur de recherche, le Dr Éric Bauce pour la liberté et la latitude qu’il m’a accordées lors de la réalisation de mes travaux. Je dois exprimer ma gratitude à Éric ainsi qu’à mon co-directeur, le Dr Jacques Ibarzabal, pour l’aide financière obtenue, laquelle a permis la réalisation de cette maîtrise. Je tiens aussi à leur présenter mes sincères remerciements pour m’avoir permis de présenter mes résultats de recherche à diverses occasions, notamment lors de congrès scientifiques tel que le Symposium International sur la dynamique et les services écologiques du bois mort dans les écosystèmes forestiers. La participation à ces congrès m’a permis d’évoluer grandement au niveau de ma formation académique et professionnelle.

Merci infiniment à mon co-directeur, le Dr Christian Hébert du laboratoire sur l’écologie et la diversité des insectes forestiers (ÉcoDIF). Christian est l’un des chercheurs en entomologie les plus passionnés que j’ai pu rencontrer lors de ma formation. En plus de m’avoir permis d’effectuer mes travaux de maîtrise dans son laboratoire, celui-ci m’a constamment fait bénéficier de ses larges connaissances au cours de nombreuses discussions à saveur scientifique. Merci pour ses judicieux conseils et pour son esprit critique. Je dois aussi souligner son dévouement, son ouverture d’esprit et sa grande disponibilité offerte lors de la direction de ma maîtrise. Christian est certainement un modèle d’inspiration pour l’ensemble des étudiants qui évoluent à ses côtés.

J’adresse aussi mes remerciements au coordonnateur du consortium iFor, le Dr Richard Berthiaume, pour son encadrement, son soutien et ses multiples implications lors de ma maîtrise. C’est toujours pertinent, constructif et quelquefois inévitable (!) de discuter de différents sujets avec Richard. Un grand merci aussi à Georges Pelletier, taxonomiste au Centre de Foresterie des Laurentides, pour ses formations d’identification d’insectes qui ont grandement facilité et accéléré cette étape de ma maîtrise. Je dois également remercier Yves Dubuc, technicien du laboratoire EcoDIF, pour sa disponibilité et son aide sans limite pour les travaux de terrain et de laboratoire.

xiii Une attention particulière à mon ami de longue date, Jean-Philippe Légaré qui m’a permis de mieux connaître l’entomologie forestière en m’engageant comme auxiliaire de recherche lors de mon baccalauréat. L’expérience acquise à ce moment a constitué une source d’inspiration pour cette maîtrise. J’exprime également ma gratitude à mon généreux collègue et ami, Jonathan Boucher, pour sa disponibilité, son esprit critique et pour m’avoir épaulé tout au long de mon parcours. Merci à tous mes collègues qui ont contribué aux travaux de laboratoire et un merci exclusif aux membres du laboratoire EcoDIF de Christian Hébert : Sébastien Bélanger, Olivier Norvez, Yannick Cadorette-Breton, Francis Desjardins et Ermias T. Azeria pour les moments inoubliables que nous avons passés ensemble.

Finalement, je remercie infiniment ma famille; mes parents et leur conjoint respectif pour leur aide et leur confiance, ma sœur et mon frère pour leur amour et leurs encouragements lors de mon cheminement. Un merci très spécial à mon amour, Chantal, pour avoir été ma grande complice ainsi que d’avoir été ma source de motivation et de persévérance pendant ce processus académique et professionnel qui comporte parfois des moments moins enthousiastes. Merci à vous tous d’avoir rendu tant de choses possibles!

xiv Avant-Propos

Ce mémoire de maîtrise est constitué de deux chapitres « Saproxylic beetle succession along a 15 years postfire chronosequence in the commercial boreal forest » et « Effects of salvage logging on saproxylic beetle succession along a 12 years postfire chronosequence in the commercial boreal forest » rédigés en anglais et sous la forme d’articles scientifiques destinés à être publiés dans des journaux scientifiques au cours de l’année 2013. Ces deux chapitres sont précédés d’une introduction et suivis d’une conclusion écrites en français. Le candidat à la maîtrise a participé à la sélection des sites ainsi qu’à l’installation du dispositif d’échantillonnage. Il a également effectué l’échantillonnage et la description des attributs de l’habitat dans les sites, identifié la plupart des insectes étudiés et fait l’analyse statistique des données. Les deux chapitres furent écrits entièrement par le candidat à la maîtrise, encadré et soutenu par les commentaires et conseils des coauteurs Christian Hébert, Jacques Ibarzabal, Richard Berthiaume et Éric Bauce. Ainsi, l’étudiant a le statut d’auteur principal pour les deux articles paraissant dans ce mémoire. Cette étude a été financée par le Fonds Québécois de la Recherche sur la Nature et les Technologies (FQRNT) dans le cadre du Programme de recherche en partenariat sur l’aménagement et l’environnement forestiers-II, par le consortium de recherche iFor (Université Laval) et le Service Canadien des Forêts de Ressources Naturelles Canada (SCF-NRCan).

Introduction générale

Les écosystèmes de la forêt boréale sont régis par des perturbations naturelles telles que des feux de forêt, des épidémies d'insectes ou des chablis (Pothier, 2001; Gauthier et al., 2008). Ces évènements peuvent se produire à grande échelle et sont d’ailleurs les principales causes de la diversité du paysage en forêt boréale. Ils y modifient la structure forestière en créant, dans le temps, des mosaïques d’arbres de dimensions, d’âges et d’essences différentes (Gauthier et al., 2008).

Les feux de forêt

Les feux de forêt constituent une perturbation majeure en forêt boréale québécoise (Bergeron et al., 2001; Gauthier et al., 2008). En effet, en 2010, près de 257 000 hectares (ha) de forêt y ont brûlé (SOPFEU, 2010) alors que la moyenne annuelle est d’environ 56 000 ha (moyenne couvrant la période de 1973 à 2004, (MRNF, 2012b)). Bien que les feux de forêt ont longtemps été considérés comme des « désastres naturels » produisant des déserts biologiques, plusieurs études ont récemment mis en évidence l’importance de cette perturbation pour la dynamique et les successions des écosystèmes forestiers (Bergeron et al., 2001; Pausas et Keeley, 2009; Drapeau et al., 2010; Nappi et al., 2011). Puisque les feux de forêt peuvent brûler de grandes superficies, ils sont considérés comme étant parmi les principales causes de changements, qui sont à leur tour essentielles à la diversité, au renouvellement et à la survie des écosystèmes boréaux (Fernández Fernández et Salgado Costas, 2004; Gauthier et al., 2008; Bowman et al., 2009; Nappi et al., 2011). Le feu permet, notamment, l’établissement rapide d’une nouvelle succession végétale composée principalement d’espèces ayant la capacité de germer et de croître rapidement et facilement après le passage du feu (Pausas et al., 2004). En effet, les essences retrouvées dans les régions où la fréquence des feux est élevée se sont adaptées et ont évolué pour devenir plus résistantes et résilientes (Pausas et Keeley, 2009). À titre d’exemples, certains pins et chênes pourraient être considérés plus résistants que d’autres essences dû à leur écorce plus épaisse, ce qui peut empêcher la mortalité des tissus internes (cambium et phloème) de l’arbre (Deslauriers et al., 1996; Mauri Ortuno et al., 2009). Le pin blanc (Pinus strobus L.), grâce à un élagage naturel efficace, possède une discontinuité verticale entre la strate

1 de végétation inférieure et le houppier, ce qui fait que les cônes sont souvent protégés des flammes (Mauri Ortuno et al., 2009). Le pin gris (Pinus banksiana Lamb.) et l’épinette noire (Picea mariana (Mill.)) quant à eux possèdent des cônes sérotineux et semi- sérotineux, respectivement, dont l’ouverture et la libération des semences sont favorisées par l’exposition à la chaleur (Deslauriers et al., 1996). Ainsi, ces trois essences ont une plus grande capacité de recolonisation et initient rapidement une nouvelle succession végétale après le passage du feu.

Le bois mort : un élément clé de la diversité biologique

À l’échelle du paysage, les feux de forêt constituent une importante source de bois mort ou moribond en forêt boréale (Nappi et al., 2004; Kennedy et Fontaine, 2009). La quantité, la qualité et la diversité du bois mort sont des éléments clés pour le maintien d’une grande diversité d’organismes saproxyliques (Saint-Germain et al., 2004b; Drapeau et al., 2010; Nappi et al., 2010). Par définition, les organismes saproxyliques nécessitent, pendant au moins une partie de leur cycle vital, du bois mort ou moribond, ou des champignons du bois, ou encore la présence d’autres organismes saproxyliques (Speight, 1989). Les chicots sur pied, les débris ligneux au sol, les souches, les branches et les racines grossières représentent divers types de bois mort ou moribond retrouvés à la suite d’un feu (Harmon et al., 1986). Ceux-ci peuvent avoir subi diverses intensités de brûlage, ce qui résulte en une sévérité du feu variable créant, à l’échelle du paysage, une distribution et une variabilité complexe de la ressource bois mort. Selon la composition forestière initiale, un feu de grande superficie laisse des peuplements brûlés et non-brûlés de composition d’essences différentes et de structures d’âges variés, ce qui accroît l’hétérogénéité de la distribution de cette ressource (Nappi et al., 2011). De plus, au fil du temps après un feu, les chicots se cassent et tombent pour devenir des débris ligneux au sol (Ausmus, 1977; Swift, 1977a-b; Harmon et al., 1986). La densité du bois diminuera graduellement durant tout le processus de décomposition alors que le contenu en eau diminuera pendant le stade de chicot et augmentera par la suite, lorsque les chicots tomberont au sol et deviendront des débris ligneux (Paletto et Tosi, 2010). Par conséquent, à la suite d’un feu, le bois mort procure une différenciation temporelle de niches écologiques (Simandl, 1993) caractérisées par différentes qualités nutritives (Tinker et Knight, 2000).

2 Cette diversité spatio-temporelle du bois mort, ainsi que les conditions spécifiques observées à la suite d’un feu (augmentation de la température, diminution de la compétition inter-spécifique, une couche organique amincie et un sol minéral parfois exposé (Wikars, 1992; Nappi et al., 2004)) procurent une variété de conditions écologiques intéressantes pour plusieurs communautés végétales et animales (Butts et McComb, 2000; Nappi et al., 2003, 2004; Similä et al., 2003), particulièrement pour les insectes (Siitonen, 2001). Au fur et à mesure que le temps passe après le passage du feu, les communautés se modifient de manière relativement synchrone avec les changements des attributs de l’habitat (Nappi et al., 2011). Ainsi, l’hétérogénéité spatio-temporelle des peuplements brûlés est primordiale pour le maintien de la diversité et les successions naturelles qui y sont associées.

Importance des feux pour les coléoptères saproxyliques

La particularité des feux de forêt, comparativement aux autres perturbations naturelles, est leur capacité à générer instantanément une grande quantité de bois mort sur de grandes superficies (Tinker et Knight, 2000). Cette abondance soudaine d’arbres tués par le feu constitue des habitats de qualité pour plusieurs espèces d’insectes, notamment les coléoptères saproxyliques (Wikars, 1992; Saint-Germain et al., 2004c). Plusieurs familles de coléoptères utilisent le bois récemment brûlé pour compléter leur cycle vital et accroître leur population, ce qui permet à plusieurs espèces de persister dans le paysage (Jonsell et al., 1998). En raison de la grande quantité de bois mort retrouvé dans un brûlis, plusieurs espèces d’insectes seraient fortement associées aux brûlis (Saint-Germain et al., 2004c; Boucher et al., 2012). Certaines auraient même développé des adaptations permettant de détecter des composés volatiles émanant de la fumée ou les ondes infrarouges émises par la chaleur afin d’atteindre les brûlis plus facilement (Evans, 1966; Evans et Kuster, 1980; Schütz et al., 1999). Des espèces pyrophiles auraient donc évolué dans le contexte des feux de forêt et en seraient dépendantes (Nappi et al., 2004).

Les conditions générées pendant et immédiatement après un feu de forêt sont favorables à une phase de colonisation initiale par des coléoptères saproxyliques (Boulanger et Sirois, 2007), particulièrement pour ceux qui se retrouvent à proximité des brûlis ainsi que pour ceux qui possèdent la capacité de se déplacer sur de longues distances et de détecter

3 rapidement les brûlis afin de s’y diriger. Parfois, certaines espèces sont retrouvées dans les brûlis avant même que le feu ne soit éteint (Evans, 1966). C’est d’ailleurs le cas du Buprestidae Melanophila acuminata DeGeer, lequel détecte les ondes infrarouges ce qui lui permettrait de faire partie de la phase initiale de colonisation, en arrivant très tôt et en pondant ses œufs à la base des conifères alors que le peuplement brûle toujours (Evans, 1966). Le bois récemment brûlé procure des sites de pontes privilégiés permettant le développement d’espèces xylophages et sous-corticales puisque la valeur nutritive du cambium est alors peu affectée (Saint-Germain et al., 2004c; Boulanger et Sirois, 2007). De plus, la prolifération de champignons, majoritairement de type ascomycete, sont une source de nourriture de prédilection pour plusieurs espèces mycophages, dont la famille des Lathridiidae (Muona et Rutanen, 1994; Boulanger et Sirois, 2007). À noter que la composition forestière avant feu (essences, diamètre, hauteur, densité du peuplement, quantité de débris ligneux au sol, etc.) et la sévérité du feu s’avèrent être d’importants facteurs qui diversifient spatialement les assemblages d’espèces de la phase initiale (Hanks, 1999; Nappi et al., 2010; Boulanger et al., 2011).

Au fil du temps après un feu, d’autres phases de colonisation par les coléoptères saproxyliques se succéderont suivant les changements graduels des attributs de l’habitat brûlé (Boulanger et Sirois, 2007; Nappi et al., 2010). Les chutes de chicots au sol, jumelées à l’augmentation de l’humidité dans les débris ligneux nouvellement générés, entraîneront des changements graduels dans les assemblages d’espèces de coléoptères saproxyliques (Boulanger et Sirois, 2007; Nappi et al., 2010). Des communautés spécifiques seront également attirées par des mycéliums et des carpophores qui se développeront sur le bois alors que l’habitat sous-cortical disparaîtra progressivement, dû à la perte de l’écorce (Kaila et al., 1997; Rukke, 2002; Jonsell et Weslien, 2003). Ainsi, des coléoptères saprophages et mycophages spécifiques succéderont à la phase initiale de colonisation au fur et à mesure que le bois brûlé se décomposera et que ses propriétés initiales se modifieront (Boulanger et Sirois, 2007). De plus, certains coléoptères, se nourrissant des racines d’arbres de plantes ou de graminées ou directement des jeunes plants, seront présents lors de la régénération de la végétation (Muona et Rutanen, 1994).

4 Problématique

L’exploitation de la matière ligneuse en forêt boréale génère environ 60 % de l’activité économique de l’industrie forestière du Canada (Burton et al., 2003). Le Québec est un acteur de taille puisque 20 % des forêts canadiennes s’y retrouvent et 70 % de celles-ci sont productives commercialement (MRNF, 2010). Ainsi, les feux de forêt peuvent engendrer des pertes économiques importantes en termes de matières ligneuses si le bois brûlé n’est pas récupéré. Conséquemment, certains gouvernements provinciaux du Canada, notamment celui du Québec, favorisent, dans leurs politiques sur les forêts, la récupération du bois brûlé par l’entremise de plans spéciaux d’aménagement. Au Québec, les volumes de bois récolté dans les brûlis ont augmenté depuis les années 90 (Purdon et al., 2002; Saint- Germain et Greene, 2009), et les récentes diminutions de la possibilité forestière, annoncées en 2004 (Coulombe et al., 2004), ont probablement contribué à cette augmentation. À titre d’exemple, en 2005, environ 20 % du volume total de bois récolté dans les forêts publiques du Québec était issu des brûlis, soit près de 6.2 millions de mètres cubes (m³) (Parent, 2008; MRNF, 2012a). Selon la loi québécoise, les plans spéciaux peuvent déroger des normes d’interventions émises par le gouvernement et régissant la récolte des forêts vertes ou non- perturbées. À titre d’exemple, jusqu’à tout récemment, la récolte des bois brûlés et non- brûlés à l’intérieur d’un brûlis se faisait par coupes avec protection de la régénération et des sols (CPRS), récoltant toutes les tiges marchandes (DHP ≥ 9.1 cm) accessibles (Purdon et al., 2002; Nappi et al., 2004). Cette intensification du double régime de perturbation soit l’interaction du feu et de la coupe de récupération, s’avère de plus en plus inquiétante en termes de conservation (Lindenmayer et al., 2008; Boucher, 2011) et pourrait avoir des conséquences à grande échelle dont le dépassement de l’intervalle de variabilité naturelle (Bouchard et Munson, 2009). D’ailleurs, il a été démontré que la perte de certains attributs des forêts brûlées, causés par la récupération, avait des impacts négatifs sur la faune et la flore (Nappi et al., 2004), notamment pour les successions écologiques naturelles (Lindenmayer et Ough, 2006). De surcroît, la récupération après feu pourrait même excéder la résilience de certaines espèces fauniques qui auraient évolué et qui se seraient adaptées au régime des feux (Le Goff et al., 2008; Boucher, 2011). Il est d’ailleurs à propos de souligner que la réduction du volume de bois mort, causée par l’aménagement forestier

5 intensif (Siitonen, 2001) et la suppression soutenue des feux de forêt (Wikars, 1992), a eu pour conséquence de rendre les coléoptères saproxyliques parmi les organismes les plus à risque d’extinction en Fennoscandie (Hyvärinen et al., 2006).

