Research ArticleResearch Article

Castanea dentata Interactions and Ectomycorrhizal Colonization in Novel Ecosystems

Jenise M. Bauman, Jennifer Franklin and Amy Santas
Ecological Restoration, September 2022, 40 (3) 179-190; DOI: https://doi.org/10.3368/er.40.3.179

ABSTRACT

Trial plantings of Castanea dentata (American chestnut) are currently underway using reclaimed surface mines as reintroduction sites for hybrids bred for disease resistance to canker causing Cryphonectria parasitica (chestnut blight). Coalfields of the eastern US coincide with the species range, but little is known of its ecological interactions as it had largely been extirpated from the landscape by the 1940s. Non-native species such as Lespedeza cuneata (sericea lespedeza) have been planted extensively in the region for both agriculture and mine reclamation and often dominate disturbed sites to form novel systems in place of native forests. We investigated the influence of herbaceous species on BC2F3 chestnut hybrid establishment and ectomycorrhizal (ECM) root colonization on three surface mines in eastern Tennessee that were restored using methods to avoid soil compaction. Hybrid chestnut growth, vegetation density, and species composition surrounding each seedling was assessed after eight growing seasons. ECM fungi present on root samples was quantified and identified by DNA sequencing of the internal transcribed spacer (ITS) region. Sites differed significantly in herbaceous cover and seedling growth. Larger seedlings were found on the site that was comprised primarily of L. cuneata. Average ECM colonization was similar across sites, and groundcover vegetation diversity or L. cuneata did not influence ECM species richness. There was a difference in ECM community composition among the three sites with some fungi displaying subtle changes in species rank. Results suggest that in the absence of soil compaction, L. cuneata does not impede hybrid chestnut establishment or ECM root colonization in Appalachian afforestation projects.

Keywords:

Restoration Recap

  • This study documented that when compaction is avoided upon reclamation, non-native L. cuneata does not inhibit tree growth in soils with pH, nutrients, and soil structure adequate for BC2F3 chestnut hybrid establishment.

  • Management may de-emphasize the need for herbicide for continual maintenance of L. cuneata when methods that avoid or mitigate compaction, coupled with competitive tree species, are employed.

  • L. cuneata's abundance in the herbaceous canopies (70−87% cover) that surrounded chestnut trees did not interfere with field ectomycorrhizal root symbiosis.

  • Documenting fungal species that correlate with abiotic soil conditions may contribute to our understanding of how ECM communities reassemble during soil recovery in areas that have been severely degraded and aid in management decisions when approaching a reforestation project.

There are many challenges in restoring an extirpated species to a landscape, including a limited knowledge of ecosystem interactions prior to its loss. For species that have been absent from the landscape for an extended period, changes in climate, effects of human land use, and introduced species and pathogens may have changed the environment in a way that has a critical impact on the species of interest. Castanea dentata (American chestnut) was once an ecologically and economically important species dominating eastern forests. However, by the 1940s this tree species was largely extirpated from the eastern North American forests due to an introduced pathogen Cryphonectria parasitica, the causal agent of chestnut blight (Gravatt 1949). Even before detection of chestnut blight in 1904, the species had been impacted by dramatic shifts in fire regimes and Phytophthora root rot, which influenced forest dynamics as oaks and other less common species came to dominate the landscape (Anagnostakis 2001, Elliott and Swank 2008, Nowacki and Abrams 2008). In the latter part of the 1900s timber harvesting, agriculture, and mining were the primary impacts on forested ecosystems (Drummond and Loveland 2010). Most recently, a changing climate and a significant increase in the use of non-native, invasive species in agriculture and reclamation has further changed the dynamics of eastern US forests (Franklin et al. 2012). Given the magnitude of these changes, the potential role of American chestnut in today's forests is unknown.

Efforts to save the iconic American chestnut through selective breeding programs began as early as 1907 and continue to this day, with the largest program being carried out by The American Chestnut Foundation (TACF) (Anagnostakis 2012). Asiatic species of Castanea carry a natural resistance to chestnut blight and have been incorporated into back-crossed breeding programs that attempt to introduce blight resistance into hybrids while maintaining characteristics of the native species (Burnham 1988, Steiner et al. 2017). Research is currently assessing growth and blight resistance of multiple advanced breeding lines; however, greater time is required to ascertain field resistance to chestnut blight for range-wide reintroduction (Clark et al. 2016). Species Distribution Models have been created for C. dentata based on estimated future climatic suitability (Fei et al. 2012, Barnes and Delborne 2019). The target range for re-introduction corresponds closely with the coal fields of the eastern US (McCarthy et al. 2008). This has influenced the partnership between TACF and the Appalachian Regional Reforestation Initiative to use reclaimed mine sites as locations for initial range-wide plantings.

