The vegetation communities of unmanaged aquatic buffer zones within conifer plantations in Ireland

The vegetation communities of unmanaged aquatic buffer zones within conifer plantations in Ireland

Forest Ecology and Management 353 (2015) 59–66 Contents lists available at ScienceDirect Forest Ecology and Management journal homepage: www.elsevie...

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Forest Ecology and Management 353 (2015) 59–66

Contents lists available at ScienceDirect

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

The vegetation communities of unmanaged aquatic buffer zones within conifer plantations in Ireland Cormac Mc Conigley a,⇑, Heather Lally a,b, Mark O’Callaghan a, Philip O’Dea c, Declan Little d, Mary Kelly-Quinn a a

School of Biology and Environmental Science, University College Dublin, Belfield, Dublin 4, Ireland Marine and Freshwater Research Centre, Galway-Mayo Institute of Technology, Dublin Road, Galway City, Ireland Coillte, Church Road, Newtownmountkennedy, Co Wicklow, Ireland d Woodlands of Ireland, Seismograph House, Rathfarnham Castle, Dublin 14, Ireland b c

a r t i c l e

i n f o

Article history: Received 19 December 2014 Received in revised form 8 May 2015 Accepted 10 May 2015 Available online 29 May 2015 Keywords: Conifer plantations Forest management Riparian zones Aquatic buffer zones Vegetation communities

a b s t r a c t The riparian zone is the interface between aquatic and terrestrial habitats and forms the ecotone where the two ecosystems intersect. Areas of the riparian zone utilised for the protection of water quality are common today, and are either left undisturbed or managed to intercept or modify impacts from adjacent land uses. In Ireland, aquatic buffer zones (ABZs) are used to protect streams from the potential impacts of commercially managed conifer forests and associated high impact forestry operations. In 1991, the Forest Service (currently of the Department of Agriculture, Food and the Marine) introduced the requirement for ABZs to be put in place on all streams identified on Ordinance Survey maps at either afforestation or restocking after clearfell. The width of the ABZs range from 10–25 m depending on the slope of the river bank in combination with the susceptibility of the soils to erode. Current practice is to leave the ABZ undisturbed allowing for natural colonization by a mix of species and establishment of various habitat types. This study describes the habitats and vegetation composition of 86 naturally vegetated riparian zones (65 ABZs in commercial conifer forests and 21 control sites) on six soil types, with a view to informing their optimum management. Across all sites, 392 taxa, within 32 habitat types were identified. The most common habitats were wet grasslands and scrub. Little variation was noted between the structure and composition of plant communities in ABZs (on afforested and clearfell & replanted) and control sites within a soil type. The communities did differ across soil types between the mineral and peaty soils, which were independent of the forest type. Within a soil type, ABZs are maintaining similar habitat and species diversity to that found on control sites indicating that current forest management practices are not impacting plant diversity in the ABZs. It is noted that tree species are not a feature of the riparian zone on peat soils and thus tree planting is not recommended as a management option unless used to control water temperatures. There is scope for tree planting on mineral soils, as control sites contained woodland habitats which were absent from the ABZs of clearfell and replanted sites. Ó 2015 Elsevier B.V. All rights reserved.

1. Introduction Exotic conifer plantations, a widespread human activity, affect the hydrochemistry (Kelly-Quinn et al., 2008; Feeley et al., 2013), structure and functioning (Riipinen et al., 2010; Martínez et al., 2013) of streams and rivers that they border. However, it has been reported that the riparian vegetation can mitigate some of these effects (Quinn et al., 2004) and therefore vegetated riparian buffer

⇑ Corresponding author. E-mail address: [email protected] (C. Mc Conigley). http://dx.doi.org/10.1016/j.foreco.2015.05.009 0378-1127/Ó 2015 Elsevier B.V. All rights reserved.

zones are commonly used for this purpose (see Richardson et al., 2012). The riparian zone is the interface between the aquatic and terrestrial ecosystems. In the case of rivers, it is generally considered to include the bank and the portion of land influenced by river water during flooding (Gregory et al., 1991; Naiman and Décamps, 1997; Little et al., 2008). It has long been understood that vegetation in the riparian zone is an important factor in maintaining the health and condition of rivers (Vannote et al., 1980; Minshall et al., 1983). Shade cast by plants in the riparian zone reduce maximum stream temperatures (Caissie, 2006; Broadmeadow et al., 2011; Ryan et al., 2013), while their stems and root systems intercept sediment by slowing over-land flow,

