Evaluation of Density and Viability of Arbuscular Mycorrhizal Spores in Austrocedrus chilensis Forests Affected by Wildland Fires in Patagonia, Argentina

Simple Summary

Wildfires represent a recurrent disturbance in the Andean Patagonia region, with increasing frequency in recent decades. This study analyzed how fires could affect the density of arbuscular mycorrhizal spores in forests of Austrocedrus chilensis, an endemic arbuscular mycorrhizal conifer vulnerable to fire. Our study area was settled in a widespread area affected by wildfire, with varying severities, in February 2015. Soil samples were collected from three sites with different precipitation ranges and three different fire severities each, evaluated 10 months and 5 years after the fire. Our key findings were as follows: (a) post-fire, spores were significantly more abundant in sites affected by moderate severity fires than in unburned sites or those with high fire severity; (b) after 5 years, burned sites showed no differences between severities, but the driest sites showed no changes, since fires occurred with the same pattern and spore density; (c) seedlings from the bioassay showed less than 25% mycorrhizal colonization in burned sites; (d) wildfires reduce arbuscular mycorrhizal spores and alter soil properties; (e) in drier sites, natural recovery is limited, requiring active restoration. Post-fire management must consider that wildfires reduce mycorrhizal presence and alter soil properties.

Abstract

Background: Wildfires represent a recurrent disturbance in the Patagonia Andean region, with increasing frequency in recent decades. Austrocedrus chilensis is an arbuscular mycorrhizal (AM) endemic conifer that is particularly vulnerable to fire, a situation that will worsen with climate change. In February 2015, a wildfire affected 5700 ha of Austrocedrus chilensis forests with varying severities (WFSs). The aim of this study was to determine and compare the density of AM spores (AMSs) in soil affected by different WFS and non-affected sites, considering site features. Methods: Ninety soil samples were collected from three sites 10 months and 5 years after the fire. The AMSs density was determined, a bioassay was set, and soil physicochemical features were evaluated. Results: After the wildfires, spores were significantly more abundant in sites affected by moderate severity fires. After 5 years, burned sites showed no differences between severities, but the driest sites showed no changes since the wildfires occurred. Seedlings from the bioassay showed less than 25% mycorrhizal colonization growing in soil from burned sites, regardless of fire severity, compared with unburned soils. Conclusions: For restoration strategies, it must be considered that wildfires reduce mycorrhizal spores and mycelium and alter soil properties, indicating that both WFSs require active restoration; under drier conditions, the spore banks do not change after a 5-year period.

1. Introduction

Wildfires have been a recurrent disturbance in the Patagonia Andean region, causing significant environmental, social, and economic impacts similar to those in other countries around the world [1,2]. In these forests, the natural fire regime varies according to the precipitation regime. In general, fire frequency increases from the humid sites in the west toward xeric sites in the east [3]. In the past, fire frequency increased in the majority of Austrocedrus chilensis (D. Don) Pic. Sern et Bizarri stands after 1850, coincidentally with an increase in the number of indigenous people that inhabited the region and the beginning of European settlement. This frequency peaked at the end of the 19th century and started to decline from then on because the indigenous population decreased, and the creation of several National Parks in the region initiated the policy of fire suppression [1], while during the last three decades, the number of wildfire ignitions has risen in the whole central Patagonia Andean region [4,5]. As rains are markedly limited to autumn–winter in the region, summer provides favorable conditions for wildfire occurrence, combining the greatest water deficit and the highest temperatures [6]. Moreover, in a climate change scenario, with predictions of a 1 °C to 3 °C temperature increase and of a 10% to 30% precipitation decrease for West Patagonia [7,8], wildfire frequency is expected to rise [9,10]