Surface mining for coal has removed nearly one million hectares of native forests in the eastern US and induced long-term changes in soil properties and vegetative communities (Burger et al. 2017). The Forestry Reclamation Approach (FRA), a methodology that encourages healthy tree establishment on reclaimed mine sites, is now used extensively throughout the region (Zipper et al. 2011a). However, the sites have greatly altered soils and hydrology, and are rapidly colonized by introduced and invasive species that dominate the surrounding landscapes (Williamson and Barton 2020, Sena et al. 2021). The failure of trees to colonize older mine sites in the eastern US (reclaimed between 1977 and 2000) is likely due to soil conditions and the persistence of aggressive, non-native herbaceous communities that were seeded upon reclamation (Franklin et al. 2012). Lespedeza cuneata (sericea lespedeza), has been commonly used to revegetate during reclamation due to its tolerance to soil compaction, drought conditions, and low levels of soil organic matter (Zipper et al. 2011b). As a legume, this plant associates with nitrogen (N) fixing bacteria and was used to enhance nutrient poor soils. Lespedeza cuneata has been cited to associate readily with native arbuscular mycorrhiza, increasing access to phosphorus, which further contributes to its fast growth rates (Burrows and Pfleger 2002). Abundant and dense canopies (> 1 m tall) created by L. cuneata have been suspected to inhibit native tree establishment particularly when invasive cover exceeds 60% on reclaimed mines (Franklin et al. 2012, Sena et al. 2020).

The planting of native tree seedlings into a primarily non-native herbaceous cover has been a practice since the 1990s and is now common on a global scale (Weidlich et al. 2020). The persistence of these non-native forbs alters nutrient regimes, water relations, microbial communities, and greatly reduces the abundance of pioneer shrub and tree species that support the ectomycorrhizal fungi (ECM), an important soil symbiont to many later arriving native trees (Amaranthus and Perry 1994, Eviner and Hawkes 2008). In addition, it is hypothesized that some invasive species succeed in inhibiting growth of native trees by degrading ECM mutualism, despite having arbuscular mycorrhizal symbionts (Pinzone et al. 2018). Previous studies have documented the benefits ECM fungi has on chestnut growth in greenhouse trials, which can help mitigate stress upon field plantings (Aryal et al. 2020). Further, abandoned mine restoration projects have demonstrated that early colonization of ECM fungi with hybrid chestnut improves seedling establishment and lowers heavy metal accumulation (Bauman et al. 2012). Establishing these symbiotic relationships have been found to increase foliar carbohydrates as well as foliar nitrogen in blight resistant chestnuts, which may aid in mitigating the cost of blightresistance in exchange for plant productivity (Rieske et al. 2003). Understanding the impact of non-native species invasion on functional ECM formation becomes an integral part of successful tree establishment within mine land restoration.

The objective of this study was to evaluate BC2F3 hybrid chestnut root colonization by ECM fungi to see how that might be affected by the presence of an invasive species in restored sites in eastern Tennessee. This included percent root colonization as well as differences in ECM community composition, and whether those differences may be influenced by the presence of non-native L. cuneata. In addition, herbaceous plant species composition within varying abiotic site factors were recorded and used to determine which factors influenced chestnut growth in these restored mine sites. Ultimately, these factors were assessed to determine conditions most conducive to chestnut establishment on mine lands, and extrapolated to the broader, southern and central Appalachian regions.

Methods

Study Sites

This study used plots on three reclaimed mine sites in eastern Tennessee. Each shared an FRA recommended protocol of end-dumping to avoid soil compaction (Sweigard et al. 2017). Planting medium was prepared in 2008 by loosely incorporating a mix of weathered brown strata to yield a non-compacted medium for planting. All sites were then broadcast seeded with an herbaceous seed mix consisting of Medicago sativa (alfalfa), Panicum virgatum (switchgrass) and Solidago nemoralis (goldenrod).

The first site (36°30'03" N, 84°16'39" W) is at an elevation of 700 m (2250 ft), on a slope with an eastern exposure and coarse substrate. The seeded cover species established poorly, and the entire site was rapidly colonized by L. cuneata which had been planted in the surrounding area as part of the reclamation seed mix. Vegetation over most of the site is a continuous, dense layer up to 1 m in height, with very few small patches of bare soil remaining at the time of sampling. This is referred to as the “Zeb Mountain” site.