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increasing water absorption rates in riverbank soils and controlling erosion by stabilising banks (Castelle et al., 1994). Riparian vegetation also contributes to allochthonous inputs from the catchment, which form the primary source of course and fine particle material that fuel aquatic macroinvertebrate food webs in low order streams (Tank et al., 2010). Furthermore riparian vegetation influences the contribution of terrestrial invertebrates which can form an important part of the diet of fish (Kelly-Quinn and Bracken, 1990; Nakano et al., 1999; Ryan and Kelly-Quinn, 2014). Ireland is one of the least forested countries in Europe (MCPEE Liaison Unit Warsaw, 2007). At the beginning of the 20th century, forest cover decreased to approximately 1% (Rackham, 1986). Following the foundation of the state in 1922 the government sponsored a reforestation program focusing on marginal and upland areas unsuitable for agriculture (Giller and O’Halloran, 2004). Forest cover has steadily increased in the interim and currently stands at 10.5% (Forest Service, 2014). The majority of the forests planted comprised of fast growing non-native conifer species particularly Picea sitchensis Bongard (Sitka spruce), which form over half of the total forest estate in Ireland (Forest Service, 2014). The installation of riparian buffer zones, known as aquatic buffer zones (ABZs) in Ireland, of between 10 and 25 m width along streams has been mandatory in Irish forestry since 1991 (Forest Service, 2000). The ABZs are put in place at afforestation or in older forests at replanting after clear-felling. The width of the ABZ is dependent on the slope of the area and the likelihood for soil erosion, with steeper sites on erodible soils requiring the widest buffers (up to 25 m). Within the ABZ, forestry operations are strictly curtailed, no conifers are planted, drains end before entering the ABZ, incursions of machinery is kept to a minimum and brash mats are used to protect the soil from erosion and compaction (Forest Service, 2000). These measures are intended to allow the ABZ to retain the plant communities present before the establishment of a conifer forest, thus allowing the ABZ to maintain the functions of riparian zones outlined above. The incorporation of ABZs into Irish conifer plantations provides an opportunity to investigate management strategies that can enhance the ecology of the stream (Little et al., 2008). Before that can be achieved, a thorough knowledge is required of the mosaic of habitats and plant communities present in ABZs. This is particularly important at this time as many plantations in Ireland reach maturity. Several studies examining the responses of upland plant communities to conifer afforestation have noted a loss of species richness (Hill, 1979) and convergence toward a community dominated by a few species (Sykes et al., 1989; Wallace et al., 1992). These changes largely relate to the degree of canopy closure, site type and forest management. These factors are not expected to be major issues for ABZs as they are not under the forest canopy (Anderson, 1979; Hill, 1979; Wallace et al., 1992; Wallace and Good, 1995). There is a paucity of research describing vegetation in riparian zones of conifer plantations in Ireland. However work carried out in the UK (Wallace and Good, 1995; Broadmeadow and Nisbet, 2002) and New Zealand (Boothroyd et al., 2004; Langer et al., 2008) indicate a community similar to that of native forest will occur in the understory. This study addresses these knowledge gaps by investigating the habitats and plant communities present in ABZs across Ireland. Aquatic buffer zones instated at initial forest planting (afforestation), and those put in place after clearfell & replanting are compared to natural riparian zones in catchments with no conifer plantations in the vicinity. Across the country 86 unmanaged riparian zones (65 ABZs in conifer plantations and 21 control sites) are described and compared across six soil types. This information provides insight into the current state of the ABZs in the national forest estate and will be used to inform their management into the future.

