The importance of conifer plantations in northern Britain as a habitat for native fungi

The importance of conifer plantations in northern Britain as a habitat for native fungi

Biological Conservation 96 (2000) 241±252 The importance of conifer plantations in northern Britain as a habitat for ...

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Biological Conservation 96 (2000) 241±252

The importance of conifer plantations in northern Britain as a habitat for native fungi J.W. Humphrey a,*, A C. Newton b, A.J. Peace c, E. Holden d b

a Woodland Ecology Branch, Forest Research, Northern Research Station, Roslin, Midlothian EH25 9SY, UK Institute of Ecology and Resource Management, University of Edinburgh, Darwin Building, Kings Buildings, May®eld Rd, Edinburgh EH9 3JU, UK c Statistics & Computing Branch, Forest Research, Alice Holt Lodge, Wrecclesham, Farnham, Surrey GU10 4LH, UK d Allanaquoich, Mar Lodge Estate, Braemar, Ballater, Aberdeenshire AB35 5YJ, UK

Received 16 December 1999; received in revised form 6 May 2000; accepted 23 May 2000

Abstract Macrofungal assemblages of Sitka spruce and Scots pine plantations in northern Britain were compared to those of semi-natural pine and oak woodlands, with a focus on threatened pinewood taxa. Fungal species-richness and species-composition were related to climate, soil, vegetation and stand variables across a range of crop stages. Altogether, 419 species were recorded (12 parasites, 76 wood saprotrophs, 174 mycorrhizal species, 157 litter saprotrophs). There were no di€erences in fungal species-richness between plantations and semi-natural woodlands nor any e€ects of crop species age or type. Signi®cant positive correlations were recorded between fungal species-richness and ground vegetation diversity, and between wood saprotroph-richness and fallen deadwood volume. Each crop species type had a distinctive mycota related to di€erences in climate, tree and vegetation diversity. Over-mature stands had a higher proportion of ``late-successional'' mycorrhizal species than the other growth stages. Nineteen Red Data list fungi species were recorded; plots nearer to semi-natural pinewood areas had a higher number of species records. These results highlight the importance of planted forests as a habitat for native fungi. Habitat value could be further enhanced through increasing the area of ``old-growth'' non-intervention reserves, and locating these reserves near existing semi-natural woodland fragments. # 2000 Elsevier Science Ltd. All rights reserved. Keywords: Fungi; Conservation; Conifer plantations; Mycorrhizas; Saprotrophs

1. Introduction Although the fungal ¯ora of northern Britain is incompletely known, the fungal diversity of Scotland was recently estimated at more than 8000 species (Watling, 1997). Some fungal assemblages are particularly characteristic of Scotland, most notably those of the native or semi-natural pinewoods, which include a number of species found nowhere else in Britain (Orton, 1987; Watling, 1997; Tofts and Orton, 1998). Many of these pinewood species are of particular conservation importance. Out of the 250 fungal species listed as threatened with extinction in Britain, 42% are associated with native pinewoods (Anon, 1998c).

* Corresponding author. Tel.: +44-131-445-2176; fax: +44-131445-5124. E-mail address: [email protected] (J.W. Humphrey).

Concern about the conservation of fungi has increased in recent years, since the detection of declines in the abundance of many taxa across Europe (Derbsch and Schmitt, 1987; Fellner, 1989; Arnolds, 1991), apparently as a result of the combined e€ects of habitat loss and aerial pollution (Jansen and Van Dobben, 1987; Arnolds and De Vries, 1993). Although no evidence is available to assess whether such declines have occurred in northern Britain, these environmental factors could be as in¯uential in the UK as elsewhere (Dyke and Newton, 1999). Given the decline in the extent of native woodland habitat in the region, it is likely that many native fungi must have declined substantially in abundance over the long-term (Newton and Humphrey, 1997). Over the past 80 years, substantial areas of conifer plantation have been established in upland areas of northern Britain, and now cover some 13% of total land area (Anon, 1998a). These plantations are mostly composed of Sitka spruce (Picea sitchensis (Bong). Carr.)