The second site (36°31'30" N, 83°57'23" W) is at an elevation of 594 m (1950 ft) on a steep, west-facing slope. The seeded ground cover species established poorly, with the exception of alfalfa, which covered approximately 25% of the area on which it was planted, and persists at low levels. Paulownia tomentosa (a non-native invasive tree, princess tree), and Robinia pseudoacacia (a native leguminous tree, black locust) established on the site soon after reclamation, and quickly formed a closed canopy over portions of the site. Lespedeza cuneata has been widely planted on surrounding reclaimed mine sites, and established on the site in large, dense patches in the open areas. Some patches of bare soil remain, particularly where there is shading from the overstory, but both native and non-native species have colonized the site. This is referred to as the “Mountainside” site.

A third site, referred to as the “Premium mine” site (36°06'21" N, 84°19'28" W), is on a steep northwestern-facing slope at an elevation of 900 m (2950 ft). This is a relatively narrow and isolated site, surrounded by native hardwood forest. The seed mix established slowly, and at the time of chestnut planting was less than 10% vegetation cover of the site. Colonization of the site by vegetation has been slow with large patches of bare ground remaining at the time of sampling.

The BC2F3 breeding generation used in this study was collected as seed from TACF open-pollinated orchards at Meadowview, VA, US and has a suspected 88% inheritance from the American chestnut (C. dentata) and the remainder from the Chinese chestnut (C. mollissima) parent (Burnham 1988). Chestnut seeds were grown by the State Nursery in Georgia in 2008 and planted as one-year-old bare root seedlings in March of 2009 in each of the three mine sites. Four, 30 m × 12 m plots were established with a 20 m buffer between each plot. Within each plot, 18 BC2F3 chestnut hybrids were planted (18 chestnuts × 4 plots × 3 mine sites = total of 216 seedlings). Chestnut hybrids were interplanted with one year-old black cherry seedlings in an alternating manner using a dibble bar in a 3 × 12 grid with a 2 m spacing between each seeding. Elevation, aspect, and slope were recorded at the time of planting.

Tree and Ground Cover Measurements

In September of 2016, all surviving chestnut trees were located, which included 75 trees: 28 from the Zeb Mountain site, 27 from the Mountainside site, and 20 from the Premium mine site. A 1 m square quadrat was placed at the base of each surviving chestnut tree, with the stem at the quadrat center for ground cover measurements. The percentage of bare ground, and the percentage of ground shaded by the canopy at mid-day (at the time of measurement, between 10:00 am and 3:00 pm) were estimated to the nearest 5%. In June of 2017, the height (cm measured with meter stick) and root collar diameter (cm measured with caliper) of each tree was measured. A 1 m square quadrat was placed around each tree and ground cover species occurring within each quadrat were recorded. Most herbaceous species were identified to genus, although a small number of plants that were encountered were too immature to identify and were classified as grass, legume, or other.

Root and Soil Collection

The 75 chestnut trees measured were also selected for nondestructive root sampling in September of 2017. These sites were extremely rocky and difficult to sample with cores or probes. Soil and rock were carefully removed with a spade to expose the chestnut root system at a depth of 25 cm and a width of 45 cm. Roots from a depth of 18 cm were carefully sifted away from the rocks and soil, stored on ice, and returned to the laboratory for further analysis (detailed below).

Soil samples were also collected at this time. To ensure randomization, four trees were randomly selected and soil was collected 1 m from the base of the tree and at a depth of 0−18 cm using a hand trowel. The soil samples were mixed thoroughly, oven dried for 36 hours at 70°C and 0.50 liters were sent to the Agricultural Laboratory at Clemson University in Clemson, SC for analysis. Soil variables such as cation exchange capacity, phosphorus, potassium, calcium, copper, manganese, magnesium, zinc, boron, and sodium were measured using the Mehlich 1 extraction method followed by inductively coupled plasma mass spectrometry (Isaac 1983). Nitrogen was extracted from dried soil samples measured by an ion electrode and recorded as NO3-N ppm. At Western Washington University, total percent N in leaf tissue was measured using Thermo Electron NC Soil Analyzer Flash EA 1112 Series (Thermo Electron Corporation, Milan, Italy) and dried soil organic matter was determined using weight loss-on-ignition by measuring weight of sample before and after combustion with a muffle furnace heated to 550°C for four hours.

DNA Extraction and Purification

In the laboratory, root samples were washed using distilled water, cut into 3 cm segments, and stored in Petri dishes at 4°C and processed within 14 days. All samples were observed under a dissecting microscope for ectomycorrhizal formation determined by the presence of a fungal sheath. One hundred root tips per seedling were randomly selected from root segments. Using a grid template as a guide, the number of ECM root tips was divided by the total number of roots sampled up to 100 tips to calculate percent ECM. ECM was morphotyped based on their color and texture of sheath, emanating hyphae, and presence of rhizomorphs (Bauman et al. 2013a). Morphotypes were counted and used for the subsequent multivariate analysis (described below). Two samples per morphotype per seedling were selected for DNA extraction. A three mm segment of root tip was removed and transferred into a microcentifuge tube and stored at -70°C until extraction.