Impacts des coupes de récupération sur la faune et la flore

Le prélèvement des chicots et des arbres mourants à l’intérieur d’un brûlis diminue grandement l’abondance des insectes saproxyliques au stade larvaire, ce qui a des impacts importants sur le nombre de proies disponibles pour certains oiseaux insectivores (Morissette et al., 2002; Purdon et al., 2002). Ces impacts sont d’autant plus importants puisqu’ils surviennent dans les premières années après feu, là où l’abondance des insectes saproxyliques est la plus élevée (Boulanger et Sirois, 2007). De plus, une grande partie des sites de nidification, des gîtes, des lieux de reproduction et de chasse (ex. perchoir) pour les oiseaux et les mammifères sont éliminés (Nappi et al., 2011). La soustraction de ces tiges a également des impacts sur le potentiel de régénération des peuplements récupérés, puisque les réserves de graines semencières (sources in situ) sont extraites du territoire (Greene et Johnson, 2000; Gauthier et al., 2008). Cela est davantage problématique pour l’épinette noire dont les graines sont libérées plus tardivement comparativement au pin gris (Greene et Johnson, 1999). Cependant, même si les semences du pin gris sont libérées dans les premiers mois après le passage du feu, celles-ci ou les semis peuvent être endommagés par le passage successif de la machinerie ou encore, peuvent être défavorisés par un assèchement plus rapide des sols et une altération des lits de germination (Purdon et al., 2002; Nappi et al., 2011). La récolte des îlots non brûlés peut également avoir des conséquences négatives pour les essences moins adaptées aux conditions des brûlis (ex. le mélèze laricin (Larix laricina (Du Roy)), le sapin baumier (Abies balsamea (L.)) et l’épinette blanche (Picea glauca (Moench))) lesquelles nécessitent la présence de semenciers pour se régénérer (Galipeau et al., 1997; Greene and Johnson, 2000). Ces îlots servent également de refuges et sont donc importants pour les espèces animales moins adaptées aux nouvelles conditions du brûlis (Greene et Johnson, 2000).

La récolte du bois dans un brûlis réduit la quantité et la qualité du bois mort au sol à moyen terme puisque les chicots qui devraient tomber au sol dans les années suivant la

6 perturbation sont retirés peu de temps après le feu. Cela entraîne des conséquences irréversibles pour les communautés associées à cet habitat et ce, à long terme (Boulanger et Sirois, 2007; Nappi et al., 2011). La succession naturelle des organismes impliqués dans le recyclage des éléments nutritifs est donc rompue en partie, ce qui pourrait entraîner des répercussions sur le renouvellement de l’écosystème (Speight, 1989; Cobb et al., 2010). En effet, il a été démontré que les larves d’insectes xylophages jouent un rôle important dans le recyclage des nutriments du sol engendrant des contrecoups positifs sur la croissance des plantes colonisatrices des milieux brûlés (Cobb et al., 2010).

Développement durable et aménagement écosystémique

Depuis le début des années 90, on considère davantage le concept de conservation de la diversité biologique en tant que critère important dans l’atteinte des objectifs d’aménagement durable des forêts (CCMF, 1995, 1997; CSNF, 1998; CCMF, 2003; CSNF, 2003). C’est lors de la Conférence des Nations Unies sur l'environnement et le développement tenue à Rio de Janeiro en 1992, connue sous le nom de « Sommet de la Terre », que l’on a reconnu l’importance de la gestion durable des forêts, et duquel a découlé l’adoption de plusieurs principes d’aménagement durable (CCMF 2003). Cette « Déclaration de principe sur les forêts » a imposé des changements dans nos pratiques sylvicoles dans le but de bénéficier d’une économie soutenue à long terme tout en assurant la protection de l’environnement et le maintien des fonctions des écosystèmes. Depuis le début des années 2000, le concept d’aménagement forestier écosystémique suscite un grand intérêt au sein de la communauté scientifique de même qu’auprès des gestionnaires forestiers (Grenon et al., 2010). En adhérant à ce concept, l’aménagement forestier doit refléter l’hétérogénéité engendrée par les perturbations naturelles à l’échelle du peuplement ainsi qu’à l’échelle du paysage. Ce concept vise la protection, la conservation et la mise en valeur des ressources, trois composantes qui sont responsables du virage dans la gestion des ressources naturelles. Cette nouvelle approche vise à assurer la pérennité des ressources forestières selon les trois volets du développement durable soit économique, social et environnemental (Coulombe et al., 2004). Plus précisément, l’aménagement écosystémique vise : « […] à maintenir des écosystèmes sains et résilients en misant sur une diminution des écarts entre les paysages naturels et ceux qui sont aménagés afin d’assurer, à long

7 terme, le maintien des multiples fonctions de l’écosystème et, par conséquent, de conserver les bénéfices sociaux et économiques que l’on en retire » (Gauthier et al., 2008).

Lors des dernières années, quelques suggestions ont été faites afin que le Ministère des Ressources Naturelles et de la Faune (MRNF) établisse des lignes directrices lors de la confection des plans spéciaux d’aménagement. Ces suggestions avaient pour but d’assurer le maintien de la biodiversité et des processus naturels associés aux brûlis tout en favorisant une rentabilité économique durable des activités de récoltes après feu (Nappi et al., 2011). Ce n’est qu’en 2010, à la suite d’un projet pilote mené dans un brûlis de l’île René- Levasseur, au Québec, que les premiers plans à caractère écosystémique ont été mis en application au Québec. Les plus récentes suggestions présentées pour la province de Québec sont parues dans le document de Nappi et al. (2011) intitulé « La récolte dans les forêts brûlées : Enjeux et orientations pour un aménagement écosystémique ». Les grandes lignes de ces orientations sont le maintien des peuplements brûlés sur une surface de l’ordre de 30 % des superficies brûlées à l’échelle d’une unité d’aménagement et de 15 % à l’échelle d’un brûlis dans le but de préserver la variabilité naturelle de la biodiversité et des processus écologiques associés à cette perturbation naturelle. La superficie de forêt brûlée résiduelle, faisant partie du 30 % à conserver, devra être réévaluée à tous les cinq ans afin d’assurer une démarche d’aménagement adaptative intégrant les nouvelles connaissances sur les effets du feu et des coupes de récupération.

Contexte de l’étude

Afin d’évaluer les effets de la récupération des brûlis sur l’intégrité du milieu et sur les successions naturelles après feu, nous suggérons d’utiliser les communautés d’arthropodes et, plus précisément, les insectes saproxyliques comme indicateurs. Les arthropodes sont une composante majeure des écosystèmes et jouent un rôle capital dans leur fonctionnement (Kim, 1993). Bien que les insectes représentent plus de 80 % du règne animal, ils forment toutefois le groupe le moins étudié sur la planète (Samways, 1993). Toutefois, les coléoptères constituent l’ordre d’insectes le mieux connu et le plus diversifié, tant au plan écologique que taxinomique. Ils sont donc plus faciles et moins coûteux à échantillonner et constituent un groupe indicateur à fort potentiel pour évaluer l’impact des

8 stress anthropiques sur le milieu (Kim, 1993). La grande quantité de bois mort contenue dans les brûlis fait en sorte que les coléoptères saproxyliques sont les insectes les plus abondants et diversifiés qui colonisent ce type d’habitat (Wikars, 2002; Saint-Germain et al., 2004a; Boulanger et Sirois, 2007; Boulanger et al., 2010). Dans le cadre de cette étude, c’est donc sur ce groupe d’organismes que nous avons choisi de travailler.

Plus précisément, nous étudions les successions naturelles de huit familles de coléoptères saproxyliques (Buprestidae, Cerambycidae, Cleridae, Corylophidae, Curculionidae, Elateridae, Lathridiidae et Salpingidae) pour ensuite décrire les effets des coupes de récupération sur ces successions dans les brûlis d’épinettes noires de la forêt boréale du Québec. Ces huit familles ont été sélectionnées sur la base de leur forte association avec les peuplements brûlés (Saint-Germain et al., 2004c; Boulanger et Sirois, 2007; Boucher et al., 2012). Non seulement ces huit familles renferment des espèces qui sont presqu’exclusivement trouvées dans les brûlis, mais elles s’y retrouvent en grande abondance comparativement à d’autres types d’habitats contenant du bois mort (Boucher et al., 2012). Les coupes de récupération pourraient donc avoir des effets négatifs à long terme sur le maintien de ces espèces dans le paysage. De plus, en termes de rôles écologiques, ces huit familles ont des fonctions écologiques diversifiées (Grove, 2002; Saint-Germain et al., 2004c; Boulanger et Sirois, 2007; Boucher et al., 2012). En effet, les larves de Buprestidae et de Cerambycidae sont xylophages et se développent directement dans le phloème ou dans le xylème du bois récemment mort ou moribond (Arnett Jr. et al., 2002), alors que les Scolytinae (Curculionidae) sont sous-corticaux et se développent en se nourrissant sous l’écorce (Bright Jr., 1976). Les autres espèces de la famille des Curculionidae sont plutôt polyphages et leurs larves se nourrissent principalement de racines ou de pousses des jeunes plants (Arnett Jr. et al., 2002). Les Cleridae et les Salpingidae sont des prédateurs d’espèces saproxyliques (Muona et Rutanen, 1994; Arnett Jr. et al., 2002), alors que les Corylophidae et les Lathridiidae sont des mycophages saproxyliques (Arnett Jr. et al., 2002). Certaines espèces d’Elateridae seraient reconnues pour être saprophages creusant des galeries dans le bois davantage décomposé et pourri par les champignons de carie, d’autres seraient prédatrices de larves d’espèces saproxyliques dans le vieux bois mort et quelques-unes se nourriraient de racines de plantes en régénération ou d’arbres matures (Muona et Rutanen, 1994; Du Chatenet, 2000).

9 Plusieurs espèces des huit familles sélectionnées sont aussi connues pour être associées avec les brûlis récents (0 à 3 ans après feu) de la zone circumpolaire (Alaska, Canada, Fennoscandie, Suisse et Russie) (Wikars, 1992; Muona et Rutanen, 1994; Wikars, 1994, 1997a; Siitonen, 2001; Saint-Germain et al., 2004c; Boulanger et Sirois, 2007; Moretti et al., 2010; Boucher, 2011). Cependant, peu d’études ont évalué les successions de ces coléoptères à moyen terme (3 à 15 ans après feu) et aucune, à notre connaissance, n’a étudié les impacts des coupes de récupération sur ces successions. Boulanger et Sirois (2007) ont déjà étudié les assemblages d’arthropodes saproxyliques d’une chronoséquence d’une durée de 29 ans après feu dans les peuplements d’épinettes noires se situant dans la zone non-exploitée du nord du Québec. Leurs résultats ont démontré deux phases de colonisation par les arthropodes saproxyliques. En effet, ils indiquent la présence d’une première phase débutant tôt après feu, surtout représentée par des espèces xylophages, sous-corticales et mycophages. La deuxième phase surviendrait environ dix ans après feu, lorsque les chicots tombent au sol générant de nouveaux débris ligneux, entraînant ainsi l’arrivée de nouvelles espèces associées à ce type de bois mort, tels des saprophages et certains prédateurs. Notre étude utilise une chronoséquence plus fine que celle de Boulanger et Sirois (2007), avec six classes d’âges le long d’une chronoséquence de 15 ans après feu.

En outre, ce qui rend notre étude particulièrement unique est le fait d’avoir choisi de travailler au niveau des souches d’épinettes noires dans les brûlis se situant dans la zone exploitée de la forêt boréale du Québec. Dans cette zone, se situant sous le 52ième parallèle, les arbres sont plus gros par rapport à la forêt non-commerciale du nord du Québec. Quant aux souches, elles constituent un micro-habitat ayant des propriétés (température et humidité) différentes et plus stables que celles de la partie plus aérienne des chicots, soit le tronc (Abrahamsson et Lindbladh, 2006). La plus grande stabilité temporelle des souches en tant qu’habitat, comparativement aux chicots, vient certainement du fait que les chicots finiront par tomber au sol pour devenir des débris ligneux, alors que les souches resteront généralement en place, sauf s’il y a un déracinement par chablis. Au Québec, très peu d’études se sont penchées sur la valeur écologique des souches en tant que micro-habitat potentiel pour les coléoptères saproxyliques, même après coupe en forêt verte. Dans cette étude, nous évaluons le taux d’activité des coléoptères saproxyliques au niveau des souches

10 à l’aide d’un échantillonnage fait avec des pièges à impacts troncaux. Ce type de piège est reconnu pour être efficace dans la capture des coléoptères saproxyliques susceptibles d’utiliser le substrat sur lequel le piège est installé pour se reproduire et compléter leur cycle vital (Kaila, 1993).

Un premier volet de l’étude vise à décrire les successions naturelles après feu pour les coléoptères saproxyliques. Plus spécifiquement, nous avons d’abord comme objectifs 1) de vérifier si les souches constituent un micro-habitat plus stable que les chicots ou les débris ligneux au sol 2) de décrire la succession post-feu des coléoptères saproxyliques au niveau des souches et le long d’une chronoséquence de 15 ans (abondance, richesse en espèces, assemblages d’espèces) et 3) de déterminer combien de phases distinctes de colonisations sont détectables le long de cette chronoséquence. Ensuite, un deuxième volet vise à déterminer les impacts que peut avoir la récupération du bois brûlé sur celles-ci et à évaluer la valeur des souches résiduelles en tant que mesure de mitigation des effets de cette pratique d’aménagement. Ainsi, les objectifs en lien avec ce volet sont : 1) de déterminer si les souches résiduelles à la coupe de récupération peuvent maintenir les communautés de coléoptères saproxyliques associées aux brûlis; 2) d’explorer les effets de la récupération après feu sur la dynamique temporelle de ces communautés et 3) de déterminer si il y a des espèces qui sont associées avec des peuplements non-récupérés exclusivement, pour plusieurs années consécutives. Il est important pour nous d’identifier de telles espèces puisqu’elles sont probablement celles dont les populations sont les plus susceptibles d’être affectées négativement par les coupes de récupération.

Chapitre 1

Saproxylic beetle succession along a 15 years postfire chronosequence in the commercial boreal forest

Olivier Jeffrey1, Christian Hébert2, Jacques Ibarzabal3, Richard Berthiaume1 and Éric

Bauce1

1 Université Laval, Faculté de foresterie, de géographie et de géomatique, Pavillon Abitibi-

Price, Québec, Québec, G1K 7P4, Canada

2 Natural Resources Canada, Canadian Forest Service, Laurentian Forestry Centre, 1055 du

P.E.P.S., P.O. Box 10380, Stn. Sainte-Foy, Québec, Québec, G1V 4C7, Canada

3 Université du Québec à Chicoutimi, 555 boulevard de l’Université, Chicoutimi, Québec,

G7H 2B1, Canada

13 Résumé

Ce chapitre vise 1) à vérifier si les souches représentent un micro-habitat plus stable que les chicots et/ou les débris ligneux au sol, 2) à décrire la succession après feu des coléoptères saproxyliques au niveau des souches et 3) à déterminer combien de phases distinctes de colonisation par ces coléoptères peuvent être détectées le long d’une chronoséquence de 15 ans après le passage du feu. Les coléoptères ont été inventoriés durant l’été 2010 dans des brûlis de 1995, 2003, 2005, 2007, 2009 et 2010. Un peu plus de 5 500 coléoptères répartis dans les 8 familles étudiées ont été capturés. Les résultats ont révélé une abondance et une richesse en espèces plus élevées lors des trois premières années après feu. De plus, quatre phases de colonisation par les coléoptères saproxyliques ont été détectées durant les cinq premières années après feu. La dynamique temporelle du bois mort semble influencer les assemblages d’espèces pendant ces cinq premières années.

Mots-clés : Biodiversité, Succession, Coléoptère saproxylique, Feu de forêt, Souche, Bois mort, Piège à impact troncaux

14 Abstract

This chapter aims 1) to verify if stumps represent a more stable habitat than either snags or logs, 2) to describe postfire succession of saproxylic beetles at the stump level and 3) to determine how many distinct succession phases we may detect along a 15 postfire years chronosequence. Beetles were sampled during summer 2010 from burns of 1995, 2003, 2005, 2007, 2009 and 2010. Slightly more than 5 500 beetles among the height studied families were captured. Results revealed higher abundance and species richness during the three first postfire years. In addition, four saproxylic beetles colonization phases were detected during the five first postfire years. Temporal dynamics of deadwood seems to influence species assemblages during these five postfire years.

Keywords: Biodiversity; Succession; Saproxylic Beetles: Wildfire; Stump; Deadwood; Trunk widow trap

15 1 Introduction

Wildfires constitute one of the most important natural disturbance in boreal forest (Bergeron et al., 2001; Gauthier et al., 2008). Several studies have highlighted the importance of wildfire on forest dynamics and ecosystem’s successions (Bergeron et al., 2001; Pausas and Keeley, 2009; Drapeau et al., 2010; Nappi et al., 2011). Indeed, several plant species have the ability to germinate quickly and grow rapidly after fire, thus initiating new successions (Pausas and Keeley, 2009). These plants have evolved in a context of wildfire and are recognized as fire-adapted species (Pausas and Keeley, 2009). Therefore, fire influences stand structure and vegetation composition and thus generates new ecological conditions for a large diversity of organisms (Bergeron et al., 2002).

