Changes in sporocarp production and vegetation following wildfire in a Mediterranean Forest Ecosystem dominated by Pinus nigra in Northern Spain

Forest Ecology and Management 331 (2014) 85–92

Contents lists available at ScienceDirect

Forest Ecology and Management journal homepage: www.elsevier.com/locate/foreco

Changes in sporocarp production and vegetation following wildfire in a Mediterranean Forest Ecosystem dominated by Pinus nigra in Northern Spain Olaya Mediavilla, Juan Andrés Oria-de-Rueda 1, Pablo Martín-Pinto ⇑ Sustainable Forest Management Research Institute, Fire and Applied Mycology Laboratory, Departments of Agroforestry Sciences, and Vegetal Production and Natural Resources, University of Valladolid (Palencia), Avda. Madrid 44, 34071 Palencia, Spain

a r t i c l e

i n f o

Article history: Received 9 May 2014 Received in revised form 25 July 2014 Accepted 29 July 2014

Keywords: Fire effects Fungal production Fungal composition Host preference

a b s t r a c t In Mediterranean forests, wildfires are a common feature which profoundly alters vegetation and its associated fungal communities. While a great deal of research has been advocated to the study of plant communities affected by forest fires, our knowledge on the interactions between fire occurrence and development of fungal communities is still scarce. The aim of this work was to study the changes triggered by wildfires in the mycoflora of a Pinus nigra artificial stand in Northern Spain. Sporocarps were collected and identified from a set of three 100 m2 transects at each one and five year old burned areas and an unburned adjacent area. Then, fungal species richness, biomass production and species composition was analyzed as dependent on time after fire, and also considering aspects as edibility and fungal life form. Sporocarp production and mycorrhizal and edible species richness were strongly affected just after fire, but few differences respect to unburned areas were observed only five years after the disturbance. Also, specific fungal communities composition was correlated with successive stages after fire. This was likely because of the different vegetation composition found at different stages, with species typically connected to Pinus, Quercus and Cistus in the areas where each one of them predominated. Promoting a mixture of host species just following fire by leaving the pioneer species during the implantation of new forest stands, could result in a prompt recovery of the associated fungal community, adding extra ecological value to these forests. Ó 2014 Elsevier B.V. All rights reserved.

1. Introduction Forest wildfires are considered as the main natural disturbance in Mediterranean ecosystems and therefore one of the most important historical factors shaping the ecological landscape in these areas (Keeley et al., 2011). Pinus, Quercus and Cistus are widespread Mediterranean species. As a result of adaptation to a regime of frequent wildfires, they show a set of different fire-resilience traits like resprouting or rapid after fire germination, that enable them to efficiently recover after fire (Acácio and Holmgren, 2014; López et al., 2009; Pausas et al., 2008). Forest wildfires not only strongly affect vegetation dynamics, but also disturb their associated fungal communities (Bastias ⇑ Corresponding author. Tel.: +34 979108340; fax: +34 979108440. E-mail addresses: [email protected] (O. Mediavilla), oria@agro. uva.es (J.A. Oria-de-Rueda), [email protected] (P. Martín-Pinto). 1 Tel.: +34 979108364; fax. +34 979108440. http://dx.doi.org/10.1016/j.foreco.2014.07.033 0378-1127/Ó 2014 Elsevier B.V. All rights reserved.

et al., 2006). A dramatic change in ecological conditions due to fire triggers fungal succession (Visser, 1995). This succession depends on multiple factors, such as the intensity and duration of fire (Hart et al., 2005). Previous studies show that low-intensity fires do not seem to significantly affect ectomycorrhizal community richness (Jonsson et al., 1999). Nonetheless, high-intensity fires are detrimental to the majority of fungal species at least in the short-term (Jiménez Esquilín et al., 2007). Fungal succession patterns are partly the results of the unequal impact that forest fires have on different fungal species. For example, some saprotrophic and/or pyrophitic fungi benefit from fire since new ecological conditions favor their reproduction and growth in absence of competition (Gassibe et al., 2011). However, other species (mainly mycorrhizal) suffer the most negative consequences due to the destruction of theirs hosts (Longo et al., 2011). It is believed that the high resilience of Mediterranean ecosystems (Bradshaw et al., 2011) is partly due to the presence of active ectomycorrhizal fungal propagules in the soil (Buscardo et al.,

86

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92

2012; Izzo et al., 2006). Indeed, fungal symbiosis is crucial for postfire regeneration and survival of Pinus spp. seedlings (Buscardo et al., 2009), a widespread genus in the Mediterranean. Specifically, Pinus nigra is an example of Mediterranean species facing regeneration problems due to recurrent wildfires (Cerro-Barja et al., 2006). For this species, the role played by mycorrhizal symbiosis is crucial, since it facilitates survival and growth of the post-fire seedlings, thus promoting the natural recovery of the affected areas (Claridge et al., 2009). This positive effect is strongly conditioned upon the presence of other bridge pioneer host species, acting as reservoirs for ectomycorrhizal fungi propagules in the soil (Hernández-Rodríguez et al., 2013). Some examples are Quercus spp. and Cistus spp., widely distributed as understory species in Pinus spp. stands (Comandini et al., 2006; Moreno and Oechel, 1994) in Spain as well as in many other Mediterranean countries frequently affected by forest fires, e.g. Italy, Greece or Turkey (Moreno and Oechel, 1994). Despite these forest ecosystems being widespread, little is known about the influence of understory host species on fungal succession following wildfires, being P. nigra a particularly overlooked species. A previous work analyzed fungal production and richness associated with P. nigra stands in northeastern Spain. However, the studied forest was naturally regenerated and located on limestone susbtrate (Bonet et al., 2010). It is known that fungal communities are significantly affected by factors such as regeneration mode, soil type or forest management of the stand (Luoma et al., 2004; Oriade-Rueda et al., 2010; Smith et al., 2002). Hence, due to the importance of species coexistence, as well as the lack of studies related to fungal succession in P. nigra under varying conditions, the goal of this study was to assess the changes in sporocarp production linked to vegetation succession under a P. nigra artificial stand on siliceous soil, which was repeatedly affected by fire. Our specific aims were: (1) to analyze and compare fungal sporocarp richness in two burned areas, one and five years after fire occurred, as well as in unburned nearby areas; (2) to analyze sporocarp production according to nutritional mode (saprotrophic or ectomycorrhizal) and edibility; (3) to assess taxa composition following fire and according to predominant hosts.