Presence and type of ECM species on the root tip was determined by DNA extraction followed by PCR and DNA sequencing. Briefly, the 3 mm segment was homogenized using a bead beater and DNA was extracted using the QIAGEN® DNeasy Plant Mini Kit per manufactures protocol (QIAGEN, Germantown, MD). About 10 ng of this DNA was used for PCR amplification using primers ITS1-F (5' cttggtcatttagaggaagtaa 3') and ITS4 (5' tcctccgcttattgatatgc 3'), which targeted the highly variable internal transcribed spacer (ITS) region of ECM fungal ribosomal DNA (Gardes and Bruns 1993). PCR reactions were based on the following concentrations for a 25 μl reaction: 12.5 μl of GoTaq® Green Master Mix (Promega, Madison, WI), 0.25 μl of 25 μM of each primer, 11 μl of molecular grade water, and 1 μl of DNA template. Temperature cycling was accomplished using GeneAmp PCR systems 9700, which allowed for a programmable Thermal Cycler Heating regime as described by Gardes and Bruns (1993): The initial denaturation step was 94°C for 85 s followed by 35 amplification cycles of denaturation, annealing, and extension. The temperature and times for the first 13 cycles were 95°C for 35 s, 55°C for 55 s, and 72°C for 45 s. Cycles 14−26 and 27−35 repeated the above parameters with lengthened extension steps 120 and 180 s, respectively. When the 35 cycles were completed, the samples were programmed to incubate for 10 min at 72°C for 45 s. The PCR products were confirmed using gel electrophoresis and purified using Wizard® SV 96 Genomic DNA Purification System (Promega, Madison, WI) and DNA concentration was quantified using a Thermo Scientific 2000 1-position Spectrophotometer (Thermo Fisher Scientific, Pittsburg, PA) prior to sequencing. Sanger sequencing was performed using The Applied Biosystem ABI Prism 3730 DNA Analyzer (Plant-Microbe Genomics Facility, the Ohio State University, Columbus, OH). The DNA sequences were analyzed and edited using the Sequencher 4.2 software (Gene Codes, Ann Arbor, MI) and compared to sequences present in the GenBank (Altschul et al. 1997).

Statistical Analysis

All statistical analyses were performed using R software (Version 1.3.1093, R Development Core Team 2009). To compare BC2F3 chestnut hybrid survival by mine site, a Pearson's Chi Square (X2) was used. To understand the abiotic differences among the three sites used in this study, a multivariate analysis of variance (MANOVA) followed by one-way analysis of variance (ANOVA) and Tukey's HSD post-hoc test was used. Site was a fixed factor and the response variables included soil chemistry collected via soil samples and physical environmental features such as elevation, aspect, slope, and tree overstory. To compare the biotic variables among the three sites, individual trees were used as independent sampling units to characterize differences in chestnut growth (height, root collar diameter, canopy), herbaceous community (percent herbaceous cover, percent bare ground, organic liter, and number of herbaceous species) and ECM (percent root colonization and number of ECM species). Using site as a fixed factor, a one-way ANOVA was used followed by a Tukey's HSD post hoc test. If variables did not meet the assumptions of normality and equal variances by data transformations, a nonparametric Kruskal-Wallis analysis followed by a Dunn's post hoc test was used.

For multivariate data, the vegan package in R was used (Oksanen 2010). A permutational multivariate analysis of variance (PERMANOVA) was used to analyze ECM fungal communities using the adonis function. This was followed by a non-metric multidimensional scaling (NMDS) ordination using metaMDS function, which employed Bray-Curtis dissimilarities to determine differences in ECM species composition among mine sites. A Pearson's correlation analysis using envfit was employed to test for associations between ECM by site using all site variables that were square root transformed and standardized via Wisconsin double standardization using the Wisconsin command in R. Using the indicspecies package, a multipattern analysis was used to identify the combination of vegetation with a highest association value with the perspective sites using the multipatt function (De Caceres and Legendre 2009). For all analyses significance was determined by α = 0.05.