It was hypothesised that there would be a significant change in the community composition between ABZs and the controls sites with the control sites having greater vegetation and habitat diversity than ABZs on the same soil type. 2. Materials and method 2.1. Site selection Located in northwest Europe on the Atlantic Ocean, Ireland experiences an oceanic climate with mild winter and cool summers. There is considerable variation in rainfall, the west coast receiving 2000 mm in 250 days of rain annually while on the east coast 700 mm of rain falls in 190 days. Due to past deforestation and agriculture Ireland is largely dominated by improved grassland habitats with the most productive grass varieties chosen for cultivation to facilitate livestock production. If left to develop naturally the potential vegetation would be dominated by temperate woodlands (Cross, 1998, 2006), similar to much of western Europe (Bohn et al., 2003), but with lower diversity due to territory size, island status and glacial history. Peatland areas are dominated by species such as Eriophorum angustifolium and Molinia caerulea with dwarf shrubs Calluna vulgaris and Erica tetralix (for a detailed map of potential vegetation see Cross, 2006). Riparian zones in afforested, clearfell and replanted and control sites on six different soil types peat, peaty podzols, peaty gley, well-drained mineral (WDM), mineral gleys and mineral alluvium, were sampled, resulting in a total of 18 categories. Each ABZ site was required to have greater than 300 m length of conifer plantation on both river/stream banks, while control site were required to have no conifer plantations in the vicinity. Thus, on afforested and clearfell and replanted sites, the ABZ’s were within conifer plantations bordered by a significant forested area. For the control sites, the natural riparian zone that would form the ABZ if the area was afforested was surveyed. All of the sites (both within the forest and controls) were unmanaged and left to be colonised by native vegetation and develop naturally. A total of 86 sites were selected across the country (Table 1). The numbers of sites in some categories were low due to difficulty in finding sites matching the criteria (clearfell & replanted sites on peaty gley soil  2 and afforested sites on well drained mineral (WDM) & mineral gley soils  2). 2.2. Vegetation survey Within each ABZ, three sampling stations (10 m  10 m) were established 100 m apart, on both the left and right banks (Fig. 1), creating a total of six stations. Within each station, three relevés (2 m  2 m), 2 m apart, were sampled along a transect perpendicular to the stream. All species present within the relevés were

Table 1 Number of sites selected for vegetation surveys classified according to 18 categories (6 soil types vs. 3 forest types). Afforested Peat Peaty podzol Peaty gley WDM Mineral gley Mineral alluvium Total per forest type

Clearfell & replanted

Control

7 4 3 2 2 3

10 6 2 6 3 4

8 5 4 7 4 6

21

31

34

WDM: well drained mineral.

Total per soil 25 15 9 15 9 13

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Fig. 1. Schematic illustrating the location of vegetation survey stations along the left and right stream banks within the ABZ (grey area) with the arrangement of relevés also illustrated for each 10 m sampling station. The ABZ for afforested and clearfell & replanted sites is within the forest.

recorded using percentage cover. Species were identified to their lowest taxonomic level, where feasible. Bryophyte samples were returned to the laboratory for further identification. Percentage cover of bryophytes, Sphagnum and liverworts were recorded with presence/absence data only noted for individual species. Relevés were only taken within the ABZ, thus if the buffer zone was less than 10 m wide fewer relevés were taken in each station. Using percentage cover, rather than the more commonly used Braun-Blanquet’s abundance/dominance or Domin scale allowed for application of arithmetic operations on the data that would be invalid in ordinal scales (Podani, 2006). In the 100 m between each station, habitats present on each bank were identified, coded and recorded according to Fossitt (2000). Shannon diversity index measurements were used to compare diversity within the relevés across all categories (Shannon and Weaver, 1963). 2.3. Soil analyses Soil samples were collected from the centre of each relevé to a depth of 20 cm using a hand corer and returned to the laboratory for analysis. All soil samples were stored at 4 °C until they were analysed. Soil pH was determined by creating a suspension of fresh soil in deionised water (ratio 1:1 by volume) and measuring the pH of the resulting slurry with a glass electrode pH meter (Allen et al., 1974). To determine soil moisture content, samples were dried in an oven at 105 °C for 24 h to a constant weight on successive weighings. Organic matter content of the soil was calculated from loss on ignition (%LOS) of 1 g of dry soil heated to 550 °C for two hours and determining mass lost.