0006-3207/00/$ - see front matter # 2000 Elsevier Science Ltd. All rights reserved. PII: S0006-3207(00)00077-X


J.W. Humphrey et al. / Biological Conservation 96 (2000) 241±252

and other non-native conifer species such as Norway spruce (Picea abies L.), Douglas ®r (Pseudotsuga menziesii (Mirb.) Franco), larches (Larix spp.) and ®rs (Abies spp.). The value of upland plantations as a habitat for wildlife is increasingly being recognized (Peterken, 1987; Petty et al., 1995; Newton and Humphrey, 1997), and current UK government policy encourages appropriate management to enhance their biodiversity (Anon, 1998b). However, little is know about the importance of upland plantations as a habitat for fungi, as few mycological surveys have been undertaken (Newton and Haigh, 1998). Recent data suggests that plantations could have the potential to support a signi®cant fungal diversity; some 151 species of ectomycorrhizal fungi have been found associated with Picea spp. in the UK (Newton and Haigh, 1998), and over 300 species of fungi recorded in a survey of lowland conifer plantations in England (Ferris et al., in press). There is currently little information on whether upland conifer plantations could contribute to the conservation of the characteristic native pinewood mycota, and rare or threatened species. The objectives of this study were to describe the fungal communities associated with Scots pine and Sitka spruce plantations established on a range of di€erent sites in northern Britain, and to relate species composition and diversity to climate, site and habitat structure variables. Plantation assemblages were compared to those of native pine (Pinus sylvestris L.) and oak (Quercus spp) woodlands in similar bioclimatic zones and on comparable site types. Particular attention was paid to the presence of threatened taxa characteristic of native pinewoods, in order to evaluate whether plantations of non-native trees may provide a suitable habitat for these species, and to suggest possible management strategies for enhancing habitat quality. 2. Methods 2.1. Study sites Assessment sites were selected from within the ``uplands'' and ``foothills'' bioclimatic zones of the Forestry Commission's Ecological Site Classi®cation (ESCÐPyatt and SuaÂrez, 1997), as part of a wider study of biodiversity in plantations (Humphrey et al., 1999). The zones are delineated by annual precipitation totals (uplands: >1500 mm; foothills: 800±1500 mm). Study sites were established in Sitka spruce plantations and native oakwoods in the uplands, and in Sitka spruce, Scots pine plantations and native pinewoods in the foothills. Two replicate sites were selected for each species  bioclimatic zone combination (Table 1). At each site, a chronosequence of four 1 ha (100 m x 100 m) permanent sample plots was established in forest stands encompassing di€erent growth stages. The four stages were:

(1) pre-thicket Ð age 8±10 years, crop height 2±4 m, incomplete canopy closure; (2) mid-rotation Ð age 20±30 years, crop height 10± 20 m, complete canopy closure, no understorey; (3) economically mature Ð age 50±80, crop height 20±25 m, some understorey development; (4) over-mature [beyond economic maturity and acquiring some of the ecological characteristics of natural old-growth forests sensu Oliver (1981)] Ð age 60±250 years, crop height >20 m, canopy break-up, well developed understorey, accumulation of deadwood. Only stages 2 and 3 were available in native oakwoods owing to a lack of large areas of newly regenerating woodland. Stages 1±3 were used in pine plantations (no stage 4 plantation origin stands were available). The two native pinewood stands sampled were classed as stage 4. The total number of plots sampled was 28. The previous land cover for the plots is given in Table 1. 2.2. Fungal assessments The presence/absence of macrofungal sporocarps (macroscopic ascomycetes and basidiomycetes) was recorded in eight, 1010 m quadrats. There were two of these quadrats arranged diagonally across the centre of each 5050 m quarter of the 1 ha plot. Assessments were made over the August±October period to coincide with the main time of sporocarp production. Three visits were made to each plot at roughly monthly intervals over this period, and repeated over three years. This sampling period allowed an estimated 80% of total species to be recorded for each plot (based on asymptotes for yearly species-accumulation curves). A total census would need a much longer time period and was beyond the scope of this study. Collections were identi®ed by reference to standard texts, involving microscopic examination where necessary. Material of particularly critical taxa was dried for reference and deposited in a national herbarium (Royal Botanic Garden, Edinburgh). Species were placed into functional groups following Newton and Haigh (1998) and Ferris et al. (in press): M Ð mycorrhizal; P Ð parastic; L Ð saprotrophic species on litter and other fungi; W Ð deadwood saprotrophs. 2.3. Climatic variables Climatic information for the 28 assessment plots was obtained from datasets held within the ESC computerbased decision support system (Ray et al., 1996). For both accumulated temperature (AT) and soil moisture de®cit (MD), 30 year means have been calculated for all 10 km squares throughout Great Britain using meteorological data collected over the 1961±1990 period (Barrow et al., 1993). AT expresses the degree of warmth or

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Table 1 Location and description of conifer and oak sites surveyed. Numbers after the decimal place in the site codes refer to stand stage (1 Ð pre-thicket, 2 Ð mid-rotation, 3 Ð mature, 4 Ð over-mature) Site code

Site name

Tree species

Bioclimatic zone

1 1.1 1.2 1.3 1.4 2 2.1 2.2 2.3 2.4 5 5.1 5.2 5.3 5.4 6 6.1 6.2 6.3 6.4 9 9.1 9.2 9.3 9.4 10 10.1 10.2 10.3 10.4 15 15.2 15.3 16 16.2 16.3