Results

Tree Survival, Growth and Herbaceous Ground Cover

After eight field seasons no differences existed regarding BC2F3 chestnut hybrid survival (Table 1). When the three reclaimed mine sites were compared with regard to tree growth and groundcover vegetation, taller trees (2.0 ± 2.8 and 2.3 ± 1.7 m) with greater production of canopy shade (46 ± 5.6 % and 53 ± 5.1%) were sampled in the Mountainside and Zeb Mountain sites, respectively, when compared to trees in the Premium sites (1.4 ± 2.9 m and 33 ± 5.1%; Table 1). Root collar diameter differed among all sites with larger diameters in the plots within Zeb Mountain (4.1 ± 0.2 cm), followed by Mountainside (2.8 ± 0.4 cm), and Premium (1.8 ± 0.1 cm). Sites also differed regarding total herbaceous cover and bare ground, Zeb Mountain had the greatest cover (85 ± 1.4%) with the least amount of bare ground (0.1 ± 2.1 %), followed by Mountainside cover (70 ± 13.7%) and bare ground (16 ± 3.3 %) and Premium (41 ± 5.2%) and (19 ± 2.7%). However, the Premium site was greater in herbaceous vegetation species richness (9 ± 0.8) and organic litter (28 ± 2.6 %) when compared to Zeb Mountain species (2.6 ± 0.8) and litter (13.4 ± 2.1%) and Mountainside (5 ± 0.8) and (11.8 ± 3.1%). The complete vegetation list is reported in Supplemental Material Table S1.

The three sites, Zeb Mountain, Mountainside, and Premium, differed with regard to physical site variables and soil chemistry when tested using a MANOVA (Pillai's Trace = 1.99, F2,9 = 144.3, p < 0.001). The following site differences are listed in Table 2. The Zeb Mountain reclamation site had lower slope, eastern facing aspect, lower pH and soil K. Mountainside had the lowest elevation with the greatest overstory noted among all the sites. This site also had the greatest available soil P, Mg, Mn and Cu. The Premium reclamation site was farthest from the forest edge, had the greatest elevation, and lowest in soil Zn and B.

A NMDS ordination was used to illustrate the ECM fungal community composition within the three respective sites (Mountain Side, Zeb Mountain, and Premium; Figure 1). This pattern was supported by a PERMANOVA, which showed a significant site effect (F2,9 = 4.74, p = 0.005, Figure 1A). Site variables are imposed as vector lines on the NMDS ordination to illustrate how the biotic and abiotic factors correlated with fungal communities sampled from the various sites (Table 3). Overall, NMDS 1 explained variation in soil chemistry where NMDS 2 explained the biotic factors such as herbaceous vegetation and tree growth (Figure 1B–D). With regard to soil chemistry (Figure 1B), Cu, Mg, P, Mn, Zn correlated with fungal communities sampled from the plots in Mountain Side, where K was significantly associated with the fungal communities sampled from the trees within the Premium sites. Further, Figure 1C illustrates variables such as plant species richness and organic matter are significantly correlated to ECM community in the Premium sites, where chestnut height and basal diameter are associated with ECM species sampled from the Mountain side and Zeb Mountain sites.

When herbaceous vegetation was overlayed onto the ordination, Melilotus officinalis (sweet clover) was significantly associated with communities in the Premium site and L. cuneata was significantly correlated with communities sampled from Mountain side and Zeb Mountain (Figure 1 D). A multi-pattern analysis identified the combination of vegetation with a highest association value with the respective sites. L. cuneata was most associated as an indicator species within Zeb Mountain (87% cover) and Mountain Side (70% cover; p = 0.003). L. cuneata was not sampled in the Premium sites; within this site, M. officinalis, Parthenocissus quinquefolia (Virginia creeper), and Clematis virginiana (clematis) were indicators of this site (all p < 0.05; Supplemental Material Table S1).

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Table 1.

Mean comparison (± 1 SE) of ECM, chestnut tree growth and herbaceous vegetation from various reclamation sites: Zeb Mountain, Mountain Side and Premium. The same letters subscripted above SE do not differ at α = 0.05 based on a Tukey's HSD with an F(2,9).

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Table 2.

Physical and soil chemistry variables compared among the three sites. A MANOVA followed by an ANOVA was used to detect significant differences among the three sites: Zeb Mountain, Mountainside and Premium. Values in the table are the means with ± 1 SE. Significant differences (p < 0.05) are shown in bold text with similar letters subscripted above SE do not differ at α = 0.05 based on a Tukey's HSD.