null hypotheses of no difference to be rejected. PERMANOVA is a permutational multivariate analysis of variance, analogues to a multi factorial ANOVA but calculated through permutations. The routine first calculates the F statistic, then randomly reassigns the labels and recalculates the F statistic with the new label arrangement. The p value can be calculated based on the number of these permutations with an equal or larger F value than the original data. The rationale being that if the rearranging of the labels at random produces results as significant or more significant than the true arrangement then the true arrangement is random. PERMANOVA has the advantage of being non-parametric and is robust in its analysis of unbalanced designs. The analysis can be run on any similarity or dissimilarity matrix. In this study, species percentage cover for each relevé was used to calculate Bray-Curtis dissimilarities on which PERMANOVA was run. An advantage of this measure of dissimilarity is that it is not affected by joint absences, a common occurrence in larger ecological datasets. No pre-treatment was applied to the data in advance of analysis except reverting species back to genus level where a large percentage (>20%) were not identified to species level. In categories where over 80% were identified, the unidentified portion were split proportionality among species in the genus that were found on that site. Similarity Percentage (SIMPER) analysis was carried out to identify the key taxa contributing to the significant differences between the 18 categories which were identified by PERMANOVA. SIMPER returns the mean abundance of each taxon in the categories under comparison and their contribution to the differences observed as a percentage of the total dissimilarity. Shannon diversity indices were also used to compare diversity within the relevés across all categories.

2.4. Data analyses 3. Results Statistical analysis was carried out using PERMANOVA to test for significant differences in taxon richness, community structure, and cover of habitat types between the 18 categories. For univariate analyses, data were tested for homogeneity of variance and transformed, where necessary. Where transformation could not equalise variance, a reduced p value of 0.01 was required for the

3.1. Habitat composition A total of 32 habitats types, as defined by Fossitt (2000), were identified across all sites (Table 2). Wet grassland dominated by M. caerulea (GS4), dense bracken (HD1), and wet and dry heaths

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(HH1&3) were more commonly encountered in the ABZs and control sites on the three peaty soil types. Wet and dry heaths were also absent from the control sites on the on mineral soils. On the majority of soils, scrub habitats were more common in ABZs than the control sites, generally scrub was dominated by Salix spp., Ulex europaeus or Betula spp. (WS1) (Table 2). On the three mineral soils, several oak woodland types (WN1 & WN2) and mixed woodland types (WD1, WD2 & WD3), were generally present only on the control sites.

Table 3 Shannon diversity index values (mean ± standard error) across the 18 categories.

3.2. Diversity measures

soils generally had higher mean taxon richness than those on peaty soils (Fig. 2). Mean taxon richness varied only slightly from a low of 7.06 (afforested peat) to a maximum of 11.07 (afforested mineral gley), a value just marginally higher than on afforested, WDM soils (10.82). Analysis detected a significant interaction between soil and forest type (F10,1159, p = 0.0118). Significantly higher numbers of species were recorded on peat control sites than on either afforested (t = 4.31, p = 0.002) or clearfell & replanted (t = 3.66, p = 0.004) sites (Fig. 2). The opposite occurred on mineral gleys, with control sites recording a significantly lower numbers of species than ABZs (t = 4.47, p = 0.0136) (Fig. 2). No other comparisons within a soil type were significant. On each soil type, a number of taxa were unique to one forest type (Table 4). Control sites on well drained mineral and mineral alluvium soils had the highest number of unique taxa, 48 and 46, respectively, this is in part due to the relatively high percentage of sites on the soil type that are controls, near 50%. In most cases the largest number of unique taxa was directly related to the greatest number of sites surveyed.

Peat Peaty podzol Peaty gley Well drained mineral Mineral gley Mineral alluvium

Clearfell and replanted sites on mineral gleys had the highest mean Shannon diversity (1.82), while control sites on peaty gley had the lowest (1.33) (Table 3). There was a marginally significant interaction between soil and forest type (F10,1159, p = 0.0434). Pairwise comparison found that control sites on peat soil had significantly higher diversity than afforested (t = 3.7947, p = 0.0032) or clearfell & replanted (t = 2.7587, p = 0.0142) sites. There were no significant differences in diversity between forest types on other soil types.

3.3. Plant community composition Across all sites, 392 taxa were identified, Potentilla erecta and M. caerulea were the most commonly encountered species present in 43.2% and 43.1%, of relevés, respectively. These were followed by Juncus effusus, Pteridium aquilinum and Rubus fruticosus agg. occurring on 29.8%, 29.3% and 29.2% respectively. The ABZs on mineral

Afforested

Clearfell & replanted

Control

1.34 ± 0.04 1.71 ± 0.06 1.44 ± 0.09 1.67 ± 0.10 1.67 ± 0.14 1.59 ± 0.11

1.48 ± 0.03 1.62 ± 0.05 1.39 ± 0.07 1.42 ± 0.08 1.82 ± 0.06 1.73 ± 0.06

1.72 ± 0.04 1.63 ± 0.04 1.33 ± 0.06 1.54 ± 0.04 1.51 ± 0.05 1.61 ± 0.05

Table 2 Percentage of sites within each of the 18 categories, where habitats were identified during the survey. Fossitt (2000) habitat names