Glen A€ric Lochan Dubh, Cannich Knock®n Plodda Falls Beinn a' Mheadhoinn Strathspey Moor of Alvie An Slugan Glenmore Lodge Airgiod-meall Knapdale Dunardy Dunardy Kilmichael Dunardy Clunes South Laggan Clunes Clunes South Laggan Kielder Falstone Falstone Falstone Archie's Rigg Glentress Glentress Glentress Glentress Cardrona Taynish Taynish Taynish Beasdale/Moidart Beasdale Moidart

Scots pine


Scots pine


Sitka spruce


Sitka spruce


Sitka spruce


Sitka spruce


Native oak


Native oak


available heat energy (Bendelow and Hartnup, 1980) and is measured by the number of degree-days above 5 C. MD is expressed as the maximum accumulated amount that monthly potential evaporation exceeds precipitation (Bendelow and Hartnup, 1980), and is essentially a measure of dryness. Both MD and AT could indicate the suitability of climatic conditions for fungal fruiting. 2.4. Vegetation, stand structure and deadwood assessments Within each 5050 m quarter of the 1 ha assessment plot, vertical stand structure was assessed using a visual cover method. Four measurements, each 10 m apart were made along a North±South transect, running through the centre point of each quarter, yielding 16 measures in total for each 1 ha plot. Four vegetation strata were de®ned: S1 (®eld) 10 cm±2 m in height; S2 (shrub) 2±5 m; S3 (lower canopy) 5±15 m; and S4 (upper


Age (years)

Previous land cover

57 230 N 57 180 N 57 170 N 57 160 N

4 480 W 4 510 W 4 550 W 4 520 W

12 35 96 >150

Heath/grassland Heath/grassland Native pine Native pine

57 90 N 3 540 W 57 120 N 3 450 W 57 100 N 3 400 W 57 90 N 3 430 W

8 32 64 >150

Native pine Native pine Native pine Native pine

56 40 N 56 40 N 56 50 N 56 40 N

9 24 44 62

Sitka spruce Sitka spruce Heath/scrub Heath/scrub

57 00 N 4 520 W 56 580 N 4 590 W 56 580 N 4 590 W 57 00 N 4 530 W

8 28 62 67

Sitka spruce Heath/montane Heath/native woodland Heath

55 100 N 2 270 W 55 90 N 2 270 W 55 90 N 2 310 W 55 80 N 2 280 W

6 23 57 69

Sitka spruce Sitka spruce Grassland Grassland/mire

55 400 N 55 390 N 55 400 N 55 370 N

10 28 55 61

Sitka spruce Heath/grassland Heath/grassland Grassland

56 00 N 5 380 W 56 00 N 5 350 W

120 200

Native oak Native oak

56 530 N 5 450 W 56 470 N 5 450 W

120 200

Native oak Native oak

5 310 W 5 310 W 5 190 W 5 300 W

3 90 W 3 80 W 3 90 W 3 60 W

canopy) 15±20 m. Percentage cover of vegetation within each vertical stratum was described to the nearest 5% and expressed as a mean of the 16 stand structure measures. Stand growth measures were made within the eight internal 1010 m quadrats (see Section 2.2). In those plots where stocking density was low (e.g. stage 4 stands), the quadrats were extended (proportionately from each corner) to 2525 m to obtain a sucient sample of trees. Assessments were made of diameter at breast height (DBH), height to the base of the live crown (HTLC), and top height (TOPHT), for all trees >7 cm DBH within the designated areas. Tree species number per plot was also recorded (TREESP). Mean basal area (MBA) was calculated for each 1 ha plot following Hamilton (1975). Field and ground layer vegetation composition was assessed visually using the DOMIN cover-abundance scale (sensu Dahl and Hadac, 1941), within eight 22 m quadrats nested in the


J.W. Humphrey et al. / Biological Conservation 96 (2000) 241±252

centre of the 1010 m quadrats. DOMIN values were converted to percentage cover values by taking the midpoint percentage for each DOMIN score. Vascular plant and bryophyte diversity was calculated using the Shannon-Weiner index which takes into account the relative abundances of species as well as their occurrence (Magurran, 1988). Accumulations of fallen deadwood were recorded along two transects bisecting the 1 ha plot diagonally from the plot corners. Total volume of fallen deadwood (>7 cm DBH) was calculated using the line intercept method (Warren and Olsen, 1964). Volumes were calculated by assuming that deadwood items were cylindrical. Deadwood quality was described using a visual ®vepoint scale, where categories 1±3 corresponded to ``fresh'' decay classes (DEADF), 4 and 5 to advanced decay (DEADR) (sensu Hunter, 1990). 2.5. Soil and litter assessments Soil samples were taken from two layers: 0±5 cm and 5±10 cm, at 32 locations within each 1 ha plot (corresponding to the four corners of the eight internal vegetation assessment plots). The samples were bulked to give one sample for each layer per plot. Available P, K, Ca and Mg were obtained by extraction, using 0.5 M ammonium acetate/acetic acid solution at pH 4.5, following a modi®cation of Morgan's method (Morgan,