Forest wildfire also constitutes an important source of dead wood in boreal forest (Tinker and Knight, 2000; Siitonen, 2001; Nappi et al., 2004; Kennedy and Fontaine, 2009). Compared with other types of disturbances like windthrows, insect outbreaks and harvesting, wildfires generates large amounts of recent dead wood all at once and over large areas (Tinker and Knight, 2000). The amount, quality and diversity of dead wood is a key component at the ecosystem level for maintaining a large diversity of saproxylic species (Bader et al., 1995; Christensen and Emborg, 1996; Siitonen, 2001; Humphrey et al., 2004), those that depend on deadwood (Speight, 1989). Several saproxylic insects are known to take advantage of the sudden abundance of newly-killed trees by fire (Saint- Germain et al., 2004c; Boucher et al., 2012) and some have developed adaptations to detect heat or smoke emitted by wildfires (Evans, 1966; Wikars, 1994; Schütz et al., 1999). As time elapses after fire, snags fall to the ground (Ausmus, 1977; Swift, 1977b; Swift, 1977a; Harmon et al., 1986). Indeed, while wood density decreases progressively from the first to the last decay stage, water content decreases during the snag stage (first decay stages) but increases afterwards when snags fall to the ground and become logs (Paletto and Tosi, 2010). Thereby, dead wood provides a temporal differentiation of ecological niches (Simandl, 1993) characterized by varying food quality (Tinker and Knight, 2000). This spatio-temporal variability in dead wood benefit to numerous organisms (Butts and McComb, 2000; Nappi et al., 2003; Similä et al., 2003; Nappi et al., 2004), specially

16 saproxylic insects (Siitonen, 2001). Therefore, wildfire also provides conditions to rapidly set up new saproxylic insect successions.

Most studies addressing postfire stand colonization by saproxylic insects in North America have focused on communities shortly after fire (Saint-Germain et al., 2004b-c; Boulanger et al., 2011; Azeria et al., 2011, 2012) and particularly on xylophagous species responsible for lumber degradation (Cerezke, 1977; Saint-Germain et al., 2004a; Saint-Germain et al., 2007). The only study looking at mid-term successions was carried out in black spruce (Picea mariana (Mill.)) forests of Northern Quebec (Boulanger and Sirois 2007). The authors recognized two phases in the successional pattern of saproxylic arthropods along a 29 years chronosequence after fire. The initial colonization begins during the year of the fire while the second phase occurs only when snags fall to the ground, usually after a decade (Boulanger and Sirois, 2007). The authors indicated that this succession pattern was rather slow compared with the one described by Esseen et al. (1992) in Scandinavian old- growth boreal forests. This could be related to the cold climate, which was typical of the lower subarctic region, the decomposition rates of postfire snags being exceptionally low in this region (Boulanger and Sirois, 2006). Moreover, postfire beetle communities of the Boulanger and Sirois (2006) study were poorer than in southern boreal forest where species richness was at least twice higher (Saint-Germain et al., 2004a-b).

In order to provide knowledge for improving management of postfire boreal forest of commercial size, we studied succession of eight beetle families (Buprestidae, Cerambycidae, Cleridae, Corylophidae, Curculionidae, Elateridae, Lathridiidae and Salpingidae) in 10 burns along a 15 years chronosequence. These beetle families contained saproxylic species (Grove, 2002; Saint-Germain et al., 2004c; Boucher et al., 2012) which are closely associated with recently burned black spruce forests (Boucher et al., 2012). We selected those families because they initiate successions and should be the most vulnerable to increasing postfire salvage logging activities (Boucher et al., 2012). Moreover, we sampled beetles using trunk window traps placed at the stump level where insects may still have access to dead wood resources, even after salvage logging. Stumps persist longer in burned stands than snags, which break and fall to the ground. Thus, stumps represent a more stable microhabitat than snags and probably show different properties (e.g.

17 temperature and water content (Abrahamsson and Lindbladh, 2006)). Our objectives were 1) to verify if stumps represent a more stable habitat than either snags or logs, 2) to describe postfire succession of saproxylic beetles at stump height, along a 15 years postfire chronosequence in the commercial boreal forest and 3) to determine how many distinct succession phases we may detect along the chronosequence.

2 Methods

2.1 Study area

The study area covered 80 000 km2 and belong to two bioclimatic subdomains of North- Western Quebec, Canada: the western spruce-moss and the northern balsam fir white birch subdomains (from 48°49’ to 50°43’ N and from 70°10’ to 75°51’W). The climate was characterized by an average annual air temperature varying between 0 and -2.5°C and annual precipitations ranging from 800 to 1 000 mm (Wilson, 1971; Grondin, 1996). This continental climate was characterized by fewer precipitations than in eastern Quebec and thus provided a shorter fire cycle ranging between 90 and 190 years (MRN, 2000; Boucher et al., 2003). Black spruce stands dominated the landscape and jack pine (Pinus banksiana Lamb.), balsam fir (Abies balsamea (L.) Mill.), trembling aspen (Populus tremuloides Michx.), paper birch (Betula papyrifera Marsch.) and tamarack (Larix laricina (Du Roi) K. Koch) were observed as companion species.

2.2 Stand description

We selected twenty-four black spruce stands among ten burns corresponding to 6 different years along a chronosequence that covered 15 years after wildfire: one burn occurred in 2010, three in 2009, two in 2007, one in 2005, one in 2003 and two in 1995 (Figure 1). Four replicate stands were established in each of the six years after burn. As the number of burns available for each category (years after burn) varied along the chronosequence, the number of replicate stands in each burn also varied, from one (2009 and 1995 burns) to four (2003, 2005 and 2010 burns). This was mainly due to the poor accessibility and availability of the burns. Selected stands varied between 60 and 100 years old and burned at moderate or high severity (i.e., all trees were dead at time of site selection). All wildfires occurred in

18 the late-spring/early-summer of the coinciding year. We loaded ecoforest and burn severity maps, provided by the Ministère des ressources naturelles (MRN), in ArcMap 9.2 (ESRI, 2009) to facilitate stand selection. Each replicate stand was distanced by at least 1.5 km to ensure data independence.

Stand attributes were described within a 400 m² circular plot. Tree species and diameter at stump height (DSH, i.e. 15 cm above root or ground level (MRNF, 2005)) were recorded for every snag > 9 cm diameter at breast height (DBH). The percentage of bark, for every 5 %, covering trunks of snags was also estimated. Tree species and DBH were recorded for every dead and live sapling (> 1.30 m height and ≤ 9 cm DBH) within a circular 40 m² plot centered on the 400 m² plot. Four 11.28 m transects were drawn from the plot center toward each cardinal points to estimate the volume (m³/ha) of down dead wood (DDW). Each DDW larger than 5 cm intercepted across the transects was identified when possible and its diameter was measured perpendicularly to the log at the intersection point (Warren and Olsen, 1964). The DDW volume was estimated according to the method of Van Wagner (1982). The decay stage of each snag, stump, sapling or DDW was determined on the basis of the Hunter (1990) classification which recognizes seven decay stages for snags and five for DDW. The Hunter (1990) classification relies largely on the presence of residual branches or twigs (non applicable for stumps), residual bark and dead wood texture. Thus, stump decay stages were determined with the DDW classification, using the residual bark coverage and wood texture as criteria. A correlation matrix between deadwood attributes was produced to identify collinearity problems (r > 0.7). As adjacent decay classes of deadwood were highly correlated, we combined them into three categories: young, medium and old dead wood. For snags, Hunter (1990) decay classes were combined as follow: young = class 3, medium = classes 4 and 5, and old = classes 6 and 7 (note that the old category refer to broken snags). For stumps and DDW, decay classes were combined for: young = classes 1 and 2, medium = class 3 and old = classes 4 and 5.

In order to compare water contents of different types of burned deadwood along the chronosequence, we cut six snags in each plot and collected a 5-cm thick disk at the stump level (45 cm above root or ground level) on three of them and at DBH level (1.30 m high) on the three other snags. This allowed complete independence between samples as they

19 were also compared with 5-cm thick disks collected on three different DDW. Disks were weighted (wet weight) and then oven dried at 65°C until weight stabilized which corresponded to the dry weight (requires a minimum of 36 hours). Water content was expressed as a percentage of wet weight and was calculated with the following formula (Akbulut and Linit, 1999):

wet.weight(g)  dry.weight(g) Water.content  x100 wet.weight

2.3 Beetle sampling

Three trunk-window traps (TWT), known for their efficiency to sample saproxylic beetles (Kaila, 1993; Boulanger and Sirois, 2007), were installed at the stump height (window just above the highest root) on three different snags of each stand. All traps were installed between May 20th and 30th of 2010 and were in operation until August 26th of 2010. Traps were distanced by a minimum of five meters and disposed in order to form an equilateral triangle. These TWT consisted of a 15 × 40 cm Plexiglas® panel installed vertically and perpendicularly to the stump in order to intercept beetles. A 15 cm diameter funnel was placed below the panel to lead beetles in a 500 ml collecting vial filled with 225 ml of a 40% ethanol and 5% household vinegar solution in order to kill and preserve insects. Samples were collected bi-weekly (six collections) and taken back to the Laurentian Forestry Centre (LFC, Quebec, Canada) where they were kept at 4 °C until sorting and identification.

As suggested by Boucher et al. (2012), we focused on beetles of eight beetle families closely associated with recently burned forests. Almost all beetles of these height families were identified at the species level and vouchers have been conserved in the Forest Insect Ecology and Diversity (EcoDIF) laboratory of LFC.

2.4 Statistical analysis

All beetle species identified were retained for analysis.

20 2.4.1 Habitat attributes along the chronosequence

One way ANOVAs were used to compare the basal area of black spruce snags, stumps and dead saplings, the basal area of living saplings of black spruce and jack pine, the volume of black spruce DDW (overall and per decay categories) and the percentage of bark recovering trunks of snags between years along the postfire chronosequence. Also, in order to verify if stumps represent a more stable habitat than either snags or logs, we compared the water content of the each type of deadwood along the chronosequence. We used one- way nested ANOVAs in which the estimates from the three disks were nested in each replicate stand. When ANOVA was significant (P < 0.05), we used the least significant difference (LSD) for comparing means between treatments. Then, we used one-way nested ANOVAs with disk samples being nested within each deadwood type in each replicate plot, for comparing water content between the three types of deadwood for each postfire year separately. The nested samples was justified by the fact that we had three disk samples for each type of deadwood in each of four replicate stands representing specific number of years after fire. Two one-way ANOVAs were chosen in order to answer our specific objectives (water content variation for each deadwood type along the chronosequence and difference between each deadwood types for each postfire year). We considered the F value of the test of hypotheses using type III mean of square for site × treatment as an error term to calculate the F value. The least significant difference (LSD) at P = 0.05 was used to compare means of each significant ANOVAs. Data were transformed (log x+1) to satisfy the normality and homogeneity of variance assumptions of ANOVA except for water content data that already met ANOVA assumptions. ANOVAs were performed using the GLM procedure of the Statistical Analysis System (SAS) v.9.2 (SAS Institute Inc., 2000- 2004).

2.4.2 Saproxylic beetle abundance and species richness along the chronosequence

One way ANOVAs were used to compare abundance and species richness of beetles between years along the chronosequence. The least significant difference (LSD) was calculated to compare means for each significant ANOVA. Abundance was transformed to its 4th square root to satisfy the normality and homogeneity of variance assumptions of

21 ANOVA. ANOVAs were performed using the GLM procedure of the Statistical Analysis System (SAS) v.9.2 (SAS Institute Inc., 2000-2004).

2.4.3 Beetle assemblages along the chronosequence

To assess the effects of time since wildfire on beetle assemblages, we performed Permutational Multivariate Analysis of Variance (PERMANOVA (Anderson, 2001)) using distance matrices produced with the Bray-Curtis dissimilarity as default with the adonis procedure in the vegan package (Oksanen, 2011) of R (R-Development-Team, 2009). The adonis procedure (type III ADONIS) accounted for all habitat attributes and examined how these affected beetle assemblages between postfire years along the chronosequence. A significant P-value in PERMANOVA may indicate that differences were observed in beetle assemblages between postfire years and/or in the within-postfire years dispersion, or both (Oksanen, 2011). Habitat attributes data were normalized by column maximum and a Hellinger transformation was applied on beetle species (Legendre and Gallagher, 2001). This procedure reduces the influence of both rare and abundant species on the overall assemblage pattern. The significance of differences in beetle assemblages along the chronosequence was tested by permutations (n = 999). Differences between each combination of postfire years were tested using pairwise PERMANOVA comparisons. Permutation analysis of homogeneity of multivariate group dispersions (PERMDISP;(Anderson, 2006)) was also used to determine if beetle assemblages differed only by their within-treatment dispersion calculated from the average distance of stands to their group centroid on the basis of principal coordinate axes. For this analysis, we used the betadisper procedure of the vegan package (Oksanen, 2011) in R (R-Development-Team, 2009) for this analysis.

2.4.4 Association between saproxylic beetles and habitat attributes along the chronosequence

Multivariate analyses were used to assess which habitat attributes were linked with variation in species assemblage along the chronosequence. A canonical redundancy analysis (RDA) was conducted on Hellinger-transformed data using the RDA procedure of the vegan package (Oksanen et al., 2009) in R (R-Development-Team, 2009). According to

22 (Legendre and Gallagher, 2001), Hellinger-transformation reduces the weight of rare species which are recorded sporadically in some sites but not in others even if they might be present. Habitat attributes used in RDA were the mean basal area of young, medium and old snags and the mean volume of young, medium and old DDW. Centroids were used to position the geometrical center of stands of each postfire year.

To identify species positively associated with a specific postfire year along the chronosequence or with a combination of postfire years, we have used the point-biserial group-equalized phi coefficient (Pearson correlation), as described by De Cáceres and Legendre (2009). We applied the MULTIPATT procedure in the Indicspecies package of R (De Cáceres, 2010). This analysis calculates correlation coefficients which take into account absences outside the target groups (in our case, a postfire year or group of postfire years) as well as presences in stands of that group, thus increasing the power of the associations. Therefore, the analysis is more context dependent than indicator indices such as the IndVal index (Dufre ne and Legendre, 1997) for the determination of species-habitat associations (De Cáceres and Legendre, 2009). We performed this analysis with all species identified.

3 Results

3.1 Overview

Among the eight studied families, 5 510 beetles belonging to 72 species were caught and identified. Captures in postfire years 1 and 2 accounted for 79 % of captures with 3 090 beetles caught during the year of the fire (year 0) and 1 246 for postfire year 1. Only 1 173 (21 %) beetles were sampled in the other postfire years: 749, 199, 103 and 38 respectively 3, 5, 7 and 15 years postfire. Overall, the three most abundant species were Dryocoetes autographus (Ratzeburg), a scolytid with 1 886 specimens captured followed by Arhopalus foveicollis (Haldeman), a cerambycid with 782 specimens captured and Thanasimus undatulus nubilus (Say), a clerid with 711 specimens captured.

23 3.2 Habitat attributes along the chronosequence

Even if there was no significant difference in the total snag basal area (F 5, 18 = 2.23,

P = 0.0963), as well as for the basal area of recent (F 5, 18 = 2.25, P = 0.0939) and medium

(F 5, 18 = 1.95, P = 0.1358) snags, a decreasing trend was observed for total and recent snags between 7 and 15 years after fire (Table 1.1). Basal area of broken snags (old class) increased significantly (F5, 18 = 2.96, P = 0.0403) during the first three years and stabilized for the rest of the chronosequence. The total volume of DDW was significantly lower for the very year of the fire (F5, 18 = 6.44, P = 0.0013) while the volume of medium DDW increased progressively (F5, 18 = 5.45, P = 0.0032) along the first five postfire years of the chronosequence (Table 1.1). There was no significant change in overall and all decay categories of dead saplings along the chronosequence (Table 1.1). Living jack pine saplings began to be significantly present 15 years after fire (F 5, 18 = 4.89, P = 0.0053) while black spruce saplings were nearly absent along the chronosequence (F 5, 18 = 1.05, P = 0.4185) (Table 1.1).

Bark recovering trunks of snags was significantly highest during the first 3 postfire years (Figure 1.2). There was a significant loss of bark (> 50 %) for postfire years 5 and 7 and another significant loss for postfire year 15 in which only 13 % of bark was still covering snag boles (Figure 1.2). The water content of snags decreased significantly and abruptly (F

23, 48 = 8.86, P < 0.0001) one year after fire and continued to decrease slowly along the chronosequence (Figure 1.3A). In DDW, the water content decreased significantly (F 23,

48 = 2.45, P = 0.0045) between 3 and 5 years after fire and then, did not change until at least 15 years post fire (Figure 1.3B). No significant difference in water content was observed in stumps along the 15 years chronosequence (F 23, 48 = 1.70, P = 0.0606) (Figure 1.3C). When we compared the water content between the three different types of deadwood for each postfire year separately, we observed significant differences 1 year after fire (F11, 24 = 22.80,

P = 0.002), 3 years after fire (F11,24 = 23.30, P = 0.002) as well as 5 years after fire (F11,

24 = 6.20, P = 0.035) (Figure 1.4). For these years, water content was consistently higher in DDW compared with water content in snags and stumps according to the LSD test. Even if it was not significant in postfire years 7 and 15, we observed a trend showing greater water content in DDW and stumps compared with snags.

24 3.3 Saproxylic beetle abundance and species richness along the chronosequence

The abundance of saproxylic beetles was highest (F5, 18 = 35.10, P < 0.0001) during the year of the fire (year 0) and dropped significantly the following year (year 1) and again between years 3 and 5 after fire (Figure 1.5). After this second drop, saproxylic beetle abundance remained stable until at least 15 years after fire. Species richness also decreased with time after fire but more progressively (F5, 18 = 7.45, P = 0.0006) (Figure 1.5B).

3.4 Beetle assemblages along the chronosequence

PERMANOVA revealed that time since wildfire had significant effects on beetle assemblages (Table 1.2). Pairwise comparisons of the PERMANOVA indicated differences between postfire years for up to postfire year 5 (Table 1.2). Afterwards, no significant difference was observed. PERMDISP results revealed that time after fire had also a significant effect on the within-treatment dispersion (F5, 18 = 24.89, P = 0.001). This indicates that the differences revealed by PERMANOVA do not result only from differences in species assemblages across treatments but also from within-treatment dispersion. PERMDISP indicated differences between the same years as the pairwise comparisons of the PERMANOVA (Table 1.2).