ECM species richness or percent colonization did not differ among the sites (Table 2). The most common ECM species sampled throughout all sites where Cenococcum geophilum (39%) and Cortinarius decipiens (15%, Table 4). ECM community composition differed among the sites (PERMANOVA, F2,9 = 4.74, p = 0.005). In the Premium sites, C. geophilum significantly increased, and Cortinarius vernus and Russula pectinatoides were unique to that site (Table 4). Certain fungi such as Hebeloma vaccinum, Tomentella sp., and Thelephora terrestris were found exclusively in the Zeb Mountain sites with high L. cuneata cover. Mountainside plots harbored Pisolithus, Inocybe sp. 2, Tuber sp., and Helotiaceae. Other species found in smaller proportions throughout the sites included: Scleroderma and other Inocybe, Hebeloma and Cortinarius spp. (Table 4).

Discussion

Chestnut Growth and Vegetation

After eight years, herbaceous canopies (< 1 m in height) growing around the trees differed among the sites regarding vegetation cover and composition, bare ground and plant species richness. Zeb Mountain and Mountain Side had the largest trees growing within an herbaceous canopy that consisted primarily of non-native L. cuneata (87% and 70%, respectively). This was unexpected in the framework of invasive plant species where competitive herbaceous plants can inhibit tree survival and growth on reclaimed sites (Franklin et al. 2012). However, benefits of vegetation have been reported from other Tennessee sites with relation to higher temperatures on unshaded, coarsely-textured soils and tree growth attributed to improved water relations with associated groundcover (Franklin and Buckley 2006). Fields-Johnson et al. (2012a) also reported improved water infiltration correlated with groundcover vegetation that may have improved water movement to the subsurface on sloping mines in Virginia. Regarding chestnut establishment, Gilland and McCarthy (2012) attributed increased growth and survival of trees with higher vegetative cover that decreased solar exposure on an end-dumped reclamation site in eastern Ohio. After 15 years, pure American tree heights averaged 6.7 m (per. comm. R. Homsher). In southwestern Virginia, initial B2F3 hybrid chestnut survival and growth were greater in a less-dense ryegrass cover; however, over time, the plots that utilized the tree compatible mix (cover that ranged from 51−100%) harbored taller trees, which were similar in growth to the B2F3 hybrids on Zeb Mountain (2.3 m after 8 years; Fields-Johnson et al. 2012b, Klopf et al. 2018).

What may have been more influential was lespedeza's association with N-fixing bacteria, which is thought to enhance nutrient poor soils during reclamation in the absence of compaction (Zipper et al. 2011a). Sena et al. (2015) reported that abundant, volunteer L. cuneata recruitment on an end-dump site established in eastern Kentucky and did not interfere with tree growth (chestnut not tested) or native herbaceous species recruitment after nine field seasons. The authors further speculated that L. cuneata may aid in the soil development, particularly in the addition of bioavailable nitrogen, while allowing for the eventual colonization of shade-tolerant native species over time. This too has been observed in sites utilizing other methods to alleviate soil compaction. On legacy mine sites reclaimed by continual grading and seeding of exotic herbaceous species, deep-soil ripping utilizing a steel shank (> 1 m depth) is used to alleviate compaction. Sena et al. (2020), when employing soil ripping, reported Pinus taeda (loblolly pine) to overtop lespedeza canopies after 10 years on Kentucky coal mined landscapes but noted poor growth and survival of slower growing species such as Quercus rubra (red oak). In contrast, Lang et al. (2009) reported a significant decrease in loblolly pine survival when seedlings were planted among L. cuneata and Chamaecrista fasciculata (partridge pea) on a legacy reclamation site in Mississippi. Regarding chestnut hybrids, Bauman et al. (2017) reported B2F3 hybrid tree growth to overtop lespedeza and tall fescue co-dominant herbaceous canopies when soil ripping was applied on an eastern Ohio reclamation site. Tree heights averaged 2.3 m after eight years, which was similar to B2F3 hybrids growing on Zeb Mountain and southwestern Virginia.

Figure 1.
Figure 1.

Non-metric multidimensional scaling (NMDS) ordination using Bray-Curtis dissimilarity index (stress = 0.07) of ectomycorrhizal fungal species samples from chestnut root tips. Fungal species are illustrated using ECM codes (defined in Table 4) with diamonds (♦) representing the sampling sites. The pattern reveals that ECM species differed among the three sites, as illustrated by 95% ellipses. Panels B–D) Plot vectors indicate strength and direction of the strongest correlations between ECM species, sites, and study site variables (including tree growth). This includes tree growth measured as height (HT) and root collar diameter (RCD).

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Table 3.

Site variables reported with coordinates (NMDS1 and NMDS2 axes) and with both the r2 and p-values that explain the strength of the relationship between ECM fungal species sampled from the roots of eight-year-old chestnuts and the associated environmental variables across the sites. Significant (p ≤ 0.05) are designated by single asterisks.