ÿ 1941). Mineralised N (in NH+ 4 and NO3 form) was determined before and after a 28 day incubation period at 30 C following the ADAS/MAFF method (Anon, 1986); pH was determined in aqueous solution using the MLURI/SAC method (Anon, 1985). Organic matter content was determined by loss on ignition. Litter depth was recorded in each 1010 m vegetation assessment plot by taking a mean of 4 random measurements.

2.6. Statistical methods Fungal community composition was examined using correspondence analysis (CA) (Hill, 1973) to provide vectors summarising the main gradients of variability amongst the sample plots. As an adjunct to this analysis, Jaccard similarity coecients were calculated to test for fungal species associations between plots. This analysis calculates the proportion of fungal species occurring in two plots relative to species occurring in just one of two plots. The CA vectors (FUNGI 1-4) together with values for species richness (number of fungal species. haÿ1 by functional group) were related to climate, soil, stand, and vegetation variables (Table 2) using correlation analysis. Principal component analysis (PCA) (Goodall, 1954) was used to produce four axes (SOIL 1±4) summarising variability in the soil chemistry data. The e€ects of crop species type and stand stage on fungal species-richness were analysed using a generalised linear

Table 2 Description of environmental variables related to fungal dataa Variable





First axis of the soil PCA ("pH, K, Mg, Ca) Second axis of the soil PCA ("pH, Ca); (#P, organic, NH+ 4 ) Third axis of the soil PCA ("NOÿ 3) Fourth axis of the soil PCA ("pH, NH+ 4 ) Vertical cover ®eld layer (%) Vertical cover shrub layer (%) Vertical cover lower canopy layer (%) Vertical cover upper canopy layer (%) Top height (m) Height to live crown (m) Basal area (m2 haÿ1) No. of tree species per plot Fresh fallen deadwood volume (m3 haÿ1) Rotten fallen deadwood volume (m3 haÿ1) Litter depth (cm) Accumulated temperature (no. day degrees >5 C) Soil moisture de®cit (mm) Species richness ground vegetation (no. haÿ1) Shannon-Weiner diversity index ground vegetation (H1) Horizontal cover ground vegetation (%) Species richness vascular plants (no. haÿ1) Shannon-Weiner diversity index vascular plants (H1) Horizontal cover vascular plants (%) Species richness bryophytes (no. haÿ1) Shannon-Weiner diversity index bryophytes (H1) Horizontal cover bryophytes (%)

ÿ1.84 ÿ1.76 ÿ1.33 ÿ1.54 0 0 0 0 2.07 0 0 1 0 0 0 771 24 4 1.32 0.1 3 0.62 0.08 1 0 0.03

3.15 2.87 2.44 2.31 72.6 36.6 52.5 31.6 32.94 15.67 59.97 10 159 77 5 1417 114 56 3.21 230.35 37 2.66 150.16 21 2.36 106.99


Values are per 1 ha plot.

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mixed model with poisson error distribution and log link function (for analysis of count data). Analyses for this paper were carried out using a combination of SAS (Anon, 1990), Genstat (Anon, 1993) and CANOCO (Ter Braak, 1988) programmes. 3. Results 3.1. Principal components analysis (PCA) of soil chemistry data The ®rst 4 axes of the soil chemistry PCA accounted for 86% of the variability in the soil data. SOIL 1 indicated a gradient of increasing pH, K, Mg and Ca, SOIL 2 a gradient of increasing pH and CA and decreasing P, ÿ organic matter and NH+ 4 , SOIL3 increasing NO3 and + SOIL 4 increasing pH and NH4 (Table 2).

Fig. 1. Frequency of fungal species in relation to number of plots in which they were found.