2. Materials and methods 2.1. Study site The study area was placed in La Valdavia (Palencia, Spain). This region is characterized by a Mediterranean climate, with a dry season during summer (three months) and low temperatures and frosts during winter. The mean annual precipitation is 560 mm and the mean temperature 10.5 °C, these climatic data were provided by the closest meteorological station (Saldaña, 357471 Longitude-UTM, 4709374 Latitude-UTM and 912 m above sea level, Spanish Meteorological Agency). The area is composed of cluster, siliceous sands and shales. Soils are mainly clayey with low permeability, low organic matter content and pH ranging from moderately acidic to moderately basic. Quercus pyrenaica used to constitute the climatic dominant species in the area, but forests have been intensively exploited for wood since ancient times and only coppice stands remained until recent years. In an attempt to restore forest ecosystems, extensive areas were successfully afforested with several pine species during the 1950s and 1960s. Our study was placed in a representative 50 year old P. nigra subsp. austriaca stand spanning 1367 ha, and where Q. pyrenaica can be found in the understory at low frequency (950–960 m a.s.l., 376090 Longitude-UTM, 4711724 Latitude-UTM). Two high intensity wildfires affected two non-overlapping areas within the stand. The first wildfire took place in 2002

(118.03 ha burned). All the pines were strongly affected, so consequently the trees were cut and the area was reforested with P. nigra subsp. salzmannii. Before planting, the soil was ploughed in order to facilitate the plantation. During the summer of 2006, a second wildfire took place in a different nearby area, affecting 135.37 ha. As in the first wildfire, all affected trees were removed and the area reforested with P. nigra subsp. salzmannii. In both cases, shortly after reforestation, Cistus laurifolius rapidly took over the entire pine stand. As a result, the studied areas were composed by different plants species known to potentially host numerous fungal species (Bruns et al., 2002; Comandini et al., 2006). Specifically, at the time of data collection in 2007, there were three areas clearly differentiated: (1) Unburned area (UB): formed by a 50 year old P. nigra subsp. austriaca stand, with some scattered Q. pyrenaica root-sprouts and few dispersed C. laurifolius individuals; (2) Five year-old burned area (B02): made up of five year-old P. nigra subsp. salzmanni, with a fraction of canopy cover around 65–70%, some Q. pyrenaica individuals and abundant C. laurifolius; (3) One year-old burned area (B06): formed by recently reforested P. nigra subsp. salzmanni with abundant remains of burned pine and oak stems, a stratum of C. laurifolius seedlings and Q. pyrenaica root sprouts. 2.2. Sampling design Ecological conditions across all areas in the studied forest stand were homogenous, in terms of climate, soil, slope and altitude. Besides, special care was taken to avoid confounding factors when selecting those areas where the sampling was finally done. Differences in fungal diversity and productivity among areas prior to fire were thus unlikely. Following previous studies, three 2  50 m plots were established at each area (UB, B02 and B06), i.e. nine plots in total (Luoma et al., 1991; Ohenoja, 1989; Smith et al., 2002). All sporocarps found within the plot limits, were systematically collected during autumn, from October to December 2007. Sampling was carried out once a week following previous studies (Dahlberg, 1991; Ohenoja, 1984). Sporocarps were fully harvested, enabling an accurate identification of fungal species and avoiding bias in biomass estimations. Samples were transferred to the laboratory, stored at 4 °C and processed within 24 h after collection for taxonomic identification and fresh weight (kg ha1) measurements. Then samples were introduced in a drying oven at 35 °C until constant weight (3–4 days) and dry weight was subsequently measured. 2.3. Identification and classification Following several taxonomical references, sporocarps were identified at the species level whenever possible (AndrésRodríguez et al., 1999; Antonín and Noordeloos, 2010; Knudsen and Vesterholt, 2008). Some samples were only identified to the genus level and were conservatively grouped by genus rather by unclassified Operational Taxonomic Units (OTUs) (HernándezRodríguez et al., 2013; Oria-de-Rueda et al., 2010). Taxa were grouped into the following categories: saprotrophic/mycorrhizal, and edible/unedible, regarding as edible taxa those with a culinary interest or some market value. 2.4. Data analysis Richess (number of taxa) and fresh weight production were analyzed for both nutritional mode (mycorrhizal/saprotrophic) and edibility at each area. Data were subjected to statistical analysis using PROC MIXED in the computer software SAS/STAT (SAS Institute Inc., 2000–2004). We used the following model:

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92

yi ¼ f ðx1 ; x2 ; . . . ; xn Þ þ ei where yi is the dependent variable (fresh weight production and richness); x1, . . ., xn are fixed categorical predictors (time after fire, edibility and nutritional mode), and ei is the residual, accounting for heterogeneous variances across the different edibility and nutritional mode levels. Means were compared by LSD Fisher tests (P < 0.05). Fresh weight data were log-transformed in order to meet normality assumptions. In order to test for differences in fungal species composition and abundance amongst areas differing in successional stage, we performed a detrended correspondence analysis (DCA) (Ter Braak and Prentice, 1988), carried out with the software CANOCO for Windows version 4.5 (Ter Braak and Šmilauer, 2002), with dry weight as the response variable. Analysis was performed on the full data set (91 species and 9 plots). Model significance was tested by a Monte Carlo permutation test. 3. Results 3.1. General data At the end of the study season, a total of 1924 sporocarps were collected and classified into a total of 91 taxa. As in preceding works (Hernández-Rodríguez et al., 2013; Oria-de-Rueda et al., 2010) there were generic level taxa in which complete identification down to the species level was not possible. This was the case for Clitocybe sp., Collybia sp., Cortinarius sp., Galerina sp., Hebeloma sp., Inocybe sp., Lycoperdum sp., Mycena sp. and Tyromyces sp. (Table 1). The number of taxa, i.e. richness, per area increased following the vegetation successional stage. The lowest richness was recorded at the 1 year-old burned areas, showing significant differences respect to the 5-year-old burned areas (P = 0.002) and the unburned areas (P < 0.001). Richness was also significantly different between 5-year-old burned areas and unburned areas (P = 0.010) (Table 2). Mycorrhizal species richness was also significantly affected by wildfire. Lower values were observed in the 1-year-old burned areas, compared to the unburned ones (P = 0.002). But then five years after fire there were five times more mycorrhizal taxa (P = 0.003) than in 1-year-old areas, while no differences were found when compared to unburned areas. Similarly, saprotrophic taxa were also negatively affected by fire compared to the unburned areas (P < 0.001) but, in this case differences between five years after fire and unburned areas were also observed (P = 0.006) (Table 2). 3.2. Sporocarp average fresh weight production according to nutritional mode and edibility Unburned areas yielded an average fresh weight production of 134.14 kg ha1, 109.21 kg ha1 were picked from 5-year-old areas, and 36.36 kg ha1 from 1-year-old areas. Fresh weight production was significantly lower (P = 0.020) in 1-year-old areas than in 5-year old areas and also than in unburned areas (P = 0.006). Differences between unburned areas (UB) and areas burned five years ago (B02) were non-significant (P = 0.300). When considering separately each of the two different nutritional models, an increase of mycorrhizal production along time after fire was evident (Fig 1). A nearly-zero production was registered in 1-year-old burned areas (1.1 kg ha1), showing significant differences with 5-year-old burned areas (44.1 kg ha1) (P = 0.040) and unburned areas (89.5 kg ha1) (P = 0.002). Production in 5-year-old areas also differed significantly from that obtained in the unburned areas (P = 0.035). Saprotrophic taxa showed a clearly

87

different pattern. Average fresh weight collected in 1-year-old areas, 5-year-old areas and unburned areas were 35.3 kg ha1, 65.1 kg ha1 and 44.7 kg ha1 respectively. Significant differences among areas were not found. Edible species average fresh weight production was highest five years after fire, with a production of 77.5 kg ha1, although not significantly different from that of unburned areas 52.7 kg ha1. The lowest production was found in 1-year-old areas, decreasing to 1.6 kg ha1 and significantly different from unburned (P = 0.020) and 5-year-old burned areas (P = 0.004). 3.3. Taxa composition following fire and according to dominating hosts Regarding multi-species associations, different fungal species compositions were observed. According to our detrended correspondence analysis, a large difference of the first eigenvalue (0.922) respect to the following (0.312, 0.067, 0.018), was observed. This indicates that greater variability among areas, in terms of fungal species composition, is explained by the gradient associated with axis 1 (Fig. 2). However, our results should be regarded with some caution given the low occurrence of many fungal species (Gauch, 1982). Fungal community associations were analyzed according to different criteria. According to the successional stage, our results show a clustering of specific taxa around the three different areas (UB, B02, B06) (Fig 2). Interestingly, pyrophytic taxa, as Pholiota carbonaria or Tyromices sp. were only present in the 1-year-old-burned areas, while taxa belonging to mature stages, such as Russula albonigra and Tricholoma saponaceum, appeared in the unburned areas exclusively. Moreover, multistage taxa, as Inocybe sp., Lactarius chrysorrheus and Lycoperdon pyriforme, were next to both B02 and UB areas, meaning that those taxa occurred in both areas indifferently (Fig. 2). Attending to plant hosts species, some species forming exclusive symbiosis with either C. laurifolius, P. nigra or Q. pyrenaica were also observed (see discussion below). 4. Discussion 4.1. General data A total of 91 different fungal taxa were collected in a single autumn season, and across all areas differing in time after fire. Other studies also reported fungal richness in Pinus spp. stands and from a single autumn season: 49 taxa in a 50-year old Pinus pinaster reforestation (Oria-de-Rueda et al., 2010) and 39 taxa in a Mediterranean ecosystem dominated by Pinus pinaster (MartínPinto et al., 2006). Comparatively, our results show a higher fungal richness, likely related to the interaction between successional stage after fire and vegetation composition. An argument in favor of this view is that species-specific taxa associated to each Pinus, Quercus and Cistus where observed. Another similar study, recorded an even higher richness, 115, in a forest dominated by Pinus pinaster with few disperse C. ladanifer individuals (Gassibe et al., 2011). In this case, the study spanned three autumn seasons, having this a likely positive effect on fungal richness. Indeed, a study including observations along several seasons in areas affected by fire at different times would be highly desirable. This would provide solid data for describing how the interaction between successional stage and year-to-year meteorological variation affects different fungal taxa. Therefore differences among study areas in our work may be interpreted as mainly driven by time after fire and its associated effects, but we cannot rule out among-plot effects. Wildfires typically have both short- and long-term effects on fungal communities (Kurth et al., 2013). As a short-term effect, fire

88

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92

Table 1 Total taxa list for nutrional mode (Mode) and edibility (Edible) collected from three study areas (B06, B02 and UB) in Northern Spain. Taxa