3.5 Saproxylic beetle communities and habitat attributes along the chronosequence

The RDA ordination was significant (F = 3.49, P = 0.005, 199 permutations) and 62.6 % of the variance was explained by the first two axes (Axis 1: 47.35 %, Axis 2: 15.1 %) (Figure 1.6). Centroids of younger burned stands (0, 1 and 3 years after fire) were on the right side of the first axis of the triplot while those of older burned stands (5, 7 and 15 years after fire) were on the left of the axis. The recent snags as well as the old DDW vectors were related to younger burned sites. The recent snag vector was associated with several xylophagous beetles during the very year of the fire (year 0) while the old DDW vector pointed toward the centroids of stands burned 1 and 3 earlier, along species such as Dryocoetes autographus and Sphaeriestes virescens. Vectors related to the older burned stands (medium snags, old snags, recent DDW and medium DDW) were found on the left of the first axis. The medium snags, the old snags and the recent DDW vectors were linked to the

25 centroid of stands burned 5 years earlier while the medium DDW vector was oriented toward the centroids of stands burned 7 and 15 years earlier.

The analysis seeking species-habitat associations (Pearson’s point-biserial correlation) revealed that 19 species were significantly associated with a specific year along the postfire chronosequence or a combination of postfire years (Table 1.3). Among these, 14 species were associated with a specific year and 5 species were associated with a combination of years along the postfire chronosequence. Most species (17 out of 19) were associated with the first three years after fire while only two elaterid species of the genus Ampedus were associated with years 5 and 15 after fire (Table 1.3). The Cerambycidae family was the most represented among the nine species associated with the very year of the fire with three species, Arhopalus foveicollis (Haldeman), Monochamus s. scutellatus (Say) and Acmaeops p. proteus (Kirby). Two Buprestids and one weevil species, which are also xylophagous, were associated with the very year of the fire. The other three species associated with the very year of the fire were either mycophagous (Clypastraea fusca and Corticaria brevicornis Fall) or predator (Thanasimus u. nubilus). Two xylophagous and two Elaterid species were associated with postfire years 0 and 1 (the buprestid Buprestis maculiventris and the elaterid Ampedus evansi) or 0, 1 and 3 (the Curculionid Dryocoetes autographus and the Elaterid Eanus decoratus). Three species were associated with postfire year 1: the elaterid S. incongruous, the mycophagous Corticaria dentigera and the predator Sphaeriestes virescens. The bark beetle, Polygraphus rufipennis (Kirby), was associated with a combination of postfire years 1 and 3.

4 Discussion

Our results revealed that the onset of saproxylic beetle colonization began early after fire, with large abundance and species richness between 0 and 3 years after fire. These results are similar to Boulanger and Sirois (2007) for black spruce forests of the lichen woodland ecosystem in Northern Quebec. However, we caught twice the number of species than in Boulanger and Sirois (2007), for the same 8 beetle families, during the first two years of the chronosequence (53 species in our study versus 26 species in Boulanger and Sirois (2007)). The large abundance and species richness observed shortly after fire might have resulted

26 from the large amount of recent commercial burned snags. After fire, most trees die almost immediately compared with other disturbances like insect outbreaks or windthrow. Thus, fire-killed trees provide different substrates for saproxylic beetles (Ahnlund and Lindhe, 1992) very soon after fire. In addition, snags still had high water content in the months following fire but their defensive mechanisms are no longer functional (except in lightly burned trees that may survive). Therefore, saproxylic beetles may simply benefit from this high quality resource shortly after disturbance (Muona and Rutanen, 1994; Wikars, 1994; Saint-Germain et al., 2004b-c). Our results on water content suggest that the quality of the snag resource drops rapidly, as soon as the next postfire year. Indeed, water content in snags decreased by about 50 % 1 year after fire, which could explain why beetle abundance also dropped by 50% on the postfire year 1. Nevertheless, species richness remained similar. This suggests that saproxylic beetles take advantage of the high quality resource rapidly after fire and have the ability to find efficiently these new habitats.

Dying and freshly-killed trees are known to release volatile compounds like ethanol and monoterpenes that attract several saproxylic beetles (Erbilgin and Rafa, 2001; Lindgren and Miller, 2002b; Lindgren and Miller, 2002a). Fire-killed trees also release volatile compounds that attract saproxylic beetles but, as for water content, the quantity of volatiles released decreased significantly the year following fire (Suckling et al., 2001; Kelsey and Joseph, 2003). Some species, known as pyrophilic, have developed adaptations to detect cues as heat and smoke (Schütz et al., 1999) and/or infrared (Evans, 1966; Schütz et al., 1999; Schmitz et al., 2000). Melanophila acuminata (DeGeer) was one of two Buprestids associated with the very year of the fire and showed the highest capture. This species was also caught exclusively in burned stands (none in plots with girdle trees) by Boucher et al. (2012). Indeed, this species is known as pyrophilous (fire-dependant species for its life cycle (Klocke et al., 2011)) or strongly associated with burns in Europe (Ahnlund and Lindhe, 1992; Wikars, 1997). It has been proven that this species has adaptive traits to detect infrared (Evans, 1966; Evans and Kuster, 1980; Schmitz et al., 2008; Klocke et al., 2011) which provide the opportunity to arrive earlier in the burned area compared with other species (Evans, 1966). Therefore, M. acuminata could benefit from the high quality resources as well as from the absence of competition which could explain its dependence on fire (Wikars, 2002). Our results also showed that most species associated with the very

27 year of fire were phloeophagous or xylophagous. These first colonizers are important to initiate dead wood decomposition, as soon as the very year of fire, because they rapidly invade woody tissues through their gallery digging activities (Edmonds and Eglitis, 1989; Schowalter et al., 1992; Müller et al., 2002; Barker, 2008). These galleries increase the surface available for establishing fungi and microorganisms that are important in wood decay processes (Schowalter et al., 1992; Progar et al., 2000; Müller et al., 2002; Barker, 2008). The establishment of specific fungal communities (mostly ascomycetes) under bark, which was almost intact, in the very year of fire could explain the rapid colonization of fungivorous species like Clypastraea fusca (Harold) and Corticaria brevicornis Fall. Predator Thanasimus u. nibilus were also very abundant and positively associated with the very year of fire which could be linked with abundant phloeophagous prey under bark (Poland and Borden, 1997; Reeve, 1997).

Despite the absence of difference in species richness between postfire years 0 and 1, beetle abundance significantly differed and the PERMANOVA showed clearly different beetle assemblages. This indicates that rapid changes occur in beetle assemblages after wildfire. After the first colonisation phase, another cohort of species, which was already present on year 0, then increased in abundance but reached lower levels than colonizing species. It included the cerambycid Rhagium inquisitor, which was nearly absent in other stands. Boulanger and Sirois (2007) also caught this species more abundantly 1 year after wildfire and this species was also nearly absent in other aged-stands after fire. This suggests that this species may need lower food quality, at the phloem/cambium interface, than species such as Monochamus s. scutellatus which was, as in Boulanger and Sirois study (2007), present almost exclusively during the very year of fire. Nevertheless, the window of adequacy of food resource also appears to be limited mostly to the year following fire. Several species previously reported to be associated with burned forests were also among the dominant species one year after fire: the predators Sphaeriestes virescens and Sericus incongruus, as well as the fungivorous Corticaria dentigera (Saint-Germain et al., 2004c; Boulanger and Sirois, 2007; Boucher et al., 2012).

Species richness and beetle assemblages differed significantly between postfire years 1 and 3 but beetle abundance was similar. This was the third succession phase by saproxylic

28 beetles. Overall, 19 species among which 14 were singletons or doubletons, were present 1 year after fire but were no longer caught 3 years after fire. According to our results, basal area of broken spruce snags increased significantly during the first 5 years which means that few sections of black spruce snags have fallen and became logs immediately between 1 and 3 years after fire. We put forward the hypothesis that fire has weakened died or moribund trees already present before fire and the increased wind velocity in such newly open stand could accelerate snag fall. This phenomenon could also be partly explained by the damage caused by wood-borer beetles that weaken stems, thus contributing to windthrow (Harmon et al., 1986). This dynamic in deadwood structure (from snags to logs) could be linked with variations in beetle assemblages and explain the third colonisation phase observed on 3 years after fire. New DDW have higher water content than snags, due to their contact with the soil (Boulanger and Sirois 2007), and this should favour fungal growth (Wei et al., 1997; Næsset, 1999; Boulanger and Sirois, 2007). This should provide new micro-habitats for saproxylic beetles and thus modify beetle assemblages. Along the chronosequence, water content was always higher in DDW than in snags and it remained highest during the first 3 postfire years in DDW, making this attribute favorable for saproxylic beetle assemblages.

The wood-borers genus Arhopalus and Monochamus, the predators Thanasimus and Sphaeriestes and the bark beetle Dryocoetes were also found in other research as early burn-associated species in boreal forests of Quebec as well as in Fennoscandia (Muona and Rutanen, 1994; Wikars, 1994; Saint-Germain et al., 2004b-c; Boucher, 2011). Those species clearly exploited burned habitats but they have also been found, in smaller numbers, in other disturbances or in mature unburned forests (Saint-Germain et al., 2004c; Boucher et al., 2012). Therefore, these species are not completely fire-dependant to complete their life cycle but they clearly take advantage of the early years after fire (0 to 3) to build-up their populations. How this is important to maintain these species on other sources of dead or moribund trees across the forest landscape remain to be determined (Wikars, 2002). Other early burn-associated species identified in our study could be more opportunistic than really associated with burned forests because they have been found abundantly in forests disturbed from other causes (Boucher et al., 2012). This is the case for Acmaeops proteus proteus (Kirby) and Polygraphus rufipennis (Kirby), which were

29 identified as burn-associated species by Saint-Germain et al. (2004c) but opportunistic by Boucher et al. (2012).

Between 5 and 15 years after fire, beetle abundance and species richness remained low and beetle assemblages similar, which constitute the fourth and last colonisation phase observed along the chronosequence. The low activity of saproxylic beetles could result from the low decomposition rate of fire-killed black spruce in boreal forest (Boulanger and Sirois, 2006; Boulanger and Sirois, 2007) that might be linked to the low water content we measured in all types of deadwood. Deadwood properties in unburned forest such as decay classes, fungal colonization, bark recovery, water content, position and orientation of the debris vary with time and are important factors that may affect the occurrence of saproxylic beetles in deadwood over time (Kaila et al., 1994; Økland, 1996; Kaila et al., 1997; Boulanger and Sirois, 2007). However, 5 years after fire, black spruce snags had lost a large fraction of their bark and were fully exposed to sunlight which decrease rapidly their water content (Bull, 1983; Harmon et al., 1986). These conditions are detrimental to fungal growth which results in slowing down wood decomposition (Gardiner, 1957; Amman, 1972; Haack et al., 1987). In addition, fire-killed wood become harder (Wikars, 1992), the nutritive value of subcortical tissues is deteriorated (Haack et al., 1987; Hanks, 1999) and cues are no longer emitted by fire or fire-killed trees. Even if stumps are more stable habitats, the deterioration of snags seems to alter the saproxylic beetles activity at this level. Boulanger and Sirois (2006) reported that 50% of burned snags were still standing 16 years after fire in northern Quebec. However, our results rather suggest that snags fall faster in the studied area, basal area of overall snags decreasing by nearly 50 % between postfire years 7 and 15. This could result from the larger trees found in our study area. Large diameter trees are more strongly attacked by xylophagous species such as Monochamus s. scutellatus (Saint-Germain et al., 2004b) and may thus be more vulnerable to windthrow (Harmon et al., 1986).

Only 2 elaterid species of the genus Ampedus were associated with late postfire years (5 to 15 years after fire) along the chronosequence. The increase in abundance of these two species could be linked to the amount of DDW with higher water content with their associated fungal development which attract numerous species (Boulanger and Sirois,

30 2007). Note that Ampedus fusculus (LeConte) is the only associated beetle that was absent at years 0, 1 and 3 after fire inclusively. In Europe, species of the genus Ampedus are generally living in the cavities of rotten wood (Du Chatenet, 2000). Thus, it is possible that they feed on wood decay fungi in their early stages (Du Chatenet, 2000). However, it is known that several Ampedus species that develop in rotten wood are frequently found with Lucanidae larvae thus, they could also be predators of those larvae in their later stages (Du Chatenet, 2000; Lassauce et al., 2012). The exact biology of several Elaterid species is still unknown in Quebec (G. Pelletier, personal communication) and diverse genera such as Ampedus, Limonius and Dalopius need revision as well as phylogenetic study (Jonhson, 2002).

Our study revealed four succession phases by saproxylic beetles along a 15 years postfire chronosequence. These results are more closer from those of Essen et al. (1992) than from those of Boulanger and Sirois (2007). Indeed, Essen et al. (1992) observed four colonisation phases by saproxylic beetles on spruce logs in old boreal forest of Fennoscandia within 20 years after tree-fall while Boulanger and Sirois observed two distinct colonisation phases at snag height along a 29 years postfire chronosequence. Therefore, our results suggest that burned stumps and logs are microhabitats with proprieties closer to those in natural forest compare to burned snags. As time elapses after fire, stumps and logs remain a more stable microhabitat than snags and thus, could be more favorable for long-term saproxylic communities.

Acknowledgments

We thank L.-P. Ménard and J.-F. Bourdon from l’Université Laval, the intern C. Ageron from IUT Saint-Étienne in France and A. Dieni and G. Meunier from l’Université de Sherbrooke for field and laboratory assistance as well as Y. Dubuc of the Canadian Forest Service of Natural Resources Canada (CFS-NRCan) for technical support. We also greatly appreciated the help of G. Pelletier, taxonomist at CFS-NRCan, for taxonomic training and for assistance for validating identifications. We also express our gratitude to J. Boucher from l’Université Laval for experimental design, site selection and statistical assistance and to L. Préfontaine for help with the English revision of this manuscript. Finally, we thank the

31 Laurentian Forestry Centre for laboratory facilities. This study was supported by the Fonds Québécois de Recherche sur la Nature et les Technologies (FQRNT) through the Programme de recherche en partenariat sur l’aménagement et l’environnement forestiers-II, by the iFor consortium (Université Laval), and by the Canadian Forest Service.

32 Tables Table 1. 1 Habitat attributes (average ± standard error) measured in postfire black spruce stands along a 15 years chronosequence.

Habitat attributes Time after fire (year) 0 1 3 5 7 15 Basal area1 of snags (m²/ha) Recent snags 29.48 ± 4.37 23.55 ± 3.91 24.24 ± 5.39 22.11 ± 4.01 30.08 ± 9.70 11.74 ± 7.31 Medium snags 0.47 ± 0.47 0.17 ± 0.11 5.97 ± 3.34 6.71 ± 3.86 6.12 ± 2.60 2.70 ± 1.57 Old snags a 0.71 ± 0.33 a 0.80 ± 0.38 ab 4.64 ± 1.19 c 4.12 ± 1.88 bc 3.66 ± 1.03 c 4.34 ± 2.13 c Total snags 30.65 ± 4.22 24.53 ± 3.44 34.95 ± 2.65 32.93 ± 8.42 39.85 ± 7.63 18.78 ± 7.32 Volume of DDW (m³/ha) Recent DDW 0.00 ± 0.00 3.56 ± 3.56 0.34 ± 0.34 9.56 ± 8.37 17.92 ± 17.46 0.00 ± 0.00 Medium DDW a 4.30 ± 1.23 a 21.07 ± 7.45 b 29.27 ± 17.43 b 58.18 ± 12.27 bc 50.85 ± 25.62 bc 94.03 ± 30.78 c Old DDW 0.79 ± 0.58 13.02 ± 8.90 39.51 ± 32.18 10.28 ± 5.88 3.88 ± 2.97 7.59 ± 7.21 Total DDW a 5.09 ± 1.25 a 37.65 ± 18.71 b 69.13 ± 49.41 b 78.02 ± 19.09 b 72.66 ± 19.24 b 101.63 ± 36.60 b Basal area1 of stumps (m²/ha) Recent stumps 0.03 ± 0.03 0.00 ± 0.00 0.00 ± 0.00 0.00 ± 0.00 0.00 ± 0.00 0.00 ± 0.00 Medium stumps a 1.13 ± 0.82 a 6.67 ± 2.61 b 4.05 ± 1.56 ab 9.86 ± 3.50 b 8.14 ± 3.46 b 9.01 ± 1.83 b Old stumps 3.34 ± 0.94 6.60 ± 2.15 5.30 ± 3.94 6.16 ± 2.43 1.80 ± 0.77 5.70 ± 3.40 Total stumps 4.49 ± 1.73 13.37 ± 4.30 9.35 ± 5.18 16.02 ± 5.46 9.94 ± 3.85 14.72 ± 4.67 Basal area of dead saplings2 (m²/ha) Recent dead saplings 4.17 ± 1.00 2.81 ± 1.51 2.88 ± 1.76 3.96 ± 1.37 1.84 ± 0.70 1.26 ± 0.51 Medium dead saplings 0.00 ± 0.00 0.33 ± 0.33 0.00 ± 0.00 0.08 ± 0.08 0.82 ± 0.56 2.56 ± 1.81 Old dead saplings 0.86 ± 0.42 0.35 ± 0.33 0.67 ± 0.39 0.16 ± 0.16 0.00 ± 0.00 0.47 ± 0.30 Total dead saplings 5.04 ± 1.36 3.50 ± 1.33 3.54 ± 1.98 4.19 ± 1.44 2.66 ± 0.31 4.29 ± 2.11 Basal area of live saplings2 (m²/ha) Living P. banksiana saplingsa 0.00 ± 0.00 a 0.00 ± 0.00 a 0.00 ± 0.00 a 0.00 ± 0.00 a 0.02 ± 0.02 a 2.55 ± 1.62 b Living P. mariana saplings 0.00 ± 0.00 0.00 ± 0.00 0.12 ± 0.12 0.00 ± 0.00 0.00 ± 0.00 0.09 ± 0.06 aLetters represent statistical differences at α < 0.05. Anova’s were done on log-transformed values 1Estimated from diameter measured at the DSH (30 cm high) 2Estimated from diameter measured at the DBH (1,3 m high)

33 Table 1. 2 PERMANOVA and PERMDISP results showing the effect of time since wildfire on beetle assemblages. Pairwise comparisons between years, using t-tests, are shown for PERMANOVA and PERMDISP.