3.2. Fungal species-richness In total, 419 fungal species were recorded; a further 168 taxa could not be identi®ed to species (usually because sporocarps of sucient quality for reliable identi®cation were not available) and were not included in the results analysis. Most species were Basidiomycetes, the main genera being Cortinarius (46 species), Galerina (21 species), Inocybe (23 species), Lactarius (20 species) Mycena (37 species) and Russula (31 species). Only 21 species were Ascomycetes. Over 50% of species were recorded only once, and no single species was recorded in all 28 plots (Fig. 1). In fact only 7 species occurred in over 20 plots (Fig. 1). These were: Entoloma cetratum, Marasmius androsaceus, Mycena ®lopes, Mycena galopus, Mycena leptocephala, Mycena rorida and Mycena sanguinolenta. Twelve parasitic species, 76 wood saprotrophs, 174 mycorrhizal species and 157 litter saprotrophs were recorded. There were no signi®cant e€ects of stand stage, crop type or climate zone on total fungal species richness or the richness of the individual functional groups (Fig. 2). Total fungal species-richness and mycorrhizal species-richness was positively correlated with vegetation diversity (vascular plants and bryophytes; P<0.05), and with bryophyte diversity alone (P<0.05). The number of wood saprotroph species was positively correlated with fresh deadwood volume (P<0.05). There were no signi®cant (P<0.05) correlations with any of the other environmental datasets. 3.3. Di€erences in fungal species Ð composition between plots Only 27 species were common to the three crop types, with 78 species recorded in Scots pine plots only, 130 speci®c to Sitka spruce and 61 speci®c to oak. Oak and pine plots shared 9 species in common, oak and spruce

Fig. 2. Number of fungal species recorded in di€erent stand growth stages of Scots pine, oak and Sitka spruce.

30 species, and pine and spruce 81 species. These di€erences between crop types are further illustrated by the correspondence analysis of plots based on their fungal species composition (Fig. 3). The ®rst axis (FUNGI 1) of the ordination separated Scots pine sites (sites 1 and 2) from the other crop types, with axis 2 (FUNGI 2) separating the oak (sites 15 and 16) from the upland and foothill Sitka spruce sites (sites 5, 6, 9 and 10). Axis 3 (FUNGI 3) indicated a further separation between the upland and foothill spruce sites. Axis 4 (FUNGI 4) did not yield readily interpretable trends and was omitted from further analysis. The eigenvalues for the four axes were 0.43, 0.4, 0.33 and 0.28 respectively. Spruce plots were characterised by large numbers of Cortinarius and Inocybe spp (all mycorrhizal species), pine plots tended to have greater numbers of Russula and Suillus spp. and fewer Cortinarius and Inocybe spp. Oak plots had their own distinctive Russula ¯ora and no Suillus spp. Correspondence analyses were carried out for each functional group of fungi, but compositional trends were only evident for the mycorrhiza (Fig. 4). Axis 1 separated Kielder spruce plots (site 9) from the other sites, with axis 2 separating oak from spruce, and spruce from pine.


J.W. Humphrey et al. / Biological Conservation 96 (2000) 241±252

Fig. 3. Correspondence analysis ordination of assessment plots based on fungal species composition (all functional groups); (A) axis 1 v axis 2, (B) axis 1 v axis 3. Data labels for plots are coded as per Table 1. Numbers preceding decimal point refer to sites, numbers after decimal point refer to stand stage (1Ðpre- thicket; 2Ðmid-rotation; 3Ðmature; 4Ð over-mature). Table 3 Pearson correlations between environmental variables and the ®rst 3 ordination axes (FUNGI 1±3, see Fig. 3) of the CA of plots based on their fungal species compositiona


Fig. 4. Correspondence analysis ordination (axis 1 v axis 2) of assessment plots based on mycorrhizal species composition. Data labels for plots are coded as per Table 1. Numbers preceding decimal point refer to sites, numbers after decimal point refer to stand stage (1Ðprethicket; 2Ðmid-rotation; 3Ðmature; 4Ðover-mature).

FUNGI 1 was positively correlated with litter depth, accumulated temperature (AT) and moisture de®cit (MD). FUNGI 2 was positively correlated with the number of tree species per plot, and species richness and diversity of vascular plants and all vegetation species combined (Table 3). FUNGI 3 was positively correlated with SOIL 4 (gradient of increasing pH and major nutrients), and negatively correlated with vertical cover of ®eld and shrub layer vegetation (S1 and S2).




0.10 0.02 0.09 0.02 0.59* 0.58* 0.69* ÿ0.16 0.01 ÿ0.10 0.03

0.08 ÿ0.23 0.07 0.73* ÿ0.19 0.60* 0.38 0.68* 0.68* 0.73* 0.68*

0.58* ÿ0.49* ÿ0.52* 0.32 0.03 0.04 0.14 0.17 ÿ0.21 0.16 ÿ0.17

Signi®cant correlations indicated by * (P<0.05).

The Jaccard similarity analysis showed that plots within individual sites tended to share more species in common than plots from di€erent sites. Typically, similarity coecients were in the range of 30±45% for individual chronosequences. Each site tended to have 20±40 species speci®c to itself. In general, the more geographically separated the sites were, the less species they had in common. Coecients were lowest for oak versus pine plots, with values all less than 15%. There were also di€erences between stand growth stages, with each stage having 20±40 unique species (Fig. 5). Following the functional classi®cation of ectomycorrhizal fungi proposed by Newton (1992), 87% of the mycorrhizal species recorded in the over-mature stage were classed as ``late-successional'' forest species, whereas only 33% of species in the prethicket stage fell into this category.