Austrocedrus chilensis is an endemic and emblematic conifer of northern Patagonian forests, which are among the areas most affected by wildfires [11]. It grows in pure or in mixed stands interspersed with Nothofagus dombeyi Mirb. Oerst. and other tree species at a latitude between 37°07′ and 43°44′ S [12]. Austrocedrus chilensis is an obligate seeder [13,14] that does not resprout after fires [15]. It is highly vulnerable to fire given its thin bark and persistent, dry branches, which favor fire access to the crown [16], as well as its high resin foliage contents, which burn explosively [17]. It forms extensive stands, which are valuable for forestry and tourism, and it is of significant ecological importance, covering an area of 100.000 ha [18]. However, fire and diseases [19,20,21], invasions by other plant species [22], and herbivory on saplings after fires [23] have generated severe degradation, resulting in great economic and environmental losses [24]. It has been declared as a ‘near threatened’ species by The IUCN (International Union for the Conservation of Nature) since 2013 [25].

Regarding its mycorrhizal status, A. chilensis is an obligated AM species [26,27]. This mutualism is of great ecological importance, involved in seedling establishment and tree growth, with major roles in absorption, transport, and translocation of nutrients (P, N, C, K, and Ca) [28], increased water availability and tolerance to drought [29,30], and resistance to pathogen attack [31]. On the other hand, AM Fungi (AMF) improves soil physicochemical properties [32] and plays a role in nutrient cycling [33], serving as key elements in ecosystem health and supporting vegetation recovery following degradation [34]. However, AMF are highly susceptible to fire [35,36,37], especially mycelial structures and colonized AM root tips. Nevertheless, AM spores (resistance structures) have been reported as not being drastically affected by fire [38,39,40]; fire heat could even break spore dormancy [41]. In any event, fire can reduce AMF diversity compared with unburned areas [33,42,43], although if edaphic, weather, and vegetation conditions are favorable, AM communities could recover in a few decades [44,45], contributing to quick ecosystem recovery after fire [46,47]. Although mycorrhizas are crucial for maintaining biodiversity and ecosystem function, there is surprisingly limited knowledge regarding the landscape-scale biogeographical patterns of AMF species and the environmental factors that affect their distributions, including their response to fire. Empirical work has linked certain edaphic properties (e.g., soil texture, pH, nitrogen, phosphorus) to AMF distributions [48,49,50]. Dispersal capabilities of AMF are likely species-specific and environmentally dependent [51], although the mechanisms are poorly understood and have not been addressed for A. chilensis forests. Biotrophic fungi (mutualists and antagonists alike) are not only affected directly by fire but also by the responses of their host [52]. Fire negatively influences overall mycorrhizal colonization through the fire-induced mortality of mycorrhizal inocula, drastically reduces active mycelium, induces changes to soil physicochemical features, and can shift plant species composition from mycorrhizal to non-mycorrhizal [46]. Additionally, host plant mortality eliminates the energy source, in the form of plant-derived photosynthates, for most mycorrhizal fungi [53].

During February and March of 2015, near Cholila (Chubut, Argentina), an extreme behavior wildfire, the most severe in the last 100 years, occurred. From the total 27,101 ha burned, approximately 5700 ha corresponded to pure and mixed A. chilensis forests [5]. Therefore, it constituted a suitable scenario to assess the AM spore bank’s postfire condition, as well as to understand its evolution over time, to properly manage restoration actions. In this sense, the aims of this study were to (1) determine and compare over time the abundance of AMSs in A. chilensis forest areas affected by different WFS, in relation to non-affected patches, as well as considering the sites’ features, and (2) analyze the behavior of the AM spore bank through its infectivity on A. chilensis seedlings.

2. Materials and Methods

2.1. Study Area

This study was carried out in three pure stands of A. chilensis forest affected by 2015 wildfire, called “Las Horquetas” fire, in Villegas and Tigre Valley (A), Cholila lake center (B), and Cholila lake east (C) in Chubut province (Figure 1).