PERMANOVA d.f. MS F R2 P Time since wildfire 5 0.593 5.88 0.62 0.0005 Residuals 18 0.101 0.38 Total 23 1.00 Pairwise comparaisons between years after fire T R2 P 15 vs. 7 1.25 0.21 0.111 7 vs. 5 1.26 0.21 0.121 5 vs. 3 1.81 0.35 0.026 3 vs. 1 1.79 0.35 0.022 1 vs. 0 1.82 0.36 0.032

PERMDISP d.f. MS F P Time since wildfire 5 0.032 24.89 0.001 Residual 18 0.001 PERMDISP between years after fire P 15 vs. 7 0.303 7 vs. 5 0.423 5 vs. 3 0.002 3 vs. 1 0.042 1 vs. 0 0.004

34 Table 1. 3 Beetle species significantly associated (p < 0.05) with a postfire year or a combination of postfire years from point-biserial group-equalized phi coefficient analysis (Pearson correlation). Number of individuals in bold refers to the total number of specimens captured for each year with which a species was positively associated. Code species refers to the abbreviations of the full name species and those codes were used in the RDA instead of the full name species. Families Code species Species Time after fire (year) Phi coefficient p 0 1 3 5 7 15 Buprestidae Bup. nut. Buprestis nuttalli Kirby 4 1 0 0 0 0 0.711 0.040 Buprestidae Mel. acu. Melanophila acuminata (DeGeer) 163 1 0 0 0 0 0.781 0.001 Cerambycidae Acm. p.pro. Acmaeops p. proteus (Kirby) 110 7 1 0 0 0 0.949 0.001 Cerambycidae Arh. fov. Arhopalus foveicollis (Haldeman) 651 63 66 2 0 0 0.860 0.001 Cerambycidae Mon. s.scu. Monochamus s. scutellatus (Say) 150 1 0 0 0 0 0.945 0.001 Cleridae Tha. u.nub. Thanasimus u. nubilus (Say) 659 45 5 2 0 0 0.938 0.001 Corylophidae Cly. fus. Clypastae fusca (Harold) 63 2 3 0 0 0 0.948 0.001 Curculionidae Hyl. con. Hylobius congener D. T., Schen. & Mars. 45 6 1 1 0 1 0.829 0.001 Lathridiidae Cor. bre. Corticaria brevicornis Fall 140 10 13 15 10 10 0.870 0.001 Buprestidae Bup. mac. Buprestis maculiventris Say 10 6 0 1 0 0 0.667 0.007 Elateridae Amp. eva. Ampedus evansi W.J. Brown 8 8 0 1 0 0 0.699 0.011 Curculionidae Dry. aut. Dryocoetes autographus (Ratzeburg) 917 469 475 23 1 1 0.677 0.023 Elateridae Ean. dec. Eanus decoratus (Mannerheim) 9 10 19 3 2 2 0.733 0.003 Elateridae Ser. inc. Sericus incongruus (LeConte) 3 29 4 4 3 6 0.863 0.001 Lathridiidae Cor. den. Corticaria dentigera LeConte 50 151 50 40 4 7 0.732 0.001 Salpingidae Sph. vir. Sphaeriestes virescens (LeConte) 12 220 21 2 1 3 0.881 0.001 Curculionidae Pol. ruf. Polygraphus rufipennis (Kirby) 1 17 31 2 0 0 0.557 0.044 Elateridae Amp. fus. Ampedus fusculus (LeConte) 0 0 0 8 2 0 0.635 0.028 Elateridae Amp. pul. Ampedus pullus Germar 8 9 0 3 7 28 0.654 0.048

35 Figure

Figure 1. 1 Study area showing selected burns with their respective time (year) since wildfire. X represent burns with a surface area less than 275 hectares.

36 125

a (%) (%) 100 a

snagssnags on on on on 75

b bark

bark b of of of of 50

25 Recovery Recovery c

0 0 1 3 5 7 15 Time after fire Figure 1. 2 ANOVA’s results comparing % bark covering on trunks of snags along the chronosequence (average ± standard error). Letters represent statistical differences at α < 0.05.

Figure 1. 3 Average (± se) water content (%) along a 15 years postfire chronosequence in black spruce boreal forest for A) snags, B) down deadwood (DDW) and C) stumps. Different letters indicate significant differences at α < 0.05.

p = 0.070 p = 0.002 p = 0.002 p = 0.035 p = 0.079 p = 0.135 40 a

a DDW Stumps 30 Snags b a 20 b

b b content (%)

content (%) b b Water Water 10

0 0 1 3 5 7 15 Time after fire (years) Figure 1. 4 Comparison of the water content (average ± se in %) between black spruce down deadwood (DDW), stumps and snags for each year, separately, along a 15 years postfire chronosequence in the boreal forest. Letters indicate significant differences at α < 0.05.

39 A

1050 a 900

750

600

Abundance 450 b

300 b 150 c c c 0 0 1 3 5 7 15

35 a 30 ab 25

20 c bc species

15 c c

Richness 10

0 0 1 3 5 7 15 Time after fire (years)

Figure 1. 5 Average (± se) abundance (A) and species richness (B) of beetles along a 15 years postfire chronosequence in black spruce boreal forest. Letters indicate significant differences at α < 0.05. Anova’s were done on fourth root-transformed values for the abundance.

Cor. Den. +

Dry. Aut. 0.4 0.4 +

Sph. vir. + 5 Ser. inc. + Pol. ruf.

+ 3 0.2 0.2 M. sna. O. sna. R. DDW O. DDW Amp. fus. 1 +

Ean. dec. Amp. eva.

+ +

0 0 0.0 0.0 7 Bup. mac. Bup. nut. +

+ Hyl. con. RDA2 (15.18%) RDA2 RDA2 (15.18%) RDA2 + M. DDW R. sna. 15 Cly. fus. +

Cor. bre. Acm. p.pro. 0.2

0.2 + - - + Amp. pul. Mel. acu. Arh. fov. + + + Mon. s.scu. +

0 0.4

0.4 Tha. u.nub. - - +

-0.5 0.0 0.5 1.0 RDA1 (47.35%) Figure 1. 6 RDA ordination of beetles caught in postfire black spruce stands of the boreal forest. Centroids of stands of each postfire time since disturbance are illustrated by the number of years in bold. Species are marked with a cross while habitat attributes are illustrated with arrows in bold. All species were used in the analysis but only those which were significantly associated with specific postfire years (or a combination of several ones) in the species-habitat associations analysis (Pearson’s point-biserial correlation; see table 3) were illustrated on the ordination. Also, only significant habitat attributes are shown: basal area of black spruce snags and volume of black spruce DDW of each decay class. The code species, presented in table 3, were used instead of the full name species. Habitat attribute codes were: R. sna. = recent snag, M. sna. = medium snag, O. sna. = old snag, R. DDW = recent down deadwood, M. DDW = medium down deadwood and O. DDW = old down deadwood.

41 References cited

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47 Swift, M. J., 1977b. The roles of fungi and animals in the immobilisation and release of nutrient elements from decomposing branch-wood. Ecological Bulletins 25, 193- 202. Tinker, D. B., Knight, D. H., 2000. Coarse woody debris following fire and logging in wyoming lodgepole pine forests. Ecosystems 3, 472-483. Van Wagner, C. E., 1982. Practical aspects of the line intersect method. In: Petawawa National Forestry Institute, C. F. S. (Ed.). Petawawa National Forestry Institute, Canadian Forestry Service, Environment Canada, Chalk River, Ontario, Canada, p. 11. Warren, W. G., Olsen, P. F., 1964. A line intersect technique for assessing logging waste. Forest Science 10, 267-276. Wei, X., Kimmins, J. P., Peel, K., Steen, O., 1997. Mass and nutrients in woody debris in harvested and wildfire-killed lodgepole pine forests in the central interior of British Columbia. Canadian Journal of Forest Research 27, 148-155. Wikars, L.-O., 1992. Forest fires and insects. Entomologist Tidskrift 113, 1-11. Wikars, L.-O., 1994. Effects of fire and ecology of fire-adapted insects. Introductory research essay no.12. In, Dept. of Zoology. University of Uppsala, Sweden. Wikars, L.-O., 1997. Pyrophilous insects in orsa Finnark, central Sweden: biology, distribution and conservation. Entomologist Tidskrift 118, 158-169. Wikars, L.-O., 2002. Dependence on fire in wood-living insects: an experiment with burned and unburned spruce and birch logs. Journal of Insect Conservation 6, 1-12. Wilson, C. V., 1971. Le climat du Québec parti 1: atlas climatique. Service météorologique du Canada, Études climatologiques no 11, 44 figures.

48 Appendix A1 Species list and number of beetles captured for each postfire year in black spruce boreal forest.

Time after fire (year) Families Species 0 1 3 5 7 15 Total Buprestidae Anthaxia inornata (Randall) 1 0 0 0 0 0 1 Buprestis maculiventris Say 10 6 0 1 0 0 17 Buprestis nuttalli Kirby 4 1 0 0 0 0 5 Dicerca lugubris LeConte 0 1 0 0 0 0 1 Melanophila acuminata (DeGeer) 163 1 0 0 0 0 164 Cerambycidae Acanthocinus pusilus Kirby 1 0 0 0 0 0 1 Acmaeops pratensis (Laicharting) 2 5 4 4 7 0 22 Acmaeops proteus proteus (Kirby) 110 7 1 0 0 0 118 Arhopalus foveicollis (Haldeman) 651 63 66 2 0 0 782 Asemun striatum (Linné) 3 3 0 0 0 0 6 Clytus ruricola (Olivier) 0 0 0 0 2 0 2 Monochamus scutellatus scutellatus (Say) 150 1 0 0 0 0 152 Pachyta lamed liturata Kirby 1 0 1 0 0 0 2 Pogonocherus mixtus Haldeman 0 0 0 1 0 0 1 Pygoleptura nigrella nigrella (Say) 0 0 0 1 0 2 3 Rhagium inquisitor (Linné) 4 115 1 1 0 0 121 Stictoleptura canadensis canadensis (Olivier) 5 2 0 2 8 2 19 Tetropium cinnamopterum Kirby 1 1 0 0 0 0 2 Trachysida aspera brevifrons (H. Howden) 0 1 0 0 0 0 1 Xylotrechus undulatus (Say) 2 0 0 0 0 0 2 Cleridae Thanasimus dubius (Fabricius) 0 0 0 1 0 0 1 Thanasimus undatulus nubilus (Say) 659 45 5 2 0 0 711 Corylophidae Clypastraea fusca (Harold) 66 2 3 0 0 0 71 Curculionidae Cossonus americanus Buchanan 0 0 0 0 0 1 1 Crypturgus borealis Swaine 0 0 0 2 0 0 2 Dryocoetes affaber (Mannerheim) 1 2 4 1 0 0 8 Dryocoetes autographus (Ratzeburg) 917 469 475 23 1 1 1886 Dryocoetes betulae Hopkins 2 0 3 1 0 0 6 Hylastes porculus Erichson 1 0 0 0 0 0 1 Hylobius congener D. T., Schen. & Mars. 45 8 1 1 0 1 56 Hylurgops rugipennis (Mannerheim) 3 1 0 0 0 0 4 Ips latidens (LeConte) 1 1 0 0 0 0 2 Pithyophtorus sp. Erichhoff 0 1 1 1 0 0 3 Polygraphus rufipennis (Kirby) 1 17 31 2 0 0 51 Rhyncolus brunneus Mannerheim 0 0 0 1 0 0 1 Rhycolus macrops 0 0 0 0 0 2 2 Trypodendron lineatum (Olivier) 0 2 1 0 0 0 3 Xyloborus sayi (Hopkins) 1 0 0 0 0 1 2 Elateridae Agriotes limosus (LeConte) 10 19 2 18 9 6 64 Ampedus apicatus (Say) 1 1 0 0 1 0 3 Ampedus evansi W. J. Brown 8 8 0 1 0 0 17 Ampedus fusculus (LeConte) 0 0 0 8 2 0 10 Ampedus laurentinus W. J. Brown 6 1 5 4 1 2 19 Ampedus luctuosus (LeConte) 6 2 0 2 7 7 24 Ampedus mixtus (Herbst) 1 2 2 1 2 6 14 Ampedus nigrinus (Herbst) 0 4 0 3 0 6 13 Ampedus pedalis Germar 0 1 1 0 1 0 3 Ampedus pullus Germar 8 9 0 3 7 28 55 Ampedus quebecensis W. J. Brown 3 5 3 4 1 3 19 Ampedus sp. Dejean 1 1 0 1 0 0 3 Ctenicera kendalli Kirby 0 1 0 0 0 0 1 Dalopius cognatus W. J. Brown 0 0 0 0 0 1 1 Dalopius sp. Eschscholtz 0 0 0 0 0 1 1 Denticollis denticornis (Kirby) 0 0 0 0 0 2 2 Eanus decoratus (Mannerheim) 9 10 19 3 2 2 45 Liotrichus spinosus (LeConte) 3 0 3 0 0 6 Neohypdonus tumescens (LeConte) 1 0 0 1 0 0 2 Pseudanostirus triundulatus (Randall) 0 0 7 22 22 17 68 Selatosomus appropinquans (Randall) 1 0 0 0 0 0 1 Sericus incongruous (LeConte) 3 29 4 4 3 6 49 Setasomus nitidula (LeConte) 0 0 0 0 1 1 2 Lathridiidae Cartodere constricta (Gyllenhal) 19 15 22 19 5 5 85 Corticaria brevicornis Fall 140 10 13 15 10 10 198

(Appendix A1 Continued) Time after fire (year) Families Species 0 1 3 5 7 15 Total Lathridiidae Corticaria dentigera LeConte 50 151 50 40 4 7 302 Cortinicara gibbosa (Herbst) 0 0 0 1 1 0 Lathridius sp. Herbst 0 2 0 0 2 0 4 Melanophthalma pumila (LeConte) 0 0 0 0 1 0 1 Melanophthalma sp Motschulsky 0 0 0 0 2 0 2 Lathridiidae sp. 2 0 0 0 0 0 2 Salpingidae Sphaeriestes virescens (LeConte) 12 220 21 2 1 3 259

50 Chapitre 2

Effects of salvage logging on saproxylic beetle succession along a 12 years postfire chronosequence in the commercial boreal forest

Olivier Jeffrey1, Christian Hébert2, Jacques Ibarzabal3, Richard Berthiaume1 and Éric

Bauce1

1 Université Laval, Faculté de foresterie, de géographie et de géomatique, Pavillon Abitibi-

Price, Québec, Québec, G1K 7P4, Canada

2 Natural Resources Canada, Canadian Forest Service, Laurentian Forestry Centre, 1055 du

P.E.P.S., P.O. Box 10380, Stn. Sainte-Foy, Québec (Québec), G1V 4C7, Canada

3 Université du Québec à Chicoutimi, 555 boulevard de l’Université, Chicoutimi, Québec,

Canada G7H 2B1

51 Résumé

Ce chapitre vise 1) à déterminer si les souches résiduelles à la coupe de récupération offrent un habitat équivalent et durable pour les coléoptères saproxyliques (abondance, richesse en espèces et assemblages d’espèces), 2) à explorer les effets de la coupe de récupération sur la dynamique temporelle de ces coléoptères et 3) à identifier des espèces de coléoptères qui pourraient être associées avec les peuplements résiduels à la coupe de récupération, pour plusieurs années consécutives après feu. Les coléoptères ont été inventoriés durant l’été 2010 dans les mêmes brûlis de 1995, 2003, 2005 et 2007 décrits au chapitre 1. Près de 1 800 coléoptères répartis dans les familles étudiées ont été capturés. La coupe de récupération a affecté l’abondance et la richesse en espèces trois ans après feu seulement alors que les assemblages d’espèces de coléoptères saproxyliques ont été altérés par la récupération pour une période d’au moins sept ans.

Mots-clés : Biodiversité, Succession, Coléoptères saproxyliques, Feu, Coupe de récupération, Souche, Bois mort, Piège à impact troncaux.

52 Abstract

This chapter aims to 1) determine if stumps left in postfire salvaged stands could maintain burned-associated beetles (abundance, species richness and species assemblages), 2) explore the effects of postfire logging on the beetle temporal dynamics and 3) determine if some beetle species could be associated with residual stands, for several consecutive years after fire. Beetles were sampled during summer 2010 in the same burns of 1995, 2003, 2005 and 2007 described in chapter 1. Nearly 1 800 beetles among the studied families were captured. Salvage logging affected the abundance and species richness only for postfire 3 while species assemblages were modified for a period of seven years.