Code

Agaricus arvensis Schaeff. Amanita citrina (Schaeff.) Pers. Amanita phalloides (Vaill. ex Fr.) Link Amanita rubescens Pers. Armillaria bulbosa (Barla) Kile & Watling Clitocybe cerussata (Fr.) P. Kumm. Clitocybe costata Kühner & Romagn. Clitocybe decembris Singer Clitocybe fragrans (With.) P. Kumm Clitocybe gibba (Pers.) P. Kumm. Clitocybe nebularis (Batsch) P. Kumm Clitocybe sp. (Fr.) Staude Collybia butyracea (Bull.) P. Kumm. Collybia cookei (Bres.) J.D. Arnold Collybia dryophila (Bull.) P. Kumm. Collybia erythropus (Pers.) P. Kumm. Collybia maculata (Alb. & Schwein.) P. Kumm. Collybia sp. (Fr.: Fr.) Staude Cortinarius sp Fr. Crepidotus variabilis. (Pers.) P. Kumm Crinipellis scabellum (Alb. & Schw.: Fr.) Murrill Cystoderma amianthinum (Scop.) Fayod Cystoderma cinnabarinum (Alb. & Schwein.) Fayod Galerina marginata (Batsch) Kühner Galerina sp.Earle Hebeloma cistophilum Maire Hebeloma crustuliniforme (Bull.) Quél. Hebeloma sinapizans (Paulet) Gillet. Hebeloma sp. Kummer Heterobasidion annosum (Fr.) Bref. Hygrophoropsis aurantiaca (Wulfen) Maire Hygrophorus arbustivus var. quercetorum Bon & C Hygrophorus chrysodon (Batsch) Fr. Hygrophorus hypothejus (Fr.) Fr. Hygrophorus gliocyclus Fr. Hypholoma fasciculare (Huds.) P. Kumm. Hypholoma sublateritium (Schaeff.) Quél. Inocybe sp. (Fr.) Laccaria laccata (Scop.) Cooke Laccaria bicolor (Maire) P.D. Orton Lactarius vellereus (Fr.) Fr. Lactarius aurantiacus (Pers.) Gray Lactarius chrysorrheus Fr. Lactarius rufus (Scop.) Fr. Lactarius sanguifluus var. violaceus (Barla) Basso Lactarius tesquorum Malençon Leccinum corsicum (Rolland) Singer Lepista inversa (Scop.) Pat. Lycoperdon foetidum Bonord. Lycoperdon molle Pers. Lycoperdon perlatum Pers. Lycoperdon pyriforme Schaeff. Lycoperdon sp. Lycoperdon umbrinum Hornem. Lyophyllum decastes (Fr.) Singer Macrolepiota affinis (Velen.) Bon Macrolepiota mastoidea (Fr.) Singer Macrolepiota procera (Scop.) Singer Marasmius androsaceus (L.) Fr. Marasmius erythropus (Pers.) Quél. Marasmius scorodonius (Fr.) Fr. Mycena epipterygia (Scop.) Gray Mycena polygramma (Bull.) Gray Mycena sp. (Pers.: Fr) S. F. Gray Mycena vulgaris (Pers.) P. Kumm. Paxillus involutus (Batsch) Fr. Pholiota carbonaria (Fr.) Singer Rhizopogon roseolus (Corda) Th. Fr. Russula albonigra (Krombh.) Fr. Russula cyanoxantha (Schaeff.) Fr. Russula chloroides (Krombh.) Bres. Russula nigricans Fr. Russula risigallina (Batsch) Sacc. Russula rubroalba (Singer) Romagn.

Agaar Amaci Amaph Amaru Armbu Clice Clico Clide Clifa Cligi Cline Clisp Colbu Colco Coldr Coler Colma Colsp Corsp Creva Crisc Cysam Cysci Galma Galsp Hebci Hebcr Hebsi Hebsp Hetan Hygau Hygar Hygch Hyghy Hyggl Hypfa Hypsu Inosp Lacla Lacbi Lacve Lacau Lacch Lacru Lacsa Lacte Lecco Lepin Lycfo Lycmo Lype Lycpy Lycsp Lycum Lyode Macaf Macma Macpr Maran Marer Marsc Mycep Mycpo Mycsp Mycvu Paxin Phoca Rhiro Rusal Ruscy Rusch Rusni Rusri Rusru

B06

B02

+ +

UB

Mode

Edible

+ +

S MY MY MY S S S S S S S S S S S S S S MY S S S S S S MY MY MY MY S S MY MY MY MY S S MY MY MY MY MY MY MY MY MY MY S S S S S S S S S S S S S S S S S S MY S MY MY MY MY MY MY MY

E UE UE E E UE E UE E E E UE E UE E UE UE UE UE UE UE UE UE UE UE UE UE UE UE UE E E E E E UE UE UE E E UE UE UE UE E UE E E UE UE E E UE E E E E E E E E UE UE UE UE UE UE E UE E UE E UE E

+ + + + + + + + + + + + +

+

+ + + + + + + + + + +

+ + + + + +

+ + +

+ + +

+ + +

+ + + + + + + + +

+ + + +

+ +

+ +

+ + + + + + + + +

+ + + + + +

+ +

+ + + + + + +

+ +

+ + +

+ + + + +

89

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92 Table 1 (continued) Taxa

Code

Russula seccion ingratae Quél. Russula virescens (Schaeff.) Fr. Russula xerampelina (Schaeff.) Fr. Stereum hirsutum (Willd.) Pers. Stropharia aeruginosa (Curtis) Quél. Suillus collinitus (Fr.) Kuntze Suillus granulatus (L.) Roussel Suillus luteus (L.) Roussel Tremella mesenterica Retz. Trichaptum abietinum (Dicks.) Ryvarden Tricholoma saponaceum (Fr.) P. Kumm. Tricholoma sulphurescens Bres. Tricholoma terreum (Schaeff.) P. Kumm. Tricholoma ustaloides Romagn. Tyromyces sp. Karst Xerocomus ferrugineus (Boud.) Bon Xerocomus subtomentosus (L.) Quél.