Three different fire severities were selected as treatments in each location. The term fire severity was born out of the need to provide a description of how fire intensity affects ecosystems [54], and is largely dependent upon the nature of the fuel available for burning, and the combustion characteristics that occur when these fuels are burned. The preliminary selection of the sampling sites was based on the fire severity classification map elaborated for Las Horquetas fire [5]. The map was made based on a severity index whose original values were obtained in the geomatic laboratory and later calibrated with field surveys in the area of the fire. To estimate the index, the authors worked with Landsat 8 OLI (USGS) satellite data with a capture date of 4/11/15, which were previously atmospherically corrected using the ATCOR 2/3 (Wessling, Germany) algorithm, and the postfire Normalized Burn Ratio. Then, during the first growing season after the fire, the fire damage was recorded in plots with radii of 15 m in different strata: litter and duff, herbaceous, shrubs and tree regeneration, shrubs and low trees, and dominant trees [55]. To record fire severity, the scale adapted from [56] was used. For this study, consider class 0 of [56] as “unburned”, classes 2 and 3 as “moderately affected”, and classes 4 and 5 as “severely affected” (Figure 2). The slopes ranged between 10% and 35% in all sites, while the aspects were mostly southeast, except the unburned site A, the severely affected site B (northeast), and the unburned site B (southwest). The three sampling sites had different annual precipitations: Site A ≅ 2000 mm; Site B ≅ 1500 mm; and Site C ≅ 800 mm [57].

2.2. Sampling Design

During January and February 2016, a sampling plot of ≅ 100 m² was set and sampled for each treatment/site, and revisited and sampled in February 2021. In each plot five soil subsamples of approximately 500 g, from the first 15 cm of soil, were collected and placed in new plastic bags and stored in a refrigerator at 4 °C until the AMS evaluation. A total of 90 soil samples were obtained (2016 and 2021), which were processed in duplicate (180 subsamples in total).

In addition, three composite soil samples (630 cm³ each) of the upper 5 cm of mineral soil were collected at each sampling site in 2016 (27 soil samples) to determine physicochemical parameters.

2.3. Sample Processing: Spore and Soil Analyses

For the extraction of soil AMS, the wet sieving and sucrose gradient method (adapted from [58]) was used. Soil subsamples of 50 to 75 g per treatment/site were placed in a 20% sodium pyrophosphate solution for 60 min and sieved with distilled water using a 25 μ sieve; the obtained fraction was centrifuged at 2000 rpm for 5 min to precipitate the AMS. After discarding the supernatant, 50 ml of 50% sucrose solution was added, homogenized, and centrifuged at 2000 rpm for 5 min. The supernatant was filtered on filter paper (103 slow, XinXing®, Hangzhou, China), placed in a Petri dish, and counted under stereomicroscope. Density of AMSs was expressed as the amount of AMSs per 100 g of dry soil by treatment [59].

Soil analyses were carried out at CIEFAP Soil Laboratory with dried 2 mm sieved fraction of soil. Physicochemical parameters analyzed were as follows: (1) current pH 1:2.5 soil/distilled water ratio [60]; (2) electrical conductivity (EC, ds/m) [61]; (3) organic matter (OM%) [62], with soil organic carbon (C) estimated from the ratio between organic matter and the van Bemmelen factor of 1.724 [63]; (4) total nitrogen (N, %) obtained by Kjeldahl method [64]; (5) available phosphorus (P, mg/kg) estimated by the Olsen method [65], recommended for slightly acidic soils such as Andisols [66]; (6) cation exchange capacity (CEC) [67]; and (7) exchangeable bases, to determine calcium (Ca, mEq/100 g) and magnesium (Mg, mEq/100 g), sodium (Na, mEq/100 g), and potassium (K, mEq/100 g) [68].