Keywords: Biodiversity; Succession; Saproxylic beetles: Wildfire; Salvage logging; Stump; Deadwood; Trunk widow trap

53 1 Introduction

The concept of ecosystem management, which aims at emulating patterns of variability generated by natural disturbances (Hunter, 1990; Schmiegelow et al., 2006; Gauthier et al., 2008) has been adopted by Canadian forest managers in order to mitigate the differences between natural and managed forests. Thus, salvage logging of naturally disturbed forests raises a paradox for forest management (Schmiegelow et al., 2006). Deadwood, which is largely produced through natural disturbances, has been identified as a key habitat component of forests, as it supports many organisms (Gauthier et al., 2008). Its maintenance is important to ensure conservation of biological diversity and ecosystem resilience (Bull, 1983; Siitonen, 2001; Saint-Germain et al., 2004b; Gauthier et al., 2008; Drapeau et al., 2010). Considering its mid and long-term effects on biodiversity, salvage logging must be carried out on the basis of strong scientific based policies.

By their forest policies, several governments have tried to cope with this issue, in which the recovery of burned forests is mainly done through special management plans that regulate the intensity of salvage logging (Saint-Germain and Greene, 2009). In Quebec, the economic losses resulting from wildfire, combined with the recent reductions in the allocated timber volumes to industries have increased the pressure for salvaging naturally disturbed forests (Purdon et al., 2002; Saint-Germain and Greene, 2009). For example, in 2005 in Quebec public forests, 6.2 millions m3 of burned trees were salvaged which represented about 20 % of the annual wood volume harvested (Parent, 2008; MRNF, 2012). This increase in salvage logging is favored by a better accessibility to burned forests due to the extensive road network (Nappi et al., 2004), as well as to technological improvements which now permit a more efficient treatment of burned trees in lumber mills (Saint- Germain and Greene, 2009). However, salvage logging constitutes a second perturbation after fire and its negative effects on wildlife and for natural recovery of the ecosystem raise conservation issues (Nappi et al., 2004, 2011; Lindenmayer and Ough, 2006). It was raised that salvage logging simplifies stand’s composition and structure as well as the forest mosaic at the landscape scale (Morissette et al., 2002; Purdon et al., 2004; Nappi et al., 2011). This modifies heterogeneity of the fauna and flora assemblages as well as their

54 ecological functions (Lindenmayer and Ough, 2006; Lindenmayer et al., 2008). A reduction in the amount of burned trees across the landscape directly results in a reduction in the availability of postfire habitats for many species. Indeed, snags are colonized by a large number of insects that are preys to many insectivorous birds (Morissette et al., 2002; Purdon et al., 2002). Moreover, postfire logging could have long term effects while affecting late-succession species, especially those that become more abundant about a decade after fire occurrence, when snags fall to the ground (Boulanger and Sirois, 2007). Such practice could thus even exceed the resilience of the ecosystem (e.g. regeneration accident that may lead to the establishment of another type of forest) and of the species that have evolved with fire and which are well-adapted to this disturbance regime (Le Goff et al., 2008; Boucher, 2011; Buma and Wessman, 2011). The removal of dead or moribund trees (e.g. snags) of commercial value over large areas thus reduces one of the key resources for the maintenance of biodiversity and of ecosystem resilience (Rondeux and Sanchez, 2010; Lassauce et al., 2011).

The effects of salvage logging on mid-term succession of saproxylic beetles in postfire stands have never been studied in the boreal forest. Indeed, studies addressing the effects of postfire logging on biodiversity have mostly focused on short-term impacts (Nappi et al., 2004; Lindenmayer et al., 2008; Cahall and Hayes, 2009; Boucher, 2011). Besides, studies assessing the impact of salvage logging, by sampling in residual stands, have mainly measured the effects of varying distances to salvaged stands and/or of different percentages of salvaged areas around the residual stands (Boucher, 2011; Saint-Germain et al., 2012). While studies usually evaluate the impact of snag’s removal from burned stands, we rather investigated the ecological value of stumps that are left on site after salvage logging and which constitute a dead wood resource found in both salvaged and unsalvaged stands. Moreover, stumps persist longer than snags, which break and fall to the ground. Their ecological value as microhabitat for maintaining insect diversity over time after fire and salvage logging has never been assessed.

Our work focussed on eight beetle families (Buprestidae, Cerambycidae, Cleridae, Corylophidae, Curculionidae, Elateridae, Lathridiidae and Salpingidae) known to contain saproxylic (dependant on dead wood (Speight, 1989)) species (Grove, 2002; Saint-Germain

55 et al., 2004b; Boucher et al., 2012) and species that are closely associated with recently burned black spruce (Picea mariana (Mill.)) forests (Wikars, 1992, 1997; Siitonen, 2001; Saint-Germain et al., 2004b).

In this paper, we assessed the mid-term effects of postfire salvage logging on saproxylic beetle legacies, by sampling in both residual and salvaged black spruce stands along a chronosequence covering forests that burned 3, 5, 7 and 15 years earlier. We measured the activity of saproxylic beetles at the stump level using trunk window traps (Kaila, 1993). The aim of this study was to determine if residual stumps left after salvage logging could maintain saproxylic beetle legacies associated with burned black spruce stumps. More precisely, our objectives were 1) to determine if stumps left in postfire salvaged stands could maintain burn-associated beetles (abundance, species richness and assemblages), 2) to explore the effects of postfire logging on the beetle temporal dynamics and 3) to determine if some beetle species could be associated with residual stands, for several consecutive years after fire. These beetles would be the most vulnerable to salvage logging and their presence across the landscape could be affected.

2 Methods

2.1 Study area

The study area was mainly located in the western spruce-moss bioclimatic subdomain and partly in the northern balsam fir white birch bioclimatic of North-Western Quebec, Canada (48°49’-50°43’ N and 70°10’-74°51’W). The climate is characterized by an average annual air temperature varying between 0 and -2.5°C and annual precipitations ranging from 800 to 1 000 mm (Wilson, 1971; Grondin, 1996). This continental climate was characterized by fewer precipitation than in eastern Quebec and thus provided a shorter fire cycle (90-190 years) (MRN, 2000; Boucher et al., 2003). Black spruce stands dominated the landscape and jack pine (Pinus banksiana Lamb.), balsam fir (Abies balsamea (L.) Mill.), trembling aspen (Populus tremuloides Michx.), paper birch (Betula papyrifera Marsch.) and tamarack (Larix laricina (Du Roi) K. Koch) were observed as companion species.

56 2.2 Stand description

We selected thirty-two stands among six burns corresponding to four different years along a chronosequence that covered 12 years after wildfire: two which occurred in 2007, one in 2005, one in 2003 and two in 1995 (Figure 1). Four replicate sites were established in each combination of treatments (salvaged vs residual) and age after fire (3, 5, 7 and 15 years). Within each burn, we systematically paired residual and salvaged sites and each pairs were distanced by at least 5 km to ensure data independence. Each sites of a same pair (i.e., residual vs salvaged) were distanced by about 1.5 km. As the number of burns available for each of the four postfire years varied along the chronosequence, the number of replicates for each treatment in each burn varied from one (1995 burn) to four (2005 burn). This was mainly due to the poor accessibility and availability to the burns. Selected stands varied between 60 and 100 years old and were burned at moderate or high severity (i.e., all trees were dead at time of site selection). All wildfires occurred in the late-spring/early-summer of the coinciding year and salvage logging was carried out during the same summer/fall of fire occurrence or the following winter. We loaded ecoforest and burn severity maps, provided by the Ministère des ressources naturelles (MRN), in ArcMap 9.2 (ESRI, 2009) to facilitate stand selection.

Stand attributes were described within a 400 m² circular plot. Tree species and diameter at stump height (DSH, i.e. 15 cm above root or ground level (MRNF, 2005)) were recorded for every snag larger than 9 cm diameter at breast height (DBH). Tree species and DBH were recorded for every dead and live sapling (> 1.30 m in height and ≤ 9 cm in DBH) within a circular 40 m² plot centered on the 400 m² plot. Four 11.28 m transects were drawn from the plot center toward each cardinal points to estimate the volume (m³/ha) of down dead wood (DDW). Each DDW larger than 5 cm intercepted across the transects was identified when possible and its diameter was measured perpendicularly to the log at the intersection point (Warren and Olsen, 1964). The DDW volume was estimated according to the method of Van Wagner (1982).

In order to compare the water contents of stumps between residual and salvaged stands, we cut three snags in each residual stands and collected a 5 cm thick disk at the stump level (45 cm over root or ground level) as well as a 5 cm thick disk at 45 cm high of three

57 residual stumps in each salvaged stands. Disks were weighted (wet weight) and then oven dried at 65°C until weight stabilized which corresponded to the dry weight (requires a minimum of 36 hours). Water content was expressed as a percentage of wet weight and was calculated with the following formula (Akbulut and Linit, 1999):

wet.weight(g)  dry.weight(g) Water.content  x100 wet.weight

2.2 Beetle sampling

In each plot, three trunk-window traps (TWT), known for their efficiency to sample saproxylic beetles (Kaila, 1993; Boulanger and Sirois, 2007), were installed at stump height (window just above the highest root) on three different snags (in residual stands) or stumps (in salvaged stands). All traps were installed between May 20th and 30th of 2010 and were in operation until August 26th of 2010. Traps were distanced by a minimum of five meters and disposed in order to form an equilateral triangle. These TWT consisted of a 15 × 40 cm Plexiglas® panel installed vertically and perpendicularly to the stump in order to intercept beetles. A 15 cm diameter funnel was placed below the panel to lead beetles in a 500 ml collecting vial filled with 225 ml of a 40% ethanol and 5% household vinegar solution in order to kill and preserve insects. Samples were collected bi-weekly (six collections) and taken back to the Laurentian Forestry Centre (LFC, Quebec, Canada) where they were kept at 4 °C until sorting and identification.

We focused on beetles of eight families closely associated with burned forests (Boucher et al., 2012). Almost all beetles were identified at the species level and vouchers specimens have been deposited in the Forest Insect Ecology and Diversity (EcoDIF) lab collection at LFC.

58 2.4 Statistical analysis

2.4.1 Habitat attributes

Since it was obvious that salvage logging would have an impact on most habitat attributes, we were mainly interested in determining the duration of this impact. Thus, we used paired Student t-tests to compare various habitat attributes between residual and salvaged stands for each postfire year, separately, along the postfire chronosequence. Habitat attributes tested were the basal area of either black spruce snags or dead saplings, the basal area of living saplings of either black spruce or jack pine, the volume of black spruce DDW and the average water content of stumps. Paired Student t-tests were performed using the T- TEST procedures of the Statistical Analysis System (SAS) v.9.2 (SAS Institute Inc., 2000- 2004).

2.4.2 Beetle abundance and species richness

In chapter 1, we showed that there was not much effect of time on the abundance and richness of saproxylic beetles between 3 and 15 years after fire. In fact, only abundance in postfire year 3 was larger than for later years. As we were mainly interested to determine if stumps left in postfire salvaged stands could maintain burn-associated beetles and to explore the effects of postfire logging on the beetle temporal dynamics, we compared beetle abundance and richness for each postfire year separately using a paired Student t-test (T- TEST procedure of the Statistical Analysis System (SAS) v.9.2 (SAS Institute Inc., 2000- 2004).

2.4.3 Beetle assemblages in residual versus salvaged stands over time

Even if abundance and richness were similar between postfire years 3 and 15 (see chapter 1), beetle communities differed. To assess the effects of salvage logging in combination with time after fire as well as their interaction on beetle assemblages, we performed Permutational Multivariate Analysis of Variance (PERMANOVA) (Anderson, 2001). Distance matrices produced with the Bray-Curtis dissimilarity index were used with the adonis procedure in the vegan package (Oksanen, 2011) of R (R-Development-Team, 2009). The adonis procedure accounted for all habitat attributes and examined how these

59 affected beetle assemblages between residual and salvaged stands along the chronosequence. A significant P-value in PERMANOVA indicate that differences were observed in beetle assemblage between residual and salvaged stands along the chronosequence (Oksanen, 2011). Habitat attributes data were normalized by column maximum and a Hellinger transformation was applied on beetle species (Legendre and Gallagher, 2001). This procedure reduces the influence of both rare and abundant species on the overall assemblage pattern. The significance was tested by permutation (n = 999). PERMANOVA was then performed for each postfire year separately to determine if salvage logging had significant effects on beetle assemblages. Permutation analysis of homogeneity of multivariate group dispersions (PERMDISP) (Anderson, 2006) was also used to assess if beetle assemblages differed only by their within-treatment dispersion, calculated from the average distance of stand to their group centroid on the basis of principal coordinate axes. For this analysis, we used the betadisper procedure of the vegan package (Oksanen, 2011) in R (R-Development-Team, 2009).

2.4.4 Residual versus salvage stands associated beetles

To identify species positively associated with residual and/or salvaged stands for a specific postfire year or a combination of postfire years, we have used the point-biserial group- equalized phi coefficient (Pearson correlation), as described by De Cáceres and Legendre (2009). We applied the MULTIPATT procedure in the Indicspecies package of R (De Cáceres, 2010). This analysis calculates correlation coefficients which take into account absences outside the target groups as well as presences in stands of that group, thus increasing the power of the associations. Therefore, the analysis is more context dependent than indicator indices such as the IndVal index (Dufrêne and Legendre, 1997) for the determination of species-habitat associations (De Cáceres and Legendre, 2009). We performed this analysis with all species identified which allowed us to determine which species are associated with residual stands only and thus, more at risk to be affected by salvage logging, instead of those which are associated with both residual and salvage stands.

60 3 Results

3.1 Overview

Among the eight studied families, 1 812 beetles belonging to 67 species were caught and identified. A total of 1 174 beetles distributed amongst 54 species were caught in residual stands while 638 specimens representing 49 species were caught in salvaged stands. Beetles caught in postfire year 3 accounted for 60 % of all captures. Overall, the three most abundant species were the scolytid Dryocoetes autographus (Ratzeburg) with 686 specimens, the lathridid Corticaria dentigera (LeConte) with 179 specimens, and the elaterid Pseudanostirus triundulatus (Randall) with 121 specimens. Arhopalus foveicollis (Haldeman) was the most abundant cerambycid, with 87 specimens captured.

3.2 Effects of salvage logging on habitat attributes along the chronosequence

As expected, basal area in black spruce snags was significantly larger in residual stands all along the chronosequence (Figure 2.2a). The same pattern was observed for the basal area of black spruce dead saplings except for postfire year 3 which was marginally non significant (Figure 2.2c). There was no significant effect from salvage logging on the volume in black spruce DDW (Figure 2.2b). There was no significant effect from salvage logging on the basal area of living saplings of either black spruce or jack pine. The water content remained similar in stumps whether they were collected in residual or salvaged stands all along the chronosequence.

3.3 Effects of salvage logging on beetle abundance and species richness along the chronosequence

The paired student t-test revealed significant effects of salvage logging on beetle abundance and richness for postfire 3 (abundance: t = 3.01, df = 6, P = 0.0184; species richness: t = 2.50, df = 6, P = 0.0465 (Figure 2.3)). Two species were particularly affected by salvage logging for postfire 3: Dryocoetes autographus (Ratzeburg) with 475 specimens caught in residual versus 181 in salvaged stands, and Arhopalus foveicollis (Haldeman) with 66 specimens caught in residual versus 0 in salvaged stands (Table 2.3). For the other postfire

61 years, no significant effect of salvaged logging was observed for both abundance and species richness, but it was marginally significant for richness on postfire year 5 (Figure 2.3).

3.4 Effects of salvage logging on beetle assemblages along the chronosequence

PERMANOVA revealed that salvage logging, in interaction with time after fire, had significant effects on beetle assemblages (Table 2.1). PERMANOVAs done for each postfire year separately showed significant effects of salvage logging on beetle assemblages for postfire years 3 and 7 (Table 2.2). PERMDISP indicated a significant effect of salvage logging on the within-treatment dispersion only for postfire year 5 (average distance to residual stands centroid = 0.309 versus average distance to salvaged stands centroid = 0.397 (Table 2.2)) but PERMANOVA was not significant for this postfire year. Thus, it indicates that the difference observed for the effect of salvage logging (PERMANOVA) for postfire years 3 and 7 was mainly due to species assemblages across treatments. There was no significant difference in beetle assemblages 15 years after fire (Table 2.2).

3.8 Beetles associated with residual or salvaged stands along the chronosequence

The analysis seeking species-habitat associations (Pearson’s point-biserial correlation) revealed that 6 species were significantly associated with residual stands exclusively, for at least one postfire year (Table 2.3). Species associated with residual stands of postfire year 3 were the wood borer Arhopalus foveicollis (Haldeman) and the sub-cortical species Dryocoetes autographus (Ratzeburg), Dryocoetes affaber (Mannerheim) and Polygraphus rufipennis (Kirby). A mycophagous species, Cartodere constricta (Gyllenhal), was associated only with residual stands of postfire years 3 and 5. The elaterid Ampedus fusculus (LeConte) was positively associated with residual stands of postfire year 5. Two other elaterid, Eanus decoratus (Mannerheim) and Ampedus pullus Germar, were positively associated with both residual and salvaged stands of postfire year 3 and 15 respectively. Thus, salvage logging does not affect those two last species within the chronosequence. The weevil Hylobius congener D. T., Schen. & Mars. was the only species associated with salvaged stands exclusively. This species was associated with salvaged stands of postfire years 3 and 15 but was consistently more frequent in salvaged stands.