Rusin Rusvi Rusxe Stehi Strae Suico Suigr Suilu Treme Triab Trisa Trisu Trite Trius Tyrsp Xerfe Xersu

B06

B02

UB

+ + +

+

+ + +

+

+ +

+ + + + + +

+ + + + +

Mode

Edible

MY MY MY S MY MY MY MY S S MY MY MY MY S MY MY

UE E E UE UE E E E UE UE UE UE E UE UE E E

B06: areas burned in 2006; B02: areas burned in 2002; UB: unburned areas; MY: mycorrhizal; S: saprotrophic; E: edible; UE: unedible. Code: code used in the Detrended Correspondence Analysis with CANOCO.

Table 2 Total richness (number of taxa) collected from three study areas (B06, B02 and UB), and classified into mycorrhizal/saprotrophic and edible/unedible. Taxa groups

B06

B02

UB

Total

Mycorrhizal Saprotrophic Edible Unedible Total

5a 8a 4a 9a 13a

25b 19b 21b 23b 44b

23b 36c 25b 34c 59c

43 48 41 50

B06: areas burned in 2006; B02: areas burned in 2002; UB: unburned areas. Values with the same letter are not significantly different (LSD Fisher Test; P < 0.05).

causes a reduction in richness (Kutorga et al., 2012; Olsson and Jonsson, 2010). This was observed in our study too, where the richness recorded differed between 59 in unburned areas to 13 taxa in the first year after fire. Regarding the nutritional model, we also recorded neat differences between the ectomycorrhizal richness in unburned areas

(23 taxa) and 1-year-old burned areas (five taxa). This is also a commonly reported result (Buée et al., 2011; Durall et al., 1999; Grogan et al., 2000; Kutorga et al., 2012; Olsson and Jonsson, 2010) explained by the strong dependency of ectomycorrhizal symbionts on their hosts (Dahlberg, 2002; Torres and Honrubia, 1997), but also related to a decrease in spore concentration and infection potential after fire (Izzo et al., 2006; Vilariño and Arines, 1991). It is somewhat surprising nonetheless, that there was actually some presence of mycorrhizal taxa immediately after the fire. This can perhaps be explained by a survival strategy in the form of different resistance structures such as infected root tips, sclerotia or resistant spores (Carney and Bastias, 2007; Grogan et al., 2000). Nonetheless, sporocarps of ectomycorrhizal species were extremely rare in the one-year old plots, and could also be regarded as incidental. Saprotrophic species richness showed a positive trend associated with more mature successional stages. This could be due to an increased complexity of the ecosystem, with mature trees enhancing the content of organic matter in the forests (Egli et al.,

Fig. 1. Production of sporocarps according to nutritional mode, edibility, and total (kg ha1). M: Mycorrhizal, S: Saprotrophic, E: Edible, UE: Unedible, T: Total. The data are mean results ± standard error of the mean. Independent comparisons were carried out within functional groups and edibility. Values with the same letter are not significantly different (LSD Fisher Test; P < 0.05).

90

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92

Fig. 2. Detrended Correspondence Analysis for unburned areas (UB), areas burned in 2002 (B02) and areas burned in 2006 (B06). Fungal taxa are identified by the code shown in Table 1.

2010). For example, when P. nigra matures, canopy is gradually closed. As a consequence, C. laurifolius individuals die because they cannot receive enough light to survive. Thus, higher amount of accumulated organic matter is observed in the forest soil and such conditions enhance the fruiting of saprotrophic fungi (Straatsma et al., 2001). 4.2. Sporocarp average fresh weight production according to nutritional mode and edibility The average fresh sporocarp production across plots in the unburned area of our study was similar to Shubin (1988), higher Bonet et al. (2010) or lower Gassibe et al. (2011), Martín-Pinto et al. (2006) than those reported in similar studies. Sporocarp productions are very irregular over the different pine stands, depending on different factors such as, host tree species (Ishida et al., 2007), age of the stand (Fernández-Toirán et al., 2006), aspect, microclimatic conditions (Hernández-Rodríguez et al., 2013) or soil (Kennedy and Peay, 2007; Nordén et al., 2004). Thus, a meta-analytical approach, rather than single studies would be most useful in shedding light into the multiple factors upon which sporocarp production depends. Not least because many studies, including ours, focus on rather local scales. As such, our results should be regarded as unavoidably influenced by local conditions, but probably valuable nonetheless as a contribution to standing and yet to come knowledge. Fungal production seemed to quickly recover five years after fire, existing no significant differences with unburned areas. According to functional groups, mycorrhizal but not saprotrophic productions increased with time after fire. This result seems to contrast with those obtained for richness, where saprotrophic richness did significantly differ among areas. An explanation for this apparent contradiction lies in the fact that some saprotrophic species were particularly prolific like P. carbonaria and Hypholoma fasciculare. Fresh weight production of desirable edible taxa, i.e. those taxa with some market value, was 131.4 kg ha1; as a reference, our results lie below those obtained by Oria-de-Rueda et al. (2010), with 294.8 kg ha1 in a Pinus pinaster forest, but are similar to those reported by Shubin (1988) who collected 153 kg ha1 in a Pinus sylvestis stand. It is noteworthy that production of edible taxa 5 years after fire, was similar to that of unburned areas, suggesting a fast after-fire recovery.