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Table 4.

Ectomycorrhizal (ECM) fungal species list generated from root tip count data from eight-year-old chestnut roots pooled from the three sites (n = 150). ECM genera are listed alphabetically along with their abbreviation used in the NMDS ordination (Figure 1). Abundance is shown as percentages generated from root tip count data. Total ECM species richness is reported below.

Hybrid chestnuts growing on the Premium site, in contrast to Zeb Mountain, were considerably smaller and had evident differences in the composition of herbaceous groundcover around the trees. Specifically, after seven years, there was an increase of native groundcover species richness due to the initial seeding, no L. cuneata, and significant increase in bare ground. There were also abiotic factors that differed among this site compared to Zeb Mountain which included higher pH of 5.7 (though still within the 4−6 pH range for chestnut) and differences in soil macro and micronutrients. Sena et al. (2018) reported pH to be an important driver for soil chemistry on an eastern Kentucky end-dumped site, specifically regarding phosphorus acquisition which may have contributed to growth rates for American chestnut (4.9 m heights in 4.9 pH) after 10 years, which was comparable to Quercus spp. and Robinia pseudoacacia (black locust). Skousen et al. (2018) found that weathered brown sandstone (pH 4.6) was a much better soil replacement substrate than unweathered gray strata (pH 6.3) for native tree species; however, they noted the poor performance in chestnut hybrid survival growth and tree vigor regardless of soil type (average 44 cm in height after eight years). Mountain Side and Premium sites shared similar pH (5.7), slope and aspect, which positioned chestnut trees on steeper slopes with western exposures. This questions whether lespedeza canopies on the Mountain Side plots alleviated stressors such as increased solar radiation, increased soil temperatures, improved water relations, and nutrient deficiencies, which have been cited as associated with western exposures (Fekedulegn et al. 2003). Because this current study was not designed to test the influence of L. cuneata in an experimental design across multiple, replicated sites, authors hypothesize that when compaction has been alleviated, lespedeza canopies may have a facilitation effect on fast-growing, competitive tree species like chestnut.

Ectomycorrhizal Fungi on Chestnut

Despite the presence of the non-native L. cuneata on two of the sites, there were no differences in the amount of ECM root colonization or species richness among the three plots; all chestnut trees tended to be strongly ectomycorrhizal. This is unlike other studies within Appalachia that report invasive species success is contributed to the degraded mutualisms hypothesis. Ailanthus altissima (Tree-of-heaven) invading a mine site in central Ohio had a significant decrease in ECM roots and root:shoot ratio of red oak seedlings, suggesting inhibition of root development due to allelopathic effects (Small et al. 2010, Bauman et al. 2013b). Castellano and Gorchov (2012) reported a similar finding on red oaks in forest stands invaded by Alliaria petiolata (garlic mustard); this herbaceous species is suspected to produce antifungal compounds toxic to mycorrhizal fungi (Wolfe et al. 2008). Imperata cylindrica (Cogongrass) has also been shown to cause a decrease in ECM on loblolly pine in Mississippi (Trautwig et al. 2017).

There was a difference in ECM community composition among the three sites; communities differed in species and some fungi displayed subtle changes in species rank per site. This appeared in the ordination to be driven by site variables such as soil chemistry, size of chestnut tree, and herbaceous vegetation. This is not a novel finding, yet provides additional information to how mineral nutrients, host plants, and site conditions interact to influence fungal communities (Smith et al. 2002, Lilleskov et al. 2004, Burke et al. 2009). Fungal genera such as Cenococcum, Cortinarius, Hebeloma, Inocybe, Pisolithus, Russula, Scleroderma, Thelephora, and Tomentella have been consistently sampled from seven and eight-year-old chestnut trees planted for coal mine restoration (Bauman et al. 2018). These fungal genera may represent an early successional clade of fungi that associate with chestnut in disturbed reclamation sites and provide invaluable ecosystem functions such as biogeochemical cycling, primary productivity, and structuring of forest habitat by the creation of mycorrhizal networks (Clemmensen et al. 2013). Based on other studies, we would anticipate fungal succession to also proceed with site development. This includes the recruitment of other genera such as Amanita, Boletus, Lactarius and Laccaria, and greater richness of Russula and Cortinarius species (Palmer et al. 2008, Dulmer et al. 2014, Stephenson et al. 2017).