62 4 Discussion

4.1 Effects of salvage logging on saproxylic beetle communities along the chronosequence

Our results showed that stumps left in postfire salvaged stands, even if they had similar water content as in residual stands, did not maintained similar abundance and species richness of saproxylic beetles for up to 3 years after fire as evaluated by capturing adults with trunk windows traps. Snags of commercial size constitute a large amount of resources for saproxylic species and their removal through salvage logging may have reduced local beetle populations. In chapter 1, we showed that beetle abundance and species richness remained relatively high in unsalvaged burned stands from 0 to 3 years after fire and dropped significantly between postfire years 3 and 5. We also showed that, in residual stands burned 3 years before, bark was still covering a large portion of black spruce tree stems (mean of 82.08 %) versus the later years (mean < 50 %). Therefore, it allows the colonisation of subcortical bark species like phloem feeders in the early postfire years and it probably explains the association of bark-beetles like Dryocoetes autographus (Ratzeburg), Dryocoetes affaber (Mannerheim) and Polygraphus rufipennis (Kirby) to residual stands of postfire year 3. Indeed, we probably caught specimens of those bark-beetles that emerged from an anterior colonisation. Moreover, residual bark and rotten wood on trunks could still provide appropriate oviposition sites for some xylem-borers whose larvae may feed on lower food quality. It seems to be the case for the cerambycid Arhopalus foveicollis (Haldeman) which was associated with residual stands affected by wildfire 3 years before being captured in all the four residual sites and in about equally number in both burns of this postfire year (site 1, burn 1: 5 specimens; site 2, burn 1: 32 specimens; site 3, burn 2: 21 specimens; site 4, burn 2: 8 specimens). Nappi et al. (2010) found Arhopalus foveicollis (Haldeman) in 8 and 11 years old burned stands and suggested that it has a long life-cycle. However, the exact duration of A. foveicollis life cycle is unknown. As for subcortical species discussed above, we cannot determine, in this study, if we caught colonizing individuals coming from neighbouring habitats or emerging adults that would have colonized burned trees soon after fire. In our study, this species was associated with residual stands 3 years after fire, even if it was found in all other postfire years. We caught

63 this species abundantly and exclusively in residual stands, 3 years after fire, and less frequently but only in salvaged stands 7 and 15 years after fire. This species could develop in both types of stands because larvae develop exclusively in the stump and probably in the roots of coniferous trees (Cadorette-Breton, 2013). However, we could hypothesize that the life cycle length of this species could vary on whether the stumps are located in residual or salvaged stands. Environmental variations generated by salvage logging may lower temperatures at night, increase wind velocity (Fontaine et al., 2010) and alter hydrological functions at ground level and by consequence, could alter conditions at stump level (Peterson et al., 2009). This could lengthen larval development and result in a longer life cycle.

Even if about 50 % of bark has fallen from snags 5 years after fire (see chapter 1), the association of some fungivorous species, like Cartodere constricta and Corticaria dentigera in residual stands of postfire years 3 and 5, suggests that subcortical ascomycete fungal communities (Muona and Rutanen, 1994; Wikars, 2002) can still grow until postfire year 5. These two species persisted until postfire year 5 in residual stands but, while Corticaria dentigera was associated to both types of stands for postfire year 3, our results revealed that Cartodere constricta was more affected by salvage logging, being associated only with residual stands for both postfire years 3 and 5.

In Boucher et al. (2012), two of our six species (Arhopalus foveicollis and Dryocoetes autographus) associated with residual stands, were also identified as burn-associated species. The six species identified to be mainly associated with residual stands in our study were found in other burns in Quebec as well as in other studies in Fennoscandia (Muona and Rutanen, 1994; Wikars, 1994; Saint-Germain et al., 2004a-b; Boucher et al., 2012). However, they were also found in mature unburned forests or in other type of disturbances (Saint-Germain et al., 2004c; Boucher et al., 2012). For instance, Polygraphus rufipennis was associated with small gap dynamic disturbance in Boucher et al. (2012). Therefore, the six species mainly associated with residual stands in our study are not fire-dependant to complete their life cycle but clearly take advantage of the large amount of dead wood produced by fire. Whether this build-up in their population dynamics is important for their

64 maintenance in other sources of dead or moribund trees until the next fire (Wikars, 2002) remains to be determined.

Despite the absence of difference in abundance and species richness between residual and salvaged stands of postfire years 5, 7 and 15, salvage logging seems to disturb the natural saproxylic beetle assemblages until postfire year 7. Removing snags obviously disturbed the early beetle assemblages, those that took advantage of the sudden abundance of fire- killed trees until postfire year 3. This modifies species composition as well as the abundance of species that were present. As time elapses in residual stands, saproxylic communities will change as a response to ongoing modification of wood properties (Simandl, 1993). Properties of deadwood such as decay classes, fungal colonization, and position of debris which vary with time since tree death are the main factors driving the colonization of insect communities in dead wood (Økland, 1996; Kaila et al., 1997). Even if we did not obtain significant difference in the amount of DDW between residual and salvaged stands, we think that the intake of woody debris to the ground is not the same between the two types of stands. In residual stands, a proportion of snags breaks and/or falls to the ground and become logs during the first 3 years after fire; (Saint-Germain and Greene, 2009); see chapter 1). Therefore, after some years, a new substrate may become suitable for a new beetle assemblage that cannot be observed in salvage stands where snags were removed and so, no new DDW can be generated. The similar amount of DDW in salvaged stands could be explained by the intake of logging debris during the salvage operations and/or, by the intake of merchantable salvaged wood left by the industries on the cutting patch due to excessive damages caused by wood-boring insects and woodpeckers (personal observations on the field). Thus, this DDW was generated all at once in salvaged stands (at time of salvage operations) compared with a gradual intake of new DDW in the residual stands during the first postfire years. Therefore, the DDW generated in the two stands did not occur at the same time and in the same condition of decomposition which could explain the difference in beetle assemblages 5 and 7 years after fire.

There was no difference in the abundance, species richness and beetle assemblages between residual and salvaged stands of 15 years postfire. Snags had lost about 50 % of their bark 5 years after fire (see chapter 1) and thus, trunks of fire-killed trees began to be harder

65 (Wikars, 1992) and the nutritional value of subcortical tissues rapidly deteriorates (Haack et al., 1987; Hanks, 1999). Only 15 % of bark was still present on trunks of snags 15 years postfire (see chapter 1). These denuded trunks were under strong sunlight, which slowed down wood decomposition as well as fungal development which is detrimental to saproxylic beetles (Gardiner, 1957; Amman, 1972; Haack et al., 1987) and other decomposer organisms (Laiho and Prescott, 2004). Thus, 15 years postfire, the activity rate of saproxylic beetles in both, residual and salvaged stands, was low because snags and stumps were poorly suitable for colonisation by most species.

4.2 Saproxylic beetles associated exclusively with salvaged stands

Hylobius congener was the only species mainly associated with salvaged stands exclusively. Even if this species was not very abundant, it was always more frequent in salvaged stands. This weevil is known to girdle root collar and make seedling debarking damage in plantation of various coniferous species (Welty and Houseweart, 1985). New seedlings established through plantation after salvage logging provide an alternative source of food for this species which is attracted to the harvested area by the odour of freshly exposed resin (Welty and Houseweart, 1985). Moreover, Martin (1964) showed that H. congener uses the inner bark of logs and stumps for food and breeding in pine plantations. Thus, the resin odour emitted from stumps of postfire stands may have attracted H. congener in salvaged stands in the early years.

4.3 Effect of salvage logging on elaterids’ association along the chronosequence

The exact biology of the elaterid species which have showed positive associations in this study is still unknown in Quebec (G. Pelletier, personal communication) and diverse genera such as Ampedus need revision as well as phylogenetic study (Jonhson, 2002). However, Eanus is a genus being related to genus Ctenicera (Du Chatenet, 2000) and thus, could also be predator which may explain its association with postfire year 3 where prey should be more abundant in both stands, residual and salvaged, compared with later years.

66 5 Conclusion

Concerning the saproxylic beetle assemblages, our results show that salvage logging has a disrupting effect on postfire beetle legacies that lasts up to 7 years after fire. This seems to be more important 3 years after fire, as it negatively affects both beetles’ abundance and species richness. This larger effect happens when beetle’s activity is high in unsalvaged burned stands (see chapter 1 of this document). Paradoxically, the early postfire years (1 to 3 years) are also those where the forest industry salvage trees aimed at timber production, since the quality of fire-killed trees deteriorates rapidly (e.g., wood boring insects, wood cleaving and decaying) (Saint-Germain and Greene, 2009). This clearly demonstrates the importance of finding measures that can contribute mitigating the effects of postfire logging in the early postfire years. However, the sole presence of the stumps as snags remnants in salvaged stands seems to be enough as a mitigating measure for some species. Indeed, some species mainly associated with residual stands in this study (e.g. Polygraphus rufipennis and Cartodere constricta) are also common in unburned boreal forest (Légaré, 2010; Boucher et al., 2012). Residual stumps from salvage logging, though, seem to have a certain ecological value for some species, such as Ampedus pullus and Eanus decoratus, and thus could lessen the impact of this practice for such species. The forthcoming harvesting of stumps, in salvaged stands, for bioenergy purposes (Victorsson and Jonsell, 2013) or even the use of scarification in site rehabilitation would reduce the availability and the quality of this alternative habitat, thus worsen the impacts of salvage logging on saproxylic beetle legacies. Our results on both habitat attributes and beetle assemblages, show that salvage logging done over the last 20 years has modified, at the stand level, the natural ecological trajectories of postfire legacies, which is in contradiction with ecosystem management that aims to reduce the gap between managed and natural forests.

Acknowledgments

67 appreciated the help of G. Pelletier, taxonomist at CFS-NRCan, for taxonomic training and for assistance for validating identifications. We also express our gratitude to J. Boucher from l’Université Laval for experimental design, site selection and statistical assistance and to L. Préfontaine for help with the English revision of this manuscript. Finally, we thank the Laurentian Forestry Centre for laboratory facilities. This study was supported by the Fonds Québécois de Recherche sur la Nature et les Technologies (FQRNT) through the Programme de recherche en partenariat sur l’aménagement et l’environnement forestiers-II, by the iFor consortium (Université Laval), and by the Canadian Forest Service.

68 Tables

Table 2. 1 PERMANOVA results showing the effect of salvage logging, time since disturbance and their interaction on beetle assemblages. Significant effects are outlined in bold (α = 0.05).

PERMANOVA Df MS F R2 p Salvage logging (SL) 1 0.410 2.945 0.059 0.0075 Time since disturbance (TSD) 3 0.826 5.930 0.358 0.0005 SL × TSD 3 0.233 1.672 0.101 0.0385 Residuals 24 0.139 0.482 Total 31 1.000

Table 2. 2 PERMANOVA results showing the by-year effects of salvage logging on postfire beetles assemblages. Significant effects are outlined in bold (α = 0.05).

PERMANOVA PERMDISP 15 years after fire d.f. MS F R2 p d.f. MS F p Treatment (residual vs salvaged stands) 1 0.262 1.51 0.20 0.094 1 0.001 0.317 0.581 Analysis residuals 6 0.174 0.80 6 0.003 Total 7 1.00 7 years after fire Treatment (residual vs salvaged stands) 1 0.497 2.11 0.33 0.033 1 0.004 2.318 0.184 Analysis residuals 6 0.167 0.67 6 0.002 Total 7 1.00 5 years after fire Treatment (residual vs salvaged stands) 1 0.228 1.34 0.18 0.217 1 0.015 9.934 0.030 Analysis residuals 6 0.170 0.82 6 0.002 Total 7 1.00 3 years after fire Treatment (residual vs salvaged stands) 1 0.120 2.70 0.31 0.031 1 0.00001 0.008 0.931 Analysis residuals 6 0.045 0.69 6 0.001 Total 7 1.00

Table 2. 3 Beetle species significantly associated (p <0.05) with residual or salvaged stands of a postfire year or a combination of postfire years from point - biserial group-equalized phi coefficient analysis (Pearson correlation). Number of individuals in bold refers to the total n umber of specimens captured for each treatment and year with which a species was positively associated.

Families Species Time after fire (year) Phi p and treatment coefficient 3 5 7 15

Res. Sal. Res. Sal. Res. Sal. Res. Sal. Cerambycidae Arhopalus foveicollis (Haldeman) 66 0 2 8 0 9 0 2 0.789 0.001 Curculionidae Dryocoetes autographus (Ratzeburg) 475 181 23 5 2 0 2 0 0.856 0.001 Hylobius congener D. T., Schen. & Mars. 1 5 1 3 0 2 1 9 0.630 0.029 Dryocoetes affaber (Mannerheim) 4 1 1 0 0 0 0 0 0.663 0.044 Polygraphus rufipennis (Kirby) 31 10 2 1 0 0 0 0 0.589 0.013 Lathridiidae Corticaria dentigera LeConte 50 60 40 7 4 5 7 6 0.841 0.001 Cartodere constricta (Gyllenhal) 22 7 19 1 5 0 5 2 0.582 0.048 Elateridae Eanus decoratus (Mannerheim) 19 17 3 2 2 0 2 0 0.831 0.001 Ampedus fusculus (LeConte) 0 0 8 1 2 0 0 0 0.634 0.026 Ampedus pullus Germar 0 1 3 6 7 21 28 24 0.694 0.007

71 Figures

Figure 2. 1 Study area showing selected burns with their respective time (year) since wildfire. The X mark represents a 15 year burn that was smaller than 275 hectares.

Figure 2. 2 Student paired t-test’s results comparing A) the basal area (m²/ha) of snags (average ± standard error), B) the volume (m³/ha) of DDW (average ± standard error) and C) the basal area of dead saplings (average ± standard error) between residual and salvaged stands for each postfire years. Values in bold indicate significant results (p < 0.05).

73 A)

250 p = 0.0184 p = 0.2165 p = 0.6325 p = 0.4917

200 Residual Salvaged

150

Abundance Abundance Abundance 100

0 3 5 7 15 B)

p = 0.0465 p = 0.0582 p = 0.4714 p = 0.6124 20

Richness species species Richness Richness

0 3 5 7 15 Time after fire (years) Figure 2. 3 Student paired t-test’s results comparing A) the abundance (average ± standard error) and B) the species richness (average ± standard error), between residual and salvaged stands for each n umber of years after fire. Values in bold indicate significant results (p < 0.05).

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79 Appendix A2 Species list and number of beetles captured for each treatment of each postfire year in black spruce boreal forest.

Time after fire (year) and treatment Families Species 3 5 7 15 Res. Sal. Res. Sal. Res. Sal. Res. Sal. Total Buprestidae Buprestis maculiventris Say 0 0 1 0 0 0 0 0 1 Cerambycidae Acmaeops pratensis (Laicharting) 4 1 4 2 7 0 0 0 18 Acmaeops proteus proteus (Kirby) 1 1 0 0 0 0 0 0 2 Acmaopsoides rufula (Haldeman) 0 0 0 0 0 0 0 2 2 Arhopalus foveicollis (Haldeman) 66 0 2 8 0 9 0 2 87 Clytus ruricola (Olivier) 0 0 0 0 2 1 0 0 3 Evodinus monticola (Randall) 0 0 0 1 0 0 0 0 1 Lepturobosca chrysocoma (Kirby) 0 0 0 2 0 1 0 0 3 Monochamus scutellatus scutellatus (Say) 0 1 0 0 0 0 0 0 1 Pachyta lamed liturata Kirby 1 0 0 0 0 0 0 0 1 Pogonocherus mixtus Haldeman 0 0 1 0 0 0 0 0 1 Pygoleptura nigrella nigrella (Say) 0 1 1 0 0 1 2 0 5 Rhagium inquisitor (Linné) 1 1 1 0 0 0 0 1 4 Stictoleptura canadensis canadensis (Olivier) 0 0 2 1 8 7 2 1 21 Trachysida mutabillis (Newman) 0 0 0 0 0 1 0 0 1 Cleridae Thanasimus dubius (Fabricius) 0 0 1 0 0 0 0 0 1 Thanasimus undatulus nubilus (Say) 5 1 2 0 0 0 0 1 9 Corylophidae Clypastraea fusca (Harold) 3 1 0 0 0 0 0 0 4 Curculionidae Auleutes epilobii (Paykull) 0 0 0 1 0 1 0 0 2 Cossonus americanus (Buchanan) 0 0 0 0 0 0 1 0 1 Crypturgus borealis Swayne 0 0 2 0 0 0 0 0 2 Dryocoetes affaber (Mannerheim) 4 1 1 0 0 0 0 0 6 Dryocoetes autographus (Ratzeburg) 475 181 23 5 1 0 1 0 686 Dryocoetes betulae Hopkins 3 0 1 0 0 0 0 0 4 Hylobius congener D. T., Schen. & Mars. 1 5 1 3 0 2 1 8 21 Ips latidens (LeConte) 0 0 0 0 0 0 0 2 2 Lepyrus palustris (Scopoli) 0 0 0 0 0 0 0 1 1 Pissodes strobe (Peck) 0 0 0 0 0 0 0 1 1 Pithyophthorus sp. Eichhoff 1 0 1 1 0 0 0 0 3 Polygraphus rufipennis (Kirby) 31 10 2 1 0 0 0 0 44 Rhyncolus brunneus Mannerheim 0 0 1 0 0 0 0 0 1 Rhyncolus macrops Buchanan 0 0 0 0 0 0 2 0 2 Tripodendron lineatum 1 0 0 0 0 0 0 0 1 Xyleborus sayi (Hopkins) 0 0 0 1 0 0 1 0 2 Elateridae Agriotes limosus (LeConte) 2 3 18 25 9 3 6 2 68 Ampedus apicatus (Say) 0 0 0 0 1 0 0 0 1 Ampedus evansi W. J. Brown 0 0 1 1 0 0 0 0 2 Ampedus fusculus (LeConte) 0 0 8 1 2 0 0 0 11 Ampedus laurentinus W. J. Brown 5 0 4 2 1 1 2 0 15 Ampedus luctuosus (LeConte) 0 0 2 1 7 0 7 2 19 Ampedus mixtus (Herbst) 2 0 1 1 2 4 6 1 17 Ampedus nigrinus (Herbst) 0 0 3 2 0 1 6 0 12 Ampedus pullus Germar 0 1 3 6 7 21 28 24 90 Ampedus quebecensis W. J. Brown 3 0 4 3 1 1 3 0 15 Ampedus sp. Dejean 0 0 1 0 0 0 0 0 1 Dalopius cognatus W. J. Brown 0 0 0 0 0 0 1 0 1 Dalopius sp. Eschscholtz 0 0 0 0 0 0 1 0 1 Denticollis denticornis 0 0 0 0 0 0 2 0 2 Eanus decoratus (Mannerheim) 19 17 3 2 2 0 2 0 45 Hypnoidus bicolor (Eschscholtz) 0 0 0 0 0 0 0 3 3 Limonius aeger LeConte 0 0 0 1 0 0 0 0 1 Liotrichus spinosus (LeConte) 3 3 4 0 0 0 0 1 11 Neohypdonus tumescens (LeConte) 0 0 1 3 0 2 0 0 6 Pseudanostirus triundulatus (Randall) 7 6 22 18 22 14 17 15 121 Selatosomus appropinquans (Randall) 0 0 0 0 0 1 0 1 2 Sericus incongruous (LeConte) 4 3 4 17 3 8 6 0 45 Setasomus arata (LeConte) 0 0 0 0 0 0 0 1 1 Setasomus nitidula (LeConte) 0 0 0 0 1 1 1 1 4 Lathridiidae Cartodere constricta (Gyllenhal) 22 7 19 1 5 0 5 2 61