4.3. Taxa composition following fire and according to dominating hosts Disturbances caused by forest fires modify competitive interactions among different species of fungi and involve the colonization of new host species (Buscardo et al., 2009; Grogan et al., 2000). In our study, 1-year-old areas correspond to the beginning of a new fungal succession associated with the development of new vegetation. During early post-fire successional stages, xerotolerant and heat-stimulated fungi are favored relative to other species (McMullan-Fisher et al., 2002; Robinson et al., 2008; Rutigliano et al., 2013). Such species can be found up to 4–6 years following the fire event. For example, P. carbonaria, which is considered a pyrophilic species (Hernández-Rodríguez et al., 2013), was only found in 1-year-old burned areas, where it accounted for a significant proportion of all fungal biomass. Our results obtained in the 5-year-old-burned areas agree with the observations of previous studies (Fernández-Toirán et al., 2006; Kutorga et al., 2012), where the pioneer species disappeared, indicating the end of initial stage of post-fire fungal succession. Forty out of the 59 taxa recorded in unburned areas were only found there. Within this group, species belonging to the genera Cystoderma and Lepista have been cited as late stage species (Gassibe et al., 2011). We also found species which appeared in all the stages before and after the wildfire. This was the case of Hypholoma fasciculare, Laccaria laccata and Stereum hirsutum. Those can be thus considered as multi-stage species, arriving in early stages of succession and persisting also in mature stands (Visser, 1995). Besides stand age effects, the change in species composition during fungal succession from unburned to burned areas was also likely due to the vegetation present in each area. Attending to the plant hosts for ectomycorrhizal species, we detected some species forming exclusive symbiosis with either C. laurifolius, P. nigra or Q. pyrenaica. We found several fungal taxa exclusively associated with Cistus, as Hebeloma cistophilum, Lactarius tesquorum, Leccinum corsicum (Comandini et al. (2006). All these species were absent from the unburned areas, where only very few C. laurifolius individuals remained. We also recoded other taxa specifically associated with either Pinus (Hygrophorus hypothejus, Suillus granulatus and S. luteus) or Quercus (Amanita phalloides, Hygrophorus arbustivus and Russula virescens). In the unburned areas where Pinus was largely dominant, only taxa exclusively associated with Pinus appeared. Quercus-specific fungi were recorded in both one and five years after fire areas, where Quercus resprouting was frequently found.

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92

Finally, we found some mycorrhizal species able to form associations with both Cistus and Pinus, such as L. laccata, Paxillus involutus, Rhizopogon roseolus, T. saponaceum, Tricholoma terreum and Xerocomus ferrugineus (Hernández-Rodríguez et al., 2013; Moreno et al., 2003). As shown in Fig. 2, the mentioned species appear on the left part of the plot, which corresponds primarily to five year-old burned areas where Cistus and Pinus genera co-occur. This highlights the key ecological role that Cistus species could play in post-fire recovery of tree stands. Indeed, pioneer shrubs may facilitate the establishment of an ectomycorrhizal network after a stand-replacing fire (Kipfer et al., 2010) and also they can act as a bridge between the pre-fire fungal communities and the emerging post-fire seedlings (Buscardo et al., 2011). A key observation in our study was the relatively high richness, as well as edible sporocarp productions recorded shortly after reforestation. This highlights the interesting ecological role that forests where several host species co-occur, could play from their early growth stages, where just after fire a prompt resprouting or germination of pioneer species occurs. During afforestation of burned areas, forest managers may consider leaving some of this pioneer existing vegetation undisturbed. This variety of hosts in the early moments of the new forest establishment could contribute to a more complex and resilient ecosystem, not only in terms of fungal diversity, but also when it comes to the indirect effect that this might have on the ecosystem. Acknowledgements We would like to express our gratitude to Raúl Fraile for the taxa classification, María Hernández Rodríguez (PhD Student, Universidad de Valladolid) and Associate Professor Pando (Departamento de Estadística e Investigación Operativa, Universidad de Valladolid) for the statistical support and Luis Santos-del-Blanco (University of Lausanne) for reviewing and improving the manuscript. References Acácio, V., Holmgren, M., 2014. Pathways for resilience in Mediterranean cork oak land use systems. Ann. For. Sci. 71, 5–13. Andrés-Rodríguez, J., Llamas-Frade, B., Terrón-Alfonso, A., Sánchez-Rodriguez, J.A., García-Prieto, O., Arrojo-Martín, E., Pérez-Jarauta, T., 1999. Guía de los hongos de la Península Ibérica. Celaryn, León. Antonín, V., Noordeloos, M.E., 2010. A Monograph of Marasmioid and Collybioid Fungi in Europe. Eching, Germany, IHW-Verlag, pp. 480. Bastias, B.A., Huang, Z.Q., Blumfield, T., Xu, Z., Cairney, J.W.G., 2006. Influence of repeated prescribed burning on the soil fungal community in an eastern Australian wet sclerophyll forest. Soil Biol. Biochem. 38, 3492–3501. Bonet, J.A., Palahí, M., Colinas, C., Pukkala, T., Fischer, C.R., Miina, J., Martínez de Aragón, J., 2010. Modelling the production and species richness of wild mushrooms in pine forests of the Central Pyrenees in northeastern Spain. Can. J. For. Res. 40, 347–356. Bradshaw, S.D., Dixon, K.W., Hopper, S.D., Lambers, H., Turner, S.R., 2011. Little evidence for fire-adapted plant traits in Mediterranean climate regions. Trends Plant Sci. 16, 69–76. Bruns, T.D., Kretzer, A.M., Thomas, R., Stendell, E.A., Bidartondo, M.I., Szaro, T.M., 2002. Current investigations of fungal ectomycorrhizal communities in the sierra national forest. USDA For. Serv. Gen. Tech. Rep. 183, 83–90. Buée, M., Maurice, J.-P., Zeller, B., Andrianarisoa, S., Ranger, J., Courtecuisse, R., Marçais, B., Le Tacon, F., 2011. Influence of tree species on richness and diversity of epigeous fungal communities in a French temperate forest stand. Fungal Ecol. 4, 22–31. Buscardo, E., Freitas, H., Pereira, J.S., de Angelis, P., 2011. Common environmental factors explain both ectomycorrhizal species diversity and pine regeneration variability in a post-fire Mediterranean forest. Mycorrhiza 21, 549–558. Buscardo, E., Rodríguez-Echeverría, S., de Angelis, P., Freitas, H., 2009. Comunidades de hongos ectomicorrícicos en ambientes propensos al fuego: compañeros esenciales para el reestablecimiento de pinares mediterráneos. Ecosistemas 18, 55–63. Buscardo, E., Rodríguez-Echeverría, S., Barrico, L., García, M.A., Freitas, H., Martín, M.P., De Angelis, P., Muller, L.A.H., 2012. Is the potential for the formation of common mycorrhizal networks influenced by fire frequency? Soil Biol. Biochem. 46, 136–144.