2.4. Bioassay Setup

A soil bioassay, to determine AMSs’ real colonization capacity, was conducted with soil sampled 10 months after fire. Austrocedrus chilensis seeds were surface-sterilized in 10% sodium hypochlorite for 10 min, sown in sterile distilled water [69], and then hydrated for 24 h in sterile distilled water before storing them at 4 °C in a refrigerator for 45 days to stratify (adapted from [70]). Non-AM seedlings were obtained in a growth chamber at 17–19 °C, 48–55% relative air humidity, and with a 16 h photoperiod with 1400 lux radiation with sterilized pumicite (autoclave 3 times at 120 °C for 30 mins; adapted from [71]) for 30 d. Three weeks later, 2 seedlings were transplanted to a clean, 250 cm³ pot filled with a 1:1 (v/v) mix of soil obtained from each sampling unit and sterilized pumicite (as mentioned above). As a control, 10 pots were filled with mixed (1:1, v/v) sterilized soil (mixing soil from each treatment from all sites) and sterilized pumicite, both autoclaved as previously mentioned. Seedlings were randomly arranged and grown for 18 months in a growing chamber, under the same set-up as described for germination. The root system of each seedling was cut into 10 mm long portions (approx. 600 mg per seedling) to fit in Tissue-Tek plastic capsules (Fisher Scientific Co., Pittsburgh, PA, USA), cleared in 10% KOH for 30 min at 100 °C under water bath and 15% H₂O₂ overnight at room temperature. Cleared samples were immersed 60 mins at 4 °C in a staining solution of 0.05% trypan blue in lactoglycerol, rinsed with tap water, and stored in lactoglycerol at 4 °C until microscopic examination [72]. Arbuscular mycorrhizal colonization percentage for each seedling root system (AM%) was estimated following [73], using the complete root system, as follows:

AM% = (number mycorrhizal intersects/total intersects) × 100

All AM structures (arbuscules, coils, vesicles and hyphae) were counted separately and considered as mycorrhizal intersects for AM%.

2.5. Statistical Analysis

Spore density and AM% were analyzed using a two-way generalized linear mixed model (2W-GLMM) with the restricted maximum likelihood estimation method (RELM) with posterior comparisons with the test DGC (test of exclusionary groups). Sites (A, B, C) were treated as blocks (incorporated as a random effect) and different treatments (HIGH, MODERATE, and UNBURNED wildfire severity) as fixed effects [74] with R-DCOM in Infostat (Windows version 2020, Córdoba, Argentina) [75]. Arbuscular mycorrhizal structures (arbuscules, coils, vesicles, and hyphae) were analyzed with a hierarchical clustering analysis using pheatmap and ggplot2 packages for Rstudio (version 2022.2.0.443, Boston, MA, USA). Soil physicochemical differences between sites and treatments were analyzed using linear mixed models (LMMs), with the site and the treatment as fixed effects and the plot as random effect (the three subsamples from the same plot were not considered as independent samples). For some variables (%OM, potassium, and sodium) a log transformation was needed. Means comparison of site and treatment were performed with Tukey Test. The analysis and the verification of the assumption of normality and homogeneity of variance were performed with the statistical software, Navure 2.3.1 (Córdoba, Argentina). To further analyze the relationships between seedlings’ mycorrhizal status and soil features, Pearson correlation tests and PCA were conducted, including AM%, AM structures, pH, OM%, electrical conductivity, and content of N, Mg, P, K, Ca, Na.

All analyses were performed at a 0.05 significance level with the statistical package InfoStat for Windows version 2020 (Córdoba, Argentina) [75] and the RStudio version 2022.2.0.443 (Boston, MA, USA) [76].

3. Results

3.1. Analyses of Soil AMS Densities and Soil Variables

In total, 24.434 spores belonging to AMF were recovered from 180 soil samples (Table S1). Ten months after the wildfire, the AM spores’ density was significantly and consistently more abundant in moderately affected sites, compared both with the unburned and the highly affected sites (Figure 2A, 2W-GLMM, F = 5.14; p = 0.0115). Also, the driest site (C) tripled the AMS density compared with the other sites for the moderately affected condition (Figure 3A).