Codes

Stone wall Exposed sand Recolonising bare soil Exposed rock Reed and large sedge swamp Tall-herb swamps Improved agricultural grassland Marsh Dry calcareous and neutral grassland Dry meadows and grassy verges Dry-humid acid grassland Wet grassland Dense bracken Dry siliceous heath Wet heath Upland blanket bog Lowland blanket bog Rich fen and flush Poor fen and flush Broadleaved woodland (Mixed) Mixed broadleaved/conifer wood Conifer woodland (Mixed) Conifer plantation Hedgerows Treelines Oak-birch-holly woodland Oak-ash-hazel woodland Riparian woodland Wet willow-alder-ash woodland Scrub Short rotation coppice Recently-felled woodland

BL1 ED1 ED3 ER FS1 FS2 GA1 GM1 GS1 GS2 GS3 GS4 HD1 HH1 HH3 PB2 PB3 PF1 PF2 WD1 WD2 WD3 WD4 WL1 WL2 WN1 WN2 WN5 WN6 WS1 WS4 WS5

Peat A

Total number of habitats across case

Peaty podzol R

C

A

R

Peaty gley C

A

R

WDM C

A

Min. gley R

C

A

Min. alluvium

R

C

14 17 17 33

10 30 10 10

25 13

33

25 25

25

17

13 33

90 20 70 40

13 100 25 50 50 38 13 13

100 75 50 25

100 67 33 67

100 67 67 67

100

50 100

67 33 33

50 25

67 33 33 17

29

67

29 29

100 33 33 33

50

25 50

50

50

67

67

17

33 10

25

25

50 50

67 25 25 75

20

12

11

25

43

50 25

67

14

50

25 100

100

50 100

67

25 25

4

8

8

100

50

50

100

100

8

8

8

6

10

5

12

13

17

25

33

33 67 33 67 33

17 17

17 50 50

33

50 17 33 50

25 50

33 33

13 14

17

25

17

14 29 57 43 14 71

25 50 50

33

75 50 25 25

33

50 30

17

17 33

50

C

33

17 17 100 75 50 50 25

R

75 25

25 29

83 33 100 83 17 17

A 33

75

67

25 8

WDM = Well Drained Mineral; A = Afforested; R = Clearfell & replanted; C = Control; Min. = Mineral.

9

12

13

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Fig. 2. Taxon richness (mean ± standard error) across soil and forest types. A = afforested R = clearfell & replanted and C = control.

Table 4 The number of taxa that were exclusive to the respective forest type across the six soil types with number of sites surveyed in brackets. Soil

Afforested

Clearfell & replanted

Control

Exclusive to ABZs

Peat Peaty podzol Peaty gley WDM Mineral gley Mineral alluvium

6 (7) 19 (4) 36 (3) 8 (2) 9 (2) 11 (3)

19 34 11 31 39 26

31 18 43 48 32 46

34 53 47 39 48 37

(10) (6) (2) (6) (3) (4)

3.4. Community structure The PERMANOVA test on the community structure (based on mean station values) identified a significant interaction between soil and forest type (F10,377 = 1.607, p = 0.0008). Pairwise comparison found significant differences between the communities on afforested and clearfell and replanted sites on peaty gleys (t = 1.78, p = 0.02) and mineral gleys (t = 1.84, p = 0.007). Communities on control and clearfell and replanted sites on mineral gleys and mineral alluvium soils also differed significantly (t = 1.81, p = 0.04) and (t = 1.41, p = 0.05) respectively. Comparisons of forest types across the range of soil types surveyed indicated that, the primary differences were between the communities on mineral soils and peaty soils (Fig. 3). The community differences by soil type can be clearly visualised in the MDS plot (Fig. 3), ABZs on mineral soils (red symbols) are grouped together on the right, while the peat soils (green symbols) are to the left of the graph. Forest type, represented by the differing symbol shapes, has no effect on the position of a point in the ordination. Soil pH and percentage loss on ignition (% LOS) affected plant community structure. Peaty soils had lower pH values (4.48 ± 0.02) and higher % LOS (39.2% ± 2.1) than the mineral soils