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3.4. Fungi of conservation importance A number of rare and threatened species were identi®ed during the survey. One, Panellus violaceofulvus, found in the Glen A€ric pre-thicket stage plot (1.1), had not been recorded previously in the UK. A second species, Cortinarius callisteus (found in Clunes Sitka spruce plots 6.2 and 6.3) had not been con®rmed previously as native to Britain. Nineteen of the species recorded were on the provisional Red Data list for British fungi (Mace and Lande, 1991, Table 4) and are therefore considered to be threatened with extinction. Ten species were found in pine plots only (2 were speci®c to the native pinewood plots), 4 in spruce only, and 4 in both spruce and pine plots. Only one threatened species was recorded in the oak plots. Most of the threatened species recorded were mycorrhizal species; litter and wood saprotrophs were less well represented (Table 4). The number of species records was signi®cantly (negatively) correlated (P<0.001) with distance of the assessment site from an extant native pinewood area (Fig. 6). As the pine plantation plots (Glen A€ric and Strathspey) were all adjacent to, or had been established upon, previously native pinewood sites, they were given a default value of 1 km. 4. Discussion 4.1. Fungal species-richness and community composition Altogether, 419 fungal species were recorded during this survey. This ®gure compares favourably with those obtained from mycological surveys of other temperate and boreal forests. For example, 527 fungi species have

Fig. 5. Number of species speci®c to each stand growth stage.


been identi®ed to date from old-growth forests of the Paci®c North West region of North America (Marcot, 1997). Jonsson etal. (1999) found 135 species of ectomycorrhizal fungi in late successional boreal Scots pine forest in Sweden (compared with a total of 174 in this current study). In contrast Dahlberg et al. (1997) recorded only 48 mycorrhizal species in oligotrophic Norway spruce stands in southern Sweden. Records for the UK suggest ®gures of 151 mycorrhizal associates of Picea, 201 for Pinus and 233 for Quercus (Newton and Haigh, 1998). Hoiland and Bendiksen (1996) recorded 140 species of wood saprotroph in a survey of 93 Norway spruce stands in central Norway (76 species were recorded in the current study). Similarly, Lindblad (1998) recorded 118 species of wood-inhabiting fungi in Norwegian oldgrowth and managed Norway spruce stands. Some studies have suggested that the fungal ¯ora of conifer stands is often less diverse than that of broadleaved stands. For example, Villeneuve et al. (1989) found that the diversity of both ectomycorrhizal and saprotroph species in Quebec forests was signi®cantly lower in conifer stands than in deciduous stands, owing mainly to the scarcity of saprotrophs in conifer mor humus. In addition, Newton and Haigh (1998) in their study of ectomycorrhizal fungi in the UK, found that exotic conifer species displayed a lower mycorrhizal diversity than would be expected from their distributional areas. In the current study no di€erences in species numbers were recorded between crop types, but there were large di€erences in species composition, with spruce and pine sharing more species in common than with oak. Some signi®cant correlations were recorded between fungal community parameters and environmental variables. In such studies, there is always the danger that environmental variables can be inter-correlated, and in some cases the variable or variables which are responsible for di€erences in fungal community composition and diversity may not have been measured directly (see for example SaÊstad, 1995). However, a number of studies have suggested that fungal diversity may be related to soil and vegetation variables (Burova, 1974; Villeneuve et al., 1989). Here, number of mycorrhizal species was positively correlated with ground and ®eld layer vegetation diversity. Similar results were recorded by Ferris et al. (in press) in conifer plantation forests in lowland England, where a positive relationship between mycorrhizal diversity and number of tree species was recorded, and attributed to patterns of host speci®city demonstrated by the fungi involved (Newton and Haigh, 1998). Host tree diversity also emerged as an in¯uential variable in the correspondence analysis undertaken in the current investigation. The positive correlations between fungal and bryophyte diversity recorded here are less readily explicable; in boreal and temperate forests the host species of ectomycorrhizal fungi are usually tree species or vascular plants (Villeneuve et al., 1989;


J.W. Humphrey et al. / Biological Conservation 96 (2000) 241±252

SaÊstad, 1995). It is conceivable that both bryophytes and fungal sporocarps tend to be associated with relatively moist microsites. Fresh fallen deadwood volume was correlated with wood saprotroph species-richness thus supporting the ®ndings of Ferris et al. (in press) from Norway spruce and Scots pine plots in southern England. The ``fresh'' decay category was an amalgam of the three decay categories used by Hunter (1990): 1 Ð no decay, 2 Ð

less than 50% bark loss (initial bark loss), and 3 Ð more than 50% bark loss (advanced bark loss). In a number of studies in boreal forests, speci®c correlations have been recorded between wood saprotroph diversity and deadwood with initial and advanced bark loss (e.g. Crites and Dale, 1998; Lindblad, 1998; Kruys et al., 1999). In Swedish Norway spruce forests, a number of Red Listed fungi showed strong preference for fallen deadwood with well-rotted bark (Kruys et al., 1999).