After 5 years (2021), the burned sites showed no differences in AMS densities between wildfire severities for sites A and B (Figure 3B). Unburned sites presented higher densities from the initial situation (Table S1), and significantly higher than burned treatments, except in the driest site (C) (2W-GLMM, F = 7.39; p = 0.0022). Also, the AMS density patterns between treatments differed when considering the site (Figure 3A,B). Interestingly, the driest site (C) did not show changes in AMS density after 5 years in the burned sites.

Regarding soil physicochemical properties, nitrogen was significantly lower in the burned treatments regardless of WFS, while soil OM%, CEC, and C/N were consistently higher in the unburned treatment compared to both burned ones (

Table 1. Physicochemical properties by site and wildfire severity. Different letters indicate significant differences (p < 0.05). Letters indicate differences between treatments within each site [Tigre and Villegas valley (A), Cholila lake center (B), and Cholila lake east (C)]. Variables without letters showed no significant differences.

Sampling Sites	A			B			C
Fire Severity	Unb *	Mod	High	Unb	Mod	High	Unb	Mod	High
pH	6.05	6.71	7.15	6.85	7.31	7.18	6.76	6.80	7.63
OM	24.38	11.04	11.02	28.72	8.95	7.46	34.76	12.95	8.51
* T	b	a	a	b	a	a	b	a	a
N	0.72	0.41	0.46	0.66	0.39	0.29	0.91	0.64	0.39
* T	b	a	a	b	a	a	b	a	a
Available P	40.66	21.65	55.23	24.01	47.29	30.45	54.03	44.48	39.68
Na	0.88	0.88	1.06	1.12	0.76	0.52	1.07	0.71	0.85
K	0.96	1.37	0.94	1.71	0.81	0.55	1.69	1.21	1.68
Ca	16.91	7.08	17.42	38.16	18.42	14.83	43.33	20	32.5
Mg	4.08	2.08	5.58	9.33	1.83	1.33	7.5	5.67	4.5
EC	0.13	0.13	0.21	0.08	0.09	0.13	0.12	0.13	0.22
C/N	16.7	13.1	11.8	24.4	11.5	12.5	19.2	9.9	10.8
* T	b	a	a	b	a	a	b	a	a
CEC	42.67	37.33	31.00	52.33	26.00	23.33	44.33	34.00	31.33
* T	b	a	a	b	a	a	b	a	a

* Unb = Unburned, Mod = Moderate, OM = %, N = %, Available P = mg/kg, Na = mEq/100 g, K = mEq/100 g, Ca = mEq/100 g, Mg = mEq/100 g, EC = ds/m, C/N = %.

). The pH was marginally significant (p < 0.08), high severity treatments showed higher pH than the moderate treatments, and both were higher than the unburned treatment. On the other hand, the soil Ca content was marginally significant (p < 0.08) between sites and treatment, but without a clear pattern (Table S2).

3.2. Seedling AM Colonization from Soil Bioassay

Ninety-five A. chilensis seedlings were evaluated from the bioassay, showing significant differences in AM% between wildfire severity treatments, and both high and moderate showed significantly lower values compared to the unburned treatment (Figure 4B, Table S3). The AM% for the moderate severity treatment represents 33.8% of the unburned treatment value, while the high severity treatment represents only 12%. The control seedlings showed no AM colonization (Figure 4, Table S3), and led to non-significant differences with the high severity treatment. Consequently, seedlings from the unburned treatment showed the greatest amounts of AM nutrient and photosynthate exchange structures (arbuscules, coils), while seedlings grown in soil affected by wildfires showed mostly explorations and storage of nutrient and photosynthate structures (hyphae and vesicles), but in small amounts, especially in wet sites (A = Tigre and Villegas valley) (Figure 4A, Table S3).

3.3. Relationships Between Seedlings AM Colonization, Spore Density, and Soil Features

Arbuscular mycorrhizal colonization in A. chilensis seedlings was negatively correlated with soil pH and EC and positively with OM% and C/N (Pearson correlation coefficient,

Table 2. Correlations index and significance of Pearson correlation analysis. Highlighted in bold letters are significant correlations.