80 (Appendix A2 Continued) Time after fire (year) and treatment Families Species 3 5 7 15 Res. Sal. Res. Sal. Res. Sal. Res. Sal. Total Lathridiidae Corticaria brevicornis Fall 13 11 15 4 10 1 10 4 68 Corticaria dentigera LeConte 50 60 40 7 4 5 7 6 179 Cortinicara gibbosa (Herbst) 0 0 1 0 1 0 0 0 2 Lathridius sp. Herbst 0 0 0 1 2 2 0 1 6 Melanophthalma pumila (LeConte) 0 0 0 0 1 0 0 0 1 Melanophthalma sp. Motschulsky 0 0 0 0 2 0 0 0 2 Salpingidae Sphaeriestes virescens (LeConte) 21 16 2 0 1 1 3 9 53

Conclusion générale

Nos résultats démontrent l’importance des brûlis pour les successions naturelles de huit familles de coléoptères fortement associées aux peuplements brûlés d’épinettes noires. En plus de confirmer l’importance des trois premières années pour l’abondance et la richesse en espèces des coléoptères saproxyliques, notre étude a dévoilé une succession rapide de quatre phases distinctes d’assemblages d’espèces durant les 15 premières années après le passage du feu. Parallèlement, plusieurs études rapportent que les communautés de coléoptères saproxyliques sont plus abondantes et plus diversifiées dans les forêts boréales brûlées comparativement à celles en forêt non-brûlées (Saint-Germain et al., 2004c; Moretti et al., 2010). Ces mêmes études soulignent, de plus, l’existence d’assemblages de coléoptères saproxyliques spécifiques aux forêts brûlées. Dans la même veine, l’étude de Boucher et al., (2012), comparant la structure des assemblages de coléoptères saproxyliques entre des brûlis et des peuplements annelés (simulation de peuplements soumis à une dynamique de perturbations par petites trouées avec des chicots récents), révèle, elle aussi, l’existence d’assemblages de coléoptères spécifiques aux brûlis. Malgré que certaines espèces semblent avoir un intérêt pour les peuplements récupérés contenant des souches, nos résultats indiquent que la coupe de récupération influence négativement l’abondance et la richesse en espèces durant les premières phases de colonisation, soit lorsque les coléoptères saproxyliques sont les plus abondants. De surcroît, nos travaux démontrent que la coupe de récupération modifie la succession naturelle des coléoptères saproxyliques en affectant les assemblages d’espèces jusqu’à sept ans après feu. Ces résultats appuient ceux de Boucher (2011) qui démontraient que la récupération avait des effets qui se répercutaient jusqu’à un minimum de 3 ans après le passage du feu.

Dynamique temporelle après feu

Les divers types de peuplements de la forêt boréale présentent des stocks de combustibles différents et sont donc affectés différemment par le feu, tant au niveau de la probabilité qu’il y ait un feu que de l’intensité qu’il peut atteindre lorsqu’il est déclenché (Gauthier et al., 2008). Autant de trajectoires de rétablissements écologiques peuvent suivre un événement de feu (Weber et Stocks, 1998). Ces trajectoires se manifestent à travers des

83 modifications dans la composition et la structure d’âges des espèces végétales, dans le recyclage des éléments nutritifs, dans la productivité de l’habitat et dans la biodiversité (Volney et Hirsch, 2005). L’évolution du peuplement brûlé procurera donc une large diversité temporelle de réponses par les organismes qui les colonisent. En ce sens, nos résultats confirment cette idée puisqu’ils rapportent des changements successifs dans les communautés et les assemblages d’espèces durant les 15 premières années après feu. De par leurs différents rôles trophiques, les assemblages d’insectes saproxyliques qui se succèdent après un feu jouent un rôle important dans la succession des écosystèmes brûlés. En effet, ces organismes contribuent à la décomposition du bois et au recyclage des éléments nutritifs ce qui rend le milieu plus productif (Speight, 1989; Boulanger et Sirois, 2007; Cobb et al., 2010) en plus de servir de proie pour d’autres organismes saproxyliques (Morissette et al., 2002; Purdon et al., 2002). D’ailleurs, les résultats présentés dans le premier chapitre de ce document ont identifié 17 espèces associées avec l’une ou plusieurs des trois premières années après le passage du feu. Parmi ces 17 espèces, on retrouvait effectivement une diversité d’organismes comprenant des xylophages, des sous-corticaux, des mycophages, des saprophages, ainsi que des prédateurs ayant tous une implication plus ou moins importante dans le fonctionnement de l’écosystème. Nos résultats ont démontré qu’en retirant le tronc des chicots, les coupes de récupération réduisent à court terme l’abondance et la richesse spécifique des 8 familles de coléoptères saproxyliques associée au bois mort. En effet, cela contribue à simplifier la composition et la structure de l’habitat brûlé.

Quant aux résultats présentés dans le deuxième chapitre de ce document, ils indiquent que quatre espèces ont une association à court terme (3 ans après feu) avec les peuplements résiduels. Il s’agit d’Arhopalus foveicollis, Dryocoetes autographus, Polygraphus rufipennis et Dryocoetes affaber. Dans le cas de Corticaria dentigera, malgré qu’elle soit associée aux deux types de peuplements 3 ans après feu, son association avec les sites résiduels persiste jusqu’à 5 ans après feu. Cette association plus durable avec les peuplements résiduels est aussi vraie pour Cartodere constricta. Parmi toutes les espèces étudiées, ce sont les populations de ces six espèces qui sont les plus susceptibles d’être affectées par les changements induits par les coupes de récupération. De plus, A. foveicollis, a également été identifiée comme une espèce associée aux brûlis, lorsque comparée à des

peuplements annelés (Boucher et al., 2012). Notre étude confirme donc la forte association de cette espèce avec les peuplements brûlés et par conséquent, sa fragilité potentielle face aux coupes de récupération. Les cinq autres espèces ont déjà été répertoriées en grand nombre en forêt boréale non perturbée du Québec (Légaré, 2010). Cependant, leur présence dans les brûlis démontre leur comportement opportuniste à profiter de l’abondance de bois soudainement mort après le passage du feu. Ainsi, les feux de forêt sont importants pour le maintient du niveau des populations de ces espèces opportunistes qui utilisent les brûlis pour accroitre leur abondance dans le paysage. La réduction de l’abondance du bois mort causée par les coupes de récupération ne procure pas seulement des effets négatifs à court ou à moyen terme sur l’abondance et la richesse en espèces des coléoptères saproxyliques. Cette pratique pourrait également avoir des conséquences plus durables sur leurs populations et même sur les métapopulations des espèces qui auraient évolué en développant des adaptations au feu et dont leur persistance à long terme dans le paysage serait dépendante des brûlis (Wikars, 2002; Boucher, 2011).

Aménagement écosystémique des brûlis

Le bois mort sur pied ou au sol constitue un attribut-clé du système forestier et son importance pour le maintien de la diversité biologique et la résilience des écosystèmes à la suite d’une perturbation n’est plus à démontrer (Bull, 1983; Siitonen, 2001; Saint-Germain et al., 2004b; Gauthier et al., 2008; Drapeau et al., 2010). Dans un contexte de mise en œuvre de l’aménagement écosystémique lors de la récupération du bois mort après feu, nos pratiques, telles qu’appliquées jusqu’à maintenant, vont à l’encontre du premier critère de développement durable. En effet, plutôt que de maintenir l’un des attributs-clés que produisent les perturbations naturelles, soit le bois mort, les coupes de récupération le retirent de l’environnement. Cela soulève un paradoxe intéressant; en effet, alors qu’on tente de s’inspirer des perturbations naturelles pour aménager la forêt boréale, nous récoltons le bois directement dans ces perturbations (Schmiegelow et al., 2006).

Dans l’optique de maintenir l’intégrité des différents écosystèmes et ainsi, mieux orienter les coupes de récupération au Québec, de nouvelles orientations pour un aménagement écosystémique des brûlis sont devenues primordiales et quelques suggestions ont été

85 apportées afin de modifier les plans spéciaux d’aménagement (Nappi et al., 2011). Le seuil de 30 % de forêt brûlée à conserver établi dans le document de Nappi et al, (2011) et présenté plus tôt dans l’introduction générale de ce document, a été inspiré de la littérature scientifique concernant les effets de la fragmentation de l’habitat sur les oiseaux et/ou les mammifères (Andrén, 1994; Radford et al., 2005; Nappi et al., 2011). Par conséquent, ce seuil de 30 % n’a pas été établi spécifiquement pour les brûlis et les espèces qui y sont associées. De plus, dans la littérature, le seuil de 30 % doit être constitué d’habitats ayant été définis comme étant convenables pour les organismes qui les utilisent (Andrén, 1994; Hanski, 2005). Ainsi, il faudrait s’assurer, d’une part, que le seuil de 30 % de forêts brûlées à conserver soit suffisant pour le cas spécifique des brûlis et, d’autre part, qu’il soit représentatif de toute la variabilité d’habitats, ayant préalablement été définis convenables, pour la diversité associée aux brûlis et ce, à long terme. L’exploitation de 70 % des brûlis jumelée à une rétention sans aucune intervention dans 30 % des autres brûlis pourrait également être une option envisageable. Ainsi, dans les brûlis non récupérés, l’ensemble des caractéristiques retrouvées à la suite du passage du feu pourrait être conservées tout en sauvant du temps et de l’énergie à définir ce qui est réellement essentiel pour l’ensemble des organismes utilisant les brûlis. Une meilleure compréhension concernant les impacts de divers niveaux de récupération après feu sur les réponses écologiques pourrait également permettre de modifier les seuils minimaux de forêt brûlée à conserver pour les brûlis. Ces seuils pourraient également varier selon les conditions de chacun des peuplements à l’intérieur d’un brûlis. Cependant, les longicornes, notamment M. s. scutellatus, creusent des galeries dans le bois dès la première année ce qui déclasse le bois pour l’industrie forestière. Il est donc primordial de concevoir les plans de récupérations rapidement ce qui laisse peu de temps pour évaluer et conserver toute la variabilité des peuplements au sein d’un même brûlis. Au Québec, le développement de nouvelles pratiques concernant l’aménagement des brûlis pourrait aussi s’inspirer en partie de certains pays. Dans les forêts boréales de Fennoscandie, certaines espèces d’insectes dépendant du bois mort sont parmi les plus à risque d’extinction à cause de la réduction de bois mort disponible (Hyvärinen et al., 2006), causée par l’aménagement forestier intensif (Siitonen, 2001) et la suppression extensive des feux de forêt (Wikars, 1992). Face à cela, des stratégies par filtre fin ont été développées afin d’assurer le maintien d’habitats spécifiques pour les espèces les plus rares

pour lesquelles, au Québec, aucune équivalence n’a encore été déterminé. Dans des forêts suédoises non-brûlées, une étude a démontré l’importance de laisser des souches plus hautes (4 m de hauteur) lors des interventions de récoltes, pour la conservation de certaines espèces d’insectes associées au bois mort et en danger d’extinction (Jonsell et al., 2004). Au Québec, la problématique du bois mort n’est pas identique à celle des pays de Fennoscandie cependant, l’ajout de telles pratiques aux normes d’interventions actuelles dans les brûlis du Québec pourraient tout de même être considéré et pourrait être appliqué à titre préventif et conservateur.

Il serait aussi important de définir le statut des espèces fortement associées aux brûlis et moins représentées dans les environnements générant d’autres sources de bois mort (Boucher et al., 2012), afin d’instaurer des stratégies adéquates permettant leur maintien dans le paysage. Ainsi, des études à caractère fondamental devraient être entreprises afin de mieux connaître la biologie et le statut de certaines espèces fortement associées au bois brûlé comme par exemple, Acmaeops pratensis, une espèce circumpolaire se retrouvant sur la liste rouge des espèces en danger d’extinction en Europe (Nieto et Alexander, 2010).

Finalement, il faudrait également s’interroger sur la nécessité de certaines pratiques post- récupérations comme la mise en andain et le scarifiage, lesquelles sont habituellement réalisées avant la remise en production. Ces deux pratiques visent à faciliter le reboisement et favoriser la croissance des plants en assurant une meilleure circulation de l’air et de l’eau dans le sol et en rendant les éléments nutritifs du sol accessibles (Doucet et Côté, 2009). Malgré ces avantages, ces deux pratiques comportent une conséquence majeure, soit l’élimination et la réduction de la qualité des deux principales sources de bois mort résiduel à la suite des coupes de récupération : les souches et les débris ligneux au sol. En effet, même si les souches restent présentes sur le territoire à la suite de la scarification et de la mise en andain, celles-ci sont fortement fragmentées par la machinerie, ce qui réduit l’habitat accessible et entraîne une perte d’humidité plus rapide, défavorisant ainsi les organismes saproxyliques. De plus, face à cette problématique, il faudrait aussi s’interroger sur l’introduction de nouvelles pratiques consistant à récupérer ce qui reste de bois mort afin de produire de la bioénergie. En Europe, jusqu’à 75 % du volume des souches peut être extrait à la suite des coupes forestières (Eräjää et al., 2010; Rabinowitsch-Jokinen et

87 Vanha-Majamaa, 2010; Victorsson et Jonsell, 2013). Ces souches sont, généralement, empilées à côté des sites d’extraction pour une période de 2 ans, avant d’être transformées en biocarburants (Victorsson et Jonsell, 2013). Dans leur étude, Victorsson et Jonsell (2013) ont comparé l’abondance des coléoptères entre des sites de coupes avec extraction de souches et des sites de coupes sans extraction, et ils ont observé une diminution de l’abondance des coléoptères de l’ordre de 70 % dans les sites avec extraction. Les auteurs ont également affirmé que cette pratique constituait un sévère piège écologique pour quatre espèces. En effet, ces quatre espèces étaient plus abondantes dans les empilements de souches extraites que dans les sites coupés adjacents et par conséquent, elles ont fort probablement été tuées lors des processus de transformation en biocarburants et ce, avant qu’elles ne puissent se reproduire. Dans un autre ordre d’idée, cette pratique pourrait s’avérer bénéfique pour les espèces saproxyliques plus abondantes dans les empilements de souches extraites si celles-ci étaient laissées suffisamment longtemps sur le territoire pour permettre de compléter le développement larvaire de ces espèces. Cependant, les résultats de cette étude démontrent clairement qu’il serait préférable d’être très vigilant et imaginatif avant d’adopter de telles pratiques à la suite des coupes de récupération au Québec.

Investigations futures

Bien que le piège à impacts troncaux soit reconnus pour être efficaces dans la capture des coléoptères saproxyliques susceptibles d’utiliser le substrat pour se reproduire et compléter leur cycle vital (Kaila, 1993), ce piège nous informe essentiellement sur le taux d’activité de ces coléoptères dans le milieu étudié. Il serait donc inexact de prétendre que tous les spécimens échantillonnés dans notre étude colonisent le bois brûlé. En effet, ce piège permet aussi la capture d’espèces touristes, probablement moins abondantes, n’étant que de passage et conséquemment, n’utilisant pas nécessairement le substrat à l’étude où même le substrat brûlé. La récolte et l’encagement des différents types de bois mort retrouvés dans un brûlis est une alternative complémentaire permettant la mise en élevage des coléoptères saproxyliques directement relié avec leur habitat. Ainsi, il devient possible d’étudier la relation directe entre un spécimen et l’habitat qu’il utilise pour se développer.

Dans le cadre d’un projet connexe, nous avons, à l’été 2011, récolté trois sections troncales de chicots, trois sections de débris ligneux au sol et trois souches avec une portion de leur système racinaire dans tous les sites résiduels à la coupe de récupération. Nous avons également récolté trois souches avec une portion de leur système racinaire dans tous les sites récupérés. Tous ces échantillons ont été encagés et entreposés, à l’abri des intempéries, au Centre de Foresterie des Laurentides à Québec. Ainsi, ce dispositif permettra de comparer les populations de coléoptères saproxyliques entre les différents habitats disponibles après feu et susceptibles d’être colonisés. Selon les différents rôles trophiques (xylophage, mycophage, saprophage, etc.) des coléoptères saproxyliques, nous devrions observer des successions dans les populations en relation avec les changements temporels des différents habitats (température, humidité, densité, etc.). Ce dispositif permettra d’étudier de façon plus spécifique la valeur écologique des souches laissées sur le territoire à la suite des coupes de récupération et de la comparer à celle des autres habitats retrouvés dans les peuplements résiduels.

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