Table 4 Threatened fungal species recorded during the surveya Name

Statusa,b (after Ing, 1992)

Statusb (after Anon, 1998c)

Plots where recorded

Functional group

Collybia acervata Cortinarius camphoratus Cortinarius laniger Cortinarius limonius Cortinarius purpurascens Cortinarius violaceus Fayodia gracilipes Hydnellum peckii Lactarius musteus Mycena purpureofusca Mycena rosella Mycena rubromarginata




Mycena urania Pseudocraterellus sinuosus Rozites caperata Russula declorans Sarcodon imbricatus Suillus ¯avidus Xeromphalina campanella


Vu LRnt LRnt LRlc Vu LRnt DD

2.4 6.2, 6.3 6.2 1.2, 2.4, 6.2, 6.3, 6.4, 1.3 5.2 5.2, 6.2 2.2 1.2, 2.1, 2.2 2.1, 10.2 1.1, 1.3, 1.4, 2.4, 6.2 1.4, 2.3, 5.1, 5.2, 5.3, 5.4, 6.1, 6.4, 10.1, 10.4 2.3 16.3 1.3, 1.4, 2.4 1.3, 1.4 2.2 1.4 2.1



Status refers to the IUCN categories of threat, as employed in Red Data lists (see Mace and Lande, 1991). Abbreviations (following the source documents from which they are taken): E, Endangered; V, Vulnerable; R, Rare.; LRnt, Lower risk near threatened; LRlc, Lower Risk least concern; DD, data de®cient; W, wood saprotrophs; M, mycorrhizal species; L, litter saprotrophs; P, parasites. b

Fig. 6. Records of threatened fungi in relation to distance of recording site from extant native pinewood area.

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Deadwood in more advanced states of decay, where the wood itself starts to rot, does not appear to support such diverse saprotroph communities (Crites and Dale, 1998; Kruys et al., 1999), although some speci®c fungi show a preference for this stage (Lindblad, 1998). The ®rst axis (FUNGI 1) of the fungi correspondence analysis (di€erences between the fungal ¯ora of pine and the other two crop types) was positively correlated with litter depth, accumulated temperature and moisture de®cit. Pine plots were generally cooler and drier and had shallower leaf litter. These results support those of Burova (1974) who found that climatic variables in¯uenced fungal community composition. Di€erences between oak and spruce plots (FUNGI 2) were related to tree species-richness and ground vegetation diversity, oak having higher values for the latter two variables. The fungal ¯ora of the foothill spruce plots was also distinct from that of the upland spruce plots (FUNGI 3). The results suggest that this may be caused by di€erences in soil properties and vegetation structure (vertical cover in the ®eld and shrub layer). Nantel and Neumann (1992) found a close correlation between basidiomycete community composition and soil variables, but there are contrasting studies. For example, Goodman and Trofymow (1998) found no e€ect of soil chemistry on the composition of mycorrhizal communities. Successional trends in ectomycorrhizal community composition have been characterised in relation to forest growth stage (Newton, 1992). In upland Britain, young forests are colonised by pioneer mycorrhizal fungi, but ``late-successional'' specialists are often restricted to older stands (see Flynn et al., 1998). Similar trends have been recorded overseas, for example Crites and Dale (1998) found a signi®cant e€ect of stand age on woodinhabiting fungi in aspen forest in Alberta. In this present study, distinct di€erences in the mycorrhizal ¯ora of the four stand ages were recorded. The over-mature stage was characterised by having more than 40 species not found in the other stages the majority of which Newton (1992) classed as ``late-successional''. Flynn et al. (1998) working in mixed stands of Sitka and Norway spruce in southern Scotland found that these late-successional species were maintained by management under an irregular silvicultural system (sensu Mathews, 1992) where some canopy cover was retained through the felling and restocking cycle. 4.2. Fungi of conservation importance One of the most striking features of the survey results was the extensive new records for rare and threatened fungi, including one species not previously recorded the UK (Panellus violaceofulvus and one not previously con®rmed as being native (Cortinarius callisteus, Orton, 1987). In addition, a further 19 species are considered to be threatened with extinction, in that they have been