Variable	% AM		Arbuscules		Coil		Hyphae		Vesicles
	r	p	r	p	r	p	r	p	r	p
pH	−0.7	0.05	−0.7	0.03	−0.6	0.00045	−0.8	0.02	−0.6	0.03
OM%	0.8	0.02	0.9	0.003	0.92	0.11	0.8	0.01	0.72	0.09
EC	−0.7	0.05	−0.65	0.06	−0.51	0.17	−0.53	0.14	−0.7	0.05
Mg	0.35	0.36	0.62	0.07	0.7	0.03	0.52	0.15	0.24	0.53
C/N	0.76	0.02	0.8	0.01	0.95	0.0001	0.8	0.01	0.66	0.05

). Arbuscle and hyphae abundance were positively correlated with OM% and C/N and negatively correlated with soil pH (Pearson correlation coefficient, Table 2). Coils’ abundance was positively correlated with OM%, Mg content, and C/N (Pearson correlation coefficient, Table 2). Finally, the vesicles’ abundance was negatively correlated with soil electrical conductivity and positively with OM% and C/N (Pearson correlation coefficient, Table 2). On the other hand, spore density (2016) was not correlated with any soil features (Table S4).

Principal component analysis (PCA) showed that the combination of AM structures, AM%, and soil features (Figure 5) could be explained by two components, including 69.6% of the variance (Tables S5 and S6); PC1 and PC2 explained 43.1% and 26.5% of the variance, respectively. PC1 discriminates between burned treatments and unburned, while PC2 discriminates between sites (A = Tigre and Villegas valley, B = Cholila lake center, and C = Cholila lake east).

The N, K, Ca, and Mg content (EV = 0.715; −0.701; −0.818; −0.680, respectively), arbuscle, vesicle, coil, and hyphae abundance (EV = −0.732, −0.788, −0.798, −0.881), and AM% (EV = −0.816) were spatial variables that contributed to PC1, grouping variables affected by fire. On the other hand, pH (EV = −0.657) OM % (EV = 0.751) and Na content (EV = 0.736) contributed to PC2, grouping site dependent variables.

4. Discussion

Arbuscular mycorrhizal colonization and spore-based studies are a gap in knowledge about South American Austrocedrus chilensis forests, except for a few works about plant mycorrhizal status from the 20th century [26,27,77]. For this reason, this study represents the first analysis regarding bank evolution and infectivity after wildfires for these native ecosystems. Our results show that wildfire events had direct effects on the AMSs communities associated with A. chilensis forests, and that these effects were affected by and varied according to fire severity and site conditions. It has been already stated that fire reduces root AM colonization [41,78,79] but increases AMS density in the soil [80]. In our study, immediately after wildfires occurred, the AMS density was significantly more abundant in the moderately affected treatment in the dry site (C), compared both with the unburned to highly affected treatments at the same site and with all treatments from the other wetter sites. On the other hand, seedlings growing in soils affected by wildfires showed low or null AM colonization regardless of WFS. The indirect effects of wildfires on AM colonization include changes in soil physicochemical properties that may affect AMF phenology and/or AM spore formation [46]. In this study, a tendency toward increase in pH and reduction in N and C was found, which can also reduce AM colonization and/or AM sporulation, although more studies are needed for A. chilensis forests. After wildfires, the surviving soil microecosystems and their ability to recover are essential to forest succession and restoration [81,82,83]. For those microorganisms beneath the surface, soil is a good insulator, especially when it is dry, and the temperature at a 2.5 cm depth may be 50 °C when the surface is 100 °C [84]. In this sense, the results of this work showed that soon after they took place, moderate severity wildfires were associated with higher AMS densities, and that this effect was more pronounced in dryer site conditions. In these xeric sites, spore abundance does not recover even 5 years after the event. The authors of [38] found that AM communities are resilient to wildfires on decadal timescales; this resilience appears to depend on the post-fire regrowth of understory vegetation and the subsequent recovery of soil chemical properties. However, regarding the direct effects of wildfires on AM fungi, results have been contradictory, reporting either negative (e.g., [85,86,87,88,89,90]), neutral (e.g., [35,46,91,92,93]), or positive results [94].