(8) (5) (4) (7) (4) (6)

(17) (10) (5) (8) (5) (7)

(pH 5.04 ± 0.01; % LOS 14.3 ± 1.1) which was independent of forest type (Fig. 3). Similarity percentage (SIMPER) analysis of comparisons between vegetation on peaty and mineral soils found that the taxa primarily responsible for the dissimilarity were M. caerulea and Bryophyta spp. On average M. caerulea accounted for 21% of the cover on sites on peaty soils, while it made up only 4% of the cover on mineral soils. Bryophyta spp. (excluding Sphagnum spp.), R. fruticosus and Luzula sylvatica were next most important, all were more frequent on mineral soils. 3.5. Buffer widths Across all soils one third of sites had a mean ABZ width below the minimum required (10 m) while three quarters of ABZs were below the minimum in at least one station (Table 5). 4. Discussion This study investigated the habitats and vegetation composition of 86 unmanaged ABZs (65 in forests and 21 control sites). As far as can be ascertained, this is the first study of its kind in Ireland.

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Fig. 3. MDS of mean percentage cover of all taxa per site with environmental variables (pH and% loss of ignition) overlaid. Soil types are represented by shapes while the forest types are denoted by their colour.

Table 5 Percentage of sites on which either the mean width (across 6 stations) or the minimum width on at least one station was below the minimum guidelines of 10 m. Forest type

n

Percentage of sites with mean ABZ width <10 m

Percentage of sites with a ABZ width <10 m in at least one station

All Afforested Clearfell & replanted

52 21 31

34.6% 28.6% 38.7%

76.9% 61.9% 87.1%

Similar to the findings of Wallace et al. (1992) wet grasslands were the most common habitat type found in the riparian zone. In the current study, these were, in most cases, dominated by M. caerulea or Juncus species. Scrub habitats dominated by Salix spp., Betula spp. or U. europaeus were also common though more so in the ABZs of afforested, and clearfell and replanted sites than on control sites. This may be due to the conifer plantations providing shelter and allowing for scrub growth to establish. These scrub species are also typically common understory trees in native Irish woodlands. On mineral soils oak woodlands were present only in controls sites. This may be a feature of the age of the ABZs, having only been established in the last 24 years there may not have been sufficient time for oak-dominated woodlands to develop (Bose et al., 2014). In addition, oak would not be expected to be a dominant species in riparian woodlands where frequent flooding and inundated soils persist. It may also reflect the paucity of native woodlands in Ireland generally (Little et al., 2008); particularly ‘old’ riparian woodlands where Quercus robur (pedunculate oak) in particular, may be encountered on the drier margins of riparian woodlands (Little et al., 2008). The short rotation time of spruce plantations may further limit the ability of slow growing species that require the stable environment of a woodland to mature (Brockerhoff et al., 2003), inclusive of successional stages. The community composition of the ABZs was similar to those identified in the riparian zones in forested catchments in the UK (Wallace et al., 1992; Broadmeadow and Nisbet, 2002). While afforestation with conifers can negatively affect the diversity of ground vegetation (Hill, 1979) this was not supported by the findings of the current study where few differences in the plant community of control sites and the ABZs were noted. Species typical of peatland habitats such as those identified by Perrin et al. (2009) dominated in all categories on peaty soils. The deleterious

effect of conifer afforestation on plant communities are associated with canopy closure and increased shading. If sufficient light penetrates the canopy, conifer afforestation may add to overall diversity (Ferris et al., 2000; Humphrey et al., 2002) and the community may evolve toward one more typical of the understory of native forests (Brockerhoff et al., 2003; Langer et al., 2008; Zhang et al., 2010). Since the ABZs are outside the plantations and do not experience a significant reduction in light levels, this almost certainly accounts for the lack of variation in the communities between forest types found in this study. On each soil type, there were taxa unique to the controls sites, however at the same time, there were species that were only found in the ABZs and missing from the controls sites. Overall, taxa missing from ABZs on all soil types were rare and infrequent on control sites, and no consistent trends in the attributes of the species unique to the control sites were identified. The key finding of the current study suggests that it is soil type which primarily dictates the vegetation communities present in the riparian zone of controls sites and in the established ABZs. There were two distinct communities, those on mineral soils and those on peaty soils. The importance of soil in determining the plant community is not unexpected. Variation in soil substrate largely explained species composition in a study of heathland and grassland habitat types in the Netherlands (Roem and Berendse, 2000). Increased availability of nutrients alters community structure and is associated with a loss of species diversity in grasslands (Berendse and Elberse, 1990). Increasing nutrient loading increases plant production resulting in the selection pressure toward taller, fast growing species (Grime, 1979) and a resultant rank sward. The pH of soil determines the availability of important nutrients; acidic soils (peaty) often have lower concentrations of nitrogen (N) and phosphorus (P) available for plant uptake (Lucas and