91

Carney, J.W.G., Bastias, B.A., 2007. Influences of fire on forest soil fungal communities. Can. J. For. Res. 37, 207–215. Cerro-barja, A., Navarro, R., Andrés-Abellán, M., Lucas-Borja, M.E., 2006. Factores que influyen en la difícil regeneración de los montes de Pinus nigra Arn., en la Serranía de Cuenca. Montes 84, 33–39. Claridge, A., Trappe, J., Hansen, K., 2009. Do fungi have a role as soil stabilizers and remediators after forest fire? For. Ecol. Manage. 257, 1063–1069. Comandini, O., Contu, M., Rinaldi, A.C., 2006. An overview of Cistus ectomycorrhizal fungi. Mycorrhiza 16, 381–395. Dahlberg, A., 1991. Ectomycorrhiza in Coniferous Forest-Structure and Dynamics of Populations and Communities. Swedish University of Agricultural Sciences, Uppsala, pp. 33. Dahlberg, A., 2002. Effects of fire on ectomycorrhizal fungi in fennoscandian boreal forests. Silva Fenn. 36, 69–80. Durall, D.M., Jones, M.D., Wright, E.F., Kroeger, P., Coates, K.D., 1999. Species richness of ectomycorrhizal fungi in cutblocks of different sizes in the Interior Cedar-Hemlock forests of northwestern British Columbia: sporocarps and ectomycorrhizae. Can. J. For. Res. 29, 1322–1332. Egli, S., Ayer, F., Peter, M., Eilmann, B., Rigling, A., 2010. Is forest mushroom productivity driven by tree growth? Results from a thinning experiment. Ann. For. Sci. 67, 509–509. Fernández-Toirán, L.M., Ágreda, T., Olano, J.M., 2006. Stand age and sampling year effect on the fungal fruit body community in Pinus pinaster forest in central Spain. Can. J. Bot. 84, 1249–1258. Gassibe, P., Fraile-Fabero, R., Hernández-Rodríguez, M., Oria-de-Rueda, J.A., MartínPinto, P., 2011. Fungal community succession following wildfire in a Mediterranean vegetation type dominated by Pinus pinaster in Northwest Spain. For. Ecol. Manage. 262, 655–662. Gauch, H.G., 1982. Multivariate Analysis in Community Ecology. Cambridge University Press, Cambridge, England, pp. 158–159. Grogan, P., Baar, J., Bruns, T.D., 2000. Below-ground ectomycorrhizal community structure in a recently burned bishop pine forest. J. Ecol. 88, 1051–1062. Hart, S.C., DeLuca, T.H., Newman, G.S., MacKenzie, M.D., Boyle, S.I., 2005. Post-fire vegetative dynamics as drivers of microbial community structure and function in forest soils. For. Ecol. Manage. 220, 166–184. Hernández-Rodríguez, M., Oria-de-Rueda, J.A., Martín-Pinto, P., 2013. Post-fire fungal succession in a Mediterranean ecosystem dominated by Cistus ladanifer L. For. Ecol. Manage. 289, 48–57. Ishida, T.A., Nara, K., Hogetsu, T., 2007. Host effects on ectomycorrhizal fungal communities: insight from eight host species in mixed conifer-broadleaf forests. New Phytol. 174, 430–440. Izzo, A., Canright, M., Bruns, T.D., 2006. The effects of heat treatments on ectomycorrhizal resistant propagules and their ability to colonize bioassay seedlings. Mycol. Res. 110, 196–202. Jiménez Esquilín, A.E., Stromberger, M.E., Massman, W.J., Frank, J.M., Shepperd, W.D., 2007. Microbial community structure and activity in a Colorado Rocky Mountain forest soil scarred by slash pile burning. Soil Biol. Biochem. 39, 1111– 1120. Jonsson, L., Dahlberg, A., Nilsson, M., Zackrisson, O., Karén, O., 1999. Ectomycorrhizal fungal communities in late-successional Swedish boreal forests, and their composition following wildfire. Mol. Ecol. 8, 205–215. Keeley, J.E., Pausas, J.G., Rundel, P.W., Bond, W.J., Bradstock, R.A., 2011. Fire as an evolutionary pressure shaping plant traits. Trends Plant Sci. 16, 406–411. Kennedy, P.G., Peay, K.G., 2007. Different soil moisture conditions change the outcome of the ectomycorrhizal symbiosis between Rhizopogon species and Pinus muricata. Plant Soil 291, 155–165. Kipfer, T., Egli, S., Ghazoul, J., Moser, B., Wohlgemuth, T., 2010. Susceptibility of ectomycorrhizal fungi to soil heating. Fungal Biol. 114, 467–472. Knudsen, H., Vesterholt, J., 2008. Funga Nordica: Agaricoid, Boletoid and Cyphelloid Genera, vol. 1. Nordsvamp, Copenhagen. Kurth, V.J., Fransioli, N., Fulé, P.Z., Hart, S.C., Gehring, C.A., 2013. Stand-replacing wildfires alter the community structure of wood-inhabiting fungi in southwestern ponderosa pine forests of the USA. Fungal Ecol. 6, 192–204. Kutorga, E., Adamonyte˙, G., Irše˙naite˙, R., Juze˙nas, S., Kasparavicˇius, J., Markovskaja,  naite˙, J., Treigiene˙, A., 2012. Wildfire and post-fire management S., Motieju effects on early fungal succession in Pinus mugo plantations, located in Curonian Spit (Lithuania). Geoderma 191, 70–79. Longo, M.S., Urcelay, C., Nouhra, E., 2011. Long term effects of fire on ectomycorrhizas and soil properties in Nothofagus pumilio forests in Argentina. For. Ecol. Manage. 262, 348–354. López, B.C., Gracia, C.A., Sabaté, S., Keenan, T., 2009. Assessing the resilience of Mediterranean holm oaks to disturbances using selective thinning. Acta Oecol. 35, 849–854. Luoma, D.L., Eberhart, J.L., Molina, R., Amaranthus, M.P., 2004. Response of ectomycorrhizal fungus sporocarp production to varying levels and patterns of green-tree retention. For. Ecol. Manage. 202, 337–354. Luoma, D.L., Frenkel, R.E., Trappe, J.M., 1991. Fruiting of hypogeous fungi in Oregon Douglas-Fir forests: seasonal and habitat variation. Mycologia 83, 335–353. Martín-Pinto, P., Vaquerizo, H., Peñalver, F., Olaizola, J., Oria-de-Rueda, J.A., 2006. Early effects of a wildfire on the diversity and production of fungal communities in Mediterranean vegetation types dominated by Cistus ladanifer and Pinus pinaster in Spain. For. Ecol. Manage. 225, 296–305. McMullan-Fisher, S.J.M., May, T.W., Keane, P.J., 2002. The macrofungal community and fire in a Mountain Ash forest in southern Australia. In: Hyde, K.D., Jones, E.B.G. (Eds.), Fungal Succession. Fungal Diversity Press, Hong Kong, pp. 57–76.