Black hyphal ECM species C. geophilum appeared in the ordination strongly associated with the Premium sites, which corresponded to smaller chestnut trees, less vegetation cover, higher elevation, and a western exposure. Although some studies have documented C. geophilum on weakened trees associated with declining oaks (Montecchio et al. 2004, Clemmensen et al. 2013), this has not been observed with chestnut tree dieback caused by Phytophthora root disease or chestnut blight (Corcobado et al. 2015, Bauman et al. 2018). Other studies report C. geophilum to increase in the absence of better competitors (Dickie et al. 2005), or conversely, selected for its ability to mitigate water stress or other most limiting soil resources (Pigott 1982, Bödeker et al. 2009). This study, as well as previous work with chestnut, has shown C. geophilum to co-exist with a robust hyphal cord producing ECM Cortinarius spp. and correspond to sites during drought periods (Bauman et al. 2017). Other fungi such as Hebeloma vaccinum was unique to Zeb Mountain, and has been previously described as a dominant fungus associated with chestnut seedlings (Bauman et al. 2013a). These results pose new questions regarding drivers of symbiont selection for chestnut seedlings on mine sites where stressors such as competition, soil conditions, water relations, and solar radiation may control for subsequent species mutualisms during establishment.

Conclusion

Tree-ground cover interactions on mine soils are complex; negative or positive interactions will occur within overlapping niche space where competition or facilitation exists on a continuum driven by soil conditions (Franklin et al. 2012). Because L. cuneata will remain a persistent, weedy species on or adjacent to many of these sites, this species will continue to be a challenge to forest restoration in Appalachian coal mined lands. In addition, pressure of seed bank, ability to regenerate with vegetative propagules, and continued use in current reclamation projects will create an ongoing threat of re-invasion throughout the Appalachian region. The ability for tap root systems, like members of Fagaceae, to access deeper water reserves will require the alleviation of soil compaction to reduce the stress of water scarcity in the soil profile dominated by herbaceous competition (Pinto et al. 2011). This current study illustrated that when compaction is alleviated and the soil chemistry is conducive for forest tree establishment, chestnut hybrids will maintain ECM symbionts and growth rates required to co-exist with L. cuneata in the early years of establishment. As these stands develop, the deeper root growth with functional ECM symbioses would prove to be a significant advantage when forbs mature and evapotranspiration rates increase to further impose water stress in the upper soil profile (Skousen et al. 2009, Sena et al. 2014). Over time, mine sites restored to forested areas will further increase the efficiency of water infiltration deeper into soil profiles via tree root penetration (Clark and Zipper 2016).

Chestnut hybrid introduction in the 21st Century will be met with many challenges that include novel insects and pathogens (including C. parasitica), as well as prolonged droughts and challenging soil conditions in the age of rapid climate change (Barnes and Delborne 2019). Previous studies have indicated predictors such as acidic pH and steeper slopes drive American chestnut range potential (Tulowiecki 2020) and have been linked to forest models within chestnut's historic range (Wang et al. 2013). However, these site attributes may have a stronger influence on chestnut establishment in disturbed and unweathered soil substrates typical of reclamation sites in formerly coal mined lands in Appalachia. Therefore, the subtle differences in disturbed soils during afforestation which influence water relations during the driest months and increase water content and humidity, may have a substantial influence on chestnut growth in the early years of establishment (Zhang et al. 2019). Long-term research remains a necessity to better understand the putative resistance of the chestnut hybrid lines required to maintain chestnut's competitive advantage within the lespedeza novel systems. Though L. cuneata appeared not to inhibit growth and ECM colonization, many other non-native species also dominate reclaimed sites and will remain a challenge for land practitioners for decades to come. In areas that have transformed into novel systems, understanding the impact these new species have on the interactions of trees and their symbionts may aid in management decisions and planting methodologies best suited to initiate native forests.

Acknowledgments

This work was supported by National Technology and Transfer funds from the U.S. Department of Interior (Office of Surface Mining), The American Chestnut Foundation Research Support Grant, The Muskingum University Muskie Fellows Program, The Muskingum University Carolyn, Glenn Hodges Student Research Fund, Research and Sponsored Program Grant from Western Washington University, and McIntyre-Stennis grant number TEN00MS-111 accession no1014114 from the USDA National Institute of Food and Agriculture. We also thank Ryan Homsher and Zachary Griffith for their help in the field and Regina Baker and Colton Wilson for their help in processing the root samples.

Figure

Chestnut. Source: The flowers and leaves of a chestnut tree. Source: The Editorial Board of the University Society Boys and Girls Bookshelf (New York, NY: The University Society, 1920).

Footnotes

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Castanea dentata Interactions and Ectomycorrhizal Colonization in Novel Ecosystems
Jenise M. Bauman, Jennifer Franklin, Amy Santas
Ecological Restoration Sep 2022, 40 (3) 179-190; DOI: 10.3368/er.40.3.179
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