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Davis, 1961). Increased deposition of N in recent years increases soil acidity; in more acidic soils the availability of potassium (K) and P decrease. This increases the N:P and N:K ratios (Roem and Berendse, 2000), both of which are important determinants of plant community composition. Acid deposition would largely only influence peaty soils as mineral soils have a greater capacity to buffer this acidity. Another key difference between the mineral and peaty soils is soil water availability. Peaty soils are often continuously waterlogged while mineral soils are generally only periodically waterlogged in riparian zones. Waterlogged soils contain localised areas with little or no available oxygen or nitrogen, potentially causing root death (Patrick and Wyatt, 1964). Small scale variations in water availability, either a deficit or surplus, can affect the community present (see review by Silvertown et al., 2015). Soil water balance can be a powerful predictor of the local community composition (Piedallu et al., 2013). It is the combined effect of soil wetness, pH, organic content and nutrient availability that influence the species that will grow in the riparian zone. In the current study the measured pH, moisture content and organic matter content differed between the mineral and peaty soil types, irrespective of forest types. One or all of these variables in combination could be responsible for the difference in community observed. In general, on peaty soils in the present study wet grasslands, heath habitats and various bog type habitats occurred while on mineral soils mixed woodland habitats were common especially on control site. In at least one station, a high percentage of the ABZs surveyed had widths below the minimum required. Width was measured from the tree stem to the stream edge, and in effect, the ABZ was shaded from each side for an average of 2.8 m by overhanging branches. As the conifers adjacent the ABZ grow, their branches will further extend into the ABZ reducing the unshaded width. In particular, it is recommended, where possible, that ABZs are widened to the minimum required and beyond, especially in high water status catchments and freshwater pearl mussel catchments. 6. Conclusions and implications for forest management This study found little difference between community structure, composition or species richness in control sites and ABZs. Thus there is little need to intervene intensively in ABZs with regard to management at establishment or after the felling of the first conifer rotation if maintaining the natural plant community is the objective. However, ABZ widths were found to be below the minimum required in many cases and this should be addressed in order to maintain optimal ABZ functioning. In addition, there was a lack of riparian woodland habitats in ABZs on mineral soils and selective planting of trees in up to 20% of the ABZ area (as described in Little et al., 2008) on sites that are isolated from native tree and shrub seed sources, or in specific locations within high water status and Margaritifera margaritifera (FPM) catchments is recommended. Tree species were not a feature on peaty soils and the planting of trees on this soil type is not advised unless to ameliorate projected elevated water temperatures in the future due to climate change, or for enhancement of instream production through increased leaf litter input. Finally, long-term monitoring of a series of ABZ sites is recommended to track changes in vegetation communities during subsequent forest rotations. Acknowledgements This project was funded by COFORD, the Department of Agriculture, Food and the Marine. We acknowledge Woodlands