listed on the provisional Red Data list for British fungi. These species were ®rst classi®ed according to the IUCN categories of threat by Ing (1992). Following revision of the criteria for classi®cation (Mace and Lande, 1991), the status of some of these species was tentatively reassessed by Scottish Natural Heritage (Anon, 1998c). The result is a more conservative assessment in many cases; for example seven species formerly considered `Vulnerable' or `Rare' are now classi®ed as `Data De®cient' (Table 4). In one sense, this emphasises the lack of information available to assess accurately the conservation status of fungal species. However, it is also clear that plantation forests provide suitable habitat for a number of these threatened species. The fact that eight species were recorded in sprucedominated plots in Clunes and Knapdale (Table 4) was unexpected and suggests a possible ability of these fungi to ``host-shift'' (Watling, 1995), as all are naturally associated with native pine forests. Information on the distribution of native pinewoods [taken from the Forestry Commission's Pinewood Inventory, Tuley (1995)] indicates that the spruce plots in Clunes forests are, on average, 5 km from existing pinewoods in the Loch Arkaig area. There may even have been fragments of native pine on rocky outcrops adjacent to plot 6.3, before the crop was established in the 1920s and 1930s (Hamilton, 1995). Thus temporal continuity of pinewood habitat within the locality could help explain the relatively large number of rare species records from Clunes forest. However, the Knapdale spruce plots are over 50 km from native pinewoods, the nearest group being in Glen Orchy and Glen Strae. It seems remarkable that these plots could have acquired pinewood fungi in the short time since planting (again in the 1920s and 1930s) and suggests that long-distance dispersal is possible in such species. There were few rare fungi recorded in Glentress and none in Kielder, re¯ecting perhaps the increasing distance of these sites from the nearest native pinewood fragment in Glen Falloch (Argyll). This theory is supported by Ferris et al. (in press) who did not record any native pinewood fungi in their survey of the mycota of lowland Scots pine, Corsican pine (Pinus nigra var maritima) and Norway spruce stands in England. However, the ecological characteristics of these fungi are poorly understood, particularly with respect to their ability to disperse and colonise new habitats. It is conceivable that some species may be translocated during forestry operations, for example in association with the root systems of planting stock. The occurrence of Suillus ¯avidus in conifer plantations in Shetland, which were established from planting stock raised in areas near to native pinewoods, suggests that this species may have been translocated in this manner (Watling 1992). Such a dispersal mechanism may also account for records of this taxon in southern England (Dickson and Leonard, 1996).


J.W. Humphrey et al. / Biological Conservation 96 (2000) 241±252

Most of the Red List species recorded were mycorrhizal; few were wood saprotrophs. This contrasts with ®ndings from Swedish boreal forests where a high proportion of Red listed fungi are wood inhabiting species (Rydin et al., 1997). The lack of threatened species in the oak plots is consistent with the suggestion by Watling (1997) that the oak woodlands of Scotland are not associated with a particularly distinctive mycota. 4.3. Conclusions and management recommendations This study yielded two main ®ndings. Firstly, although there were no e€ects of stand age or crop type on fungal species-richness, there were large di€erences in species composition between stands. The late-successional fungal assemblages associated with the over-mature growth stage in both pine and spruce need to be maintained through some form of continuous cover management system as clear-felling is inimical to these species (Westerlund, 1989; Hagerman et al., 1999). Although some fungi can survive for long periods in mycelial form, extensive felling of host trees can disrupt mycelial connectivity and reduce the extent of re-colonisation after restocking (Flynn et al., 1998). Even smaller-scale felling can be disruptive (Flynn et al., 1998), and it may be more desirable, therefore, to identify older stands known to have a high number of characteristic species as potential biological (non-intervention) retentions to allow the maturation of individual trees, accumulation of deadwood, and diversi®cation of the tree ¯ora. This would foster the development of a `reservoir' of species for colonisation of neighbouring stands managed by the patch clear-fell systems which are currently prevalent in upland Britain. The second main ®nding of the study was the unexpectedly high incidence of rare and threatened fungi in plantation stands of pine and spruce. This highlights the positive contribution that plantations of both native and exotic conifers can make to the conservation of native fungi. Plantations near extant native pinewoods seem more likely to acquire pinewood fungi, but site history may also play an important role, with small fragments of remnant habitat acting as a source of inoculum for future stands. Therefore, to help conserve native fungi, strategies for the future management of commercial conifer forests in northern Britain should consider their potential habitat value, which could be simply de®ned on the basis of proximity to native pinewoods. The Forest Habitat Network model of Peterken et al. (1995) provides an appropriate mechanism for the successful integration of plantations and native woodlands, and plans are in place to develop these principles further in a native pinewood area (Ratcli€e et al., 1998). Such approaches could encourage the rapid colonisation of newly established native pinewoods by native fungi.

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