Urretavizcaya and colleagues [95] found that natural Austrocedrus chilensis regeneration can only be expected to occur in patches that have been unaffected by fire, or in areas affected by low severity wildfires, because of low seed availability. The potential for natural regeneration of this species following large-scale disturbances that eliminate or drastically reduce the forest canopy is limited [11]. The emergence of Austrocedrus chilensis seedling patterns varied with site conditions and forest management practices [96]. Additionally, A. chilensis seedling survival improves if early successional herbaceous species and the burned canopy are not removed [11]. In this sense, herbaceous and shrub species from A. chilensis understory tended to increase their presence with wildfires [17,97]; as most of them harbor AMF [27,98], they could improve AM inoculum availability and colonization of new A. chilensis seedlings, although the specificity of A. chilensis-associated AM fungi in relation to understory AM fungal species has not yet been studied. Also, mammals, invertebrates, wind, water, and humans have all been documented as dispersal agents for AM fungi [99], although the specific role of each disperser has not yet been studied for Andean Patagonian forests. We also found low seedling colonization, mostly with explorations and storage structures (hyphae and vesicles), in treatments with soils of the severely burned forests. Despite the fact that very little is known in this regard, considering that AM symbiosis is crucial to new plant establishment [100], our findings indicate that inoculation with AM fungi should be recommended for the ecological restoration of forests affected by wildfire. Further research should examine its effects on seedling establishment in post-fire plantations.

Post-fire impacts include the significant alteration of soil properties, such as pH, OM%, EC, N, C/N ratio, and CEC [55]. It was stated that soil pH increases after wildfires [101] due to the combustion of organic matter [102]. The direct impact of lower pH on AMF is the limited sporulation and germination of spores in the rhizosphere; soils with neutral pH revealed greater amounts of AMF spores compared with acidic, pH 5.5 soils [103]. When pH rises, as our results may indicate for burned soils, stimulation of AMF sporulation can occur, increasing spore densities, as found in sites moderately affected by fire. Arbuscular mycorrhizal fungi can also be stimulated by organic matter contents [28,104]. In this sense, our seedlings revealed that AM structure colonization was consistently associated with OM soil content, even when the wildfire, regardless of WFS, showed significantly lower percentage of OM.

Arbuscular mycorrhizal spore density in natural soils/environments is highly variable [28] and experimental studies have shown that several factors could be involved in AMS sporulation, such as climatic, geological, and/or edaphic features, plant-associated phenology, human activities, and soil legacy [28,104,105,106]. In this sense, the unburned sites in this study showed a noticeable increase in AMS density after 5 years. Although our sampling design does not allow us to establish causalities, because we analyzed one plot of severity by site, considering that the sampling plots and methodology were not changed, it could be hypothesized that these differences could be related to previous climatic conditions for each sampling period (January–February 2016 and 2021). On average, the four autumns prior to the first sampling were warmer and drier, with fewer days with frost in 2016 than in 2021 (Table S7). Other studies have found that warming [107,108] and drought conditions [109] significantly decrease AMS abundance. More local studies are needed to unravel this issue, understand its dynamics, and predict evapotranspiration effects on the AMS bank.

When thinking of restoration strategies, it must be considered that wildfires would also cause extramatrical mycelium loss, although some stimulus occurs in moderately affected soils that increase spores’ abundance. Moderately burned areas have better chances of natural restoration considering the abundant AM spore bank, while artificial inoculation must be considered for severely affected areas or in burned dry sites.