92

O. Mediavilla et al. / Forest Ecology and Management 331 (2014) 85–92

Moreno, G., Díez, J., Manjón, J.L., 2003. Hongos hipógeos micorrizógenos con interés comercial recolectados en matorrales y praderas mediterráneas. Etnomicología, valor ecológico y diversidad, in: Proceedings of the I Congreso Nacional de Micología Forestal Aplicada. Soria, Spain. Moreno, J.M., Oechel, W.C., 1994. The Role of Fire in Mediterranean-type Ecosystems. Global Change in Mediterranean-type Ecosystems. SpringerVerlag, New York, USA. Nordén, B., Ryberg, M., Götmark, F., Olausson, B., 2004. Relative importance of coarse and fine woody debris for the diversity of wood-inhabiting fungi in temperate broadleaf forests. Biol. Conserv. 117, 1–10. Ohenoja, E., 1984. Fruit body production of larger fungi in Finland. 1. Introduction to the study in 1976–1978. Ann. Bot. Fenn. 21, 349–355. Ohenoja, E., 1989. Forest fertilization and fruiting body production in fungi. Atti del Cent. Sudi per la Flora Mediterr. 7, 233–253. Olsson, J., Jonsson, B.G., 2010. Restoration fire and wood-inhabiting fungi in a Swedish Pinus sylvestris forest. For. Ecol. Manage. 259, 1971–1980. Oria-de-Rueda, J.A., Hernández-Rodríguez, M., Martín-Pinto, P., Pando, V., Olaizola, J., 2010. Could artificial reforestations provide as much production and diversity of fungal species as natural forest stands in marginal Mediterranean areas? For. Ecol. Manage. 260, 171–180. Pausas, J.G., Llovet, J., Rodrigo, A., Vallejo, R., 2008. Are wildfires a disaster in the Mediterranean basin? – a review. Int. J. Wildl. Fire 17, 713. Robinson, R.M., Mellican, A.E., Smith, R.H., 2008. Epigeous macrofungal succession in the first five years following a wildfire in karri (Eucalyptus diversicolor) regrowth forest in Western Australia. Austral Ecol. 33, 807–820. Rutigliano, F.A., Migliorini, M., Maggi, O., D’Ascoli, R., Fanciulli, P.P., Persiani, A.M., 2013. Dynamics of fungi and fungivorous microarthropods in a Mediterranean maquis soil affected by experimental fire. Eur. J. Soil Biol. 56, 33–43.

Shubin, V.I., 1988. Influence of fertilizers on the fruiting of forest mushrooms. In: Vänninen, I., Raatikainen, M. (Eds.), Proceedings of the Finnish-Soviet Symposium on Timber Resources in Jyväskylä, Finland. Acta Bot. Fennica 136, 85–87. Smith, J.E., Molina, R., Huso, M.M.P., Luoma, D.L., Mckay, D., Castellano, M.A., Lebel, T., Valachovic, Y., 2002. Species richness, abundance, and composition of hypogeous and epigeous ectomycorrhizal fungal sporocarps in young, rotationage, and old-growth stands of Douglas-fir (Pseudotsuga menziesii) in the Cascade Range of Oregon, USA. Can. J. Bot. 80, 186–204. Straatsma, G., Ayer, F., Egli, S., 2001. Species richness, abundance, and phenology of fungal fruit bodies over 21 years in a Swiss forest plot. Mycol. Res. 105, 515– 523. Ter Braak, C.J.F., Prentice, I.C., 1988. A theory of gradient analysis. Adv. Ecol. Res. 18, 272–313. Ter Braak, C.J.F., Šmilauer, P., 2002. CANOCO Reference Manual and CanoDraw for Windows User’s Guide: Software for Canonical Community Ordination (version 4.5). Microcomputer power, Ithaca, New York, USA. Torres, P., Honrubia, M., 1997. Changes and effects of a natural fire on ectomycorrhizal inoculum potential of soil in a Pinus halepensis forest. For. Ecol. Manage. 96, 189–196. Vilariño, A., Arines, J., 1991. Numbers and viability vesicular-arbuscular fungal propagules field soil samples after wildfire. Soil Biol. Biochem. 23, 1083–1087. Visser, S., 1995. Ectomycorrhizal fungal succession in jack pine stands following wildfire. New Phytol. 129, 389–401.