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of Ireland and Coillte, Technical Services for their assistance and technical input into the study as well as Coillte forest management for facilitating site access and Dr. Shawn McCourt for help with the survey work carried out. References Allen, S.E., Grimshaw, H., Parkinson, J.A., Quarmby, C., 1974. Chemical Analysis of Ecological Materials. Blackwell Scientific Publications. Anderson, M., 1979. The development of plant habitats under exotic forest crops. In: Wright, S.E., Buckley, G.P. (Eds.), Ecology and Design in Amenity Land Management. Wye Recreation Ecology Research Group, Wye College Ashford, pp. 87–109. Berendse, W.T., Elberse, F., 1990. Competition and nutrient availability in heathland and grassland ecosystems. In: Grace, J., Tilman, D. (Eds.), Perspectives on Plant Competition. Academic Press, San Diego, CA, pp. 93–116. Bohn, U., Neuhäusl, R., Gollub, G., Hettwer, C., Neuhäuslová, Z., Raus, T., Schluter, H., Weber, H., 2003. Karte der natürlichen Vegetation Europas/Map of the natural vegetation of Europe. Maßstab/Scale 1: 2,500,000. Münster: Bundesamt für Naturschutz. Boothroyd, I.K.G., Quinn, J.M., Smith, B.J., Langer, E.R., Costley, K.J., Steward, G., 2004. Riparian buffers mitigate effects of pine plantation logging on New Zealand streams 1. Riparian vegetation structure, stream geomorphology and periphyton. For. Ecol. Manage. 194, 199–213. Bose, A.K., Schelhaas, M.J., Mazerolle, M.J., Bongers, F., 2014. Temperate forest development during secondary succession: effects of soil, dominant species and management. Eur. J. For. Res. 133, 511–523. Broadmeadow, S., Nisbet, T.R., 2002. The Effect of Riparian Forest Management on the Freshwater Environment. In: SNIFFER 2002. Broadmeadow, S.B., Jones, J.G., Langford, T.E.L., Shaw, P.J., Nisbet, T.R., 2011. The influence of riparian shade on lowland stream water temperatures in southern England and their viability for brown trout. River Res. Appl. 237, 226–237. Brockerhoff, E.G., Ecroyd, C.E., Leckie, A.C., Kimberley, M.O., 2003. Diversity and succession of adventive and indigenous vascular understorey plants in Pinus radiata plantation forests in New Zealand. For. Ecol. Manage. 185, 307–326. Caissie, D., 2006. The thermal regime of rivers: a review. Freshwat. Biol. 51, 1389– 1406. Castelle, A.J., Johnson, A.W., Conolly, C., 1994. Wetland and stream buffer size requirements – a review. J. Environ. Qual. 23, 878–882. Cross, J.R., 1998. An outline and map of the potential natural vegetation of Ireland. Appl. Veg. Sci. 1, 241–252. Cross, J.R., 2006. The potential natural vegetation of Ireland. Biol. Environ.: Proc. R. Irish Acad. 106, 65–116. Feeley, H.B., Bruen, M., Blacklocke, S., Kelly-Quinn, M., 2013. A regional examination of episodic acidification response to reduced acidic deposition and the influence of plantation forests in Irish headwater streams. Sci. Total Environ. 443, 173– 183. Ferris, R., Peace, A.J., Humphrey, J.W., Broome, A.C., 2000. Relationships between vegetation, site type and stand structure in coniferous plantations in Britain. For. Ecol. Manage. 136, 35–51. Forest Service, 2000. Forestry and Water Quality Guidelines. Dublin: Department of the Marine and Natural Resources. Forest Service, 2014. Ireland’s Forests – Annual Statistics. Department of Agriculture, Fisheris and Marine, Forest Service, DAF, Johnstown Castle Estate, Co., Wexford. Fossitt, J.A., 2000. A Guide to Habitats in Ireland. The Heritage Council/Chomhairle Oidhreachta, Kilkenny. Giller, P.S., O’Halloran, J., 2004. Forestry and the aquatic environment: studies in an Irish context. Hydrol. Earth Syst. Sci. 8, 314–326. Gregory, S.V., Swanson, F.J., McKee, W.A., Cummins, K.W., 1991. An ecosystem perspective of riparian zones. Bioscience 41, 540–551. Grime, J.P., 1979. Plant Strategies and Vegetation Processes. John Wiley and Sons, Chichester, UK. Hill, M.O., 1979. The development of a flora in even-aged plantations. In: Ford, E.D., Malcolm, D.C., Atterson, J. (Eds.), The Ecology of Even Aged Forest Plantations. Institute of Terrestrial Ecology, Cambridge, pp. 175–192. Humphrey, J.W., Ferris, R., Jukes, M.R., Peace, A.J., 2002. The potential contribution of conifer plantations to the UK Biodiversity Action Plan. Botanic. J. Scot. 54, 49– 62. Kelly-Quinn, M., Cruikshanks, R., Johnson, J., Matson, R., Baars, J.-R., Bruen, M. Forestry and surface water acidification. Ireland: report to the Western River Basin District Working Group; 2008. p. 81. . Kelly-Quinn, M., Bracken, J.J., 1990. A seasonal analysis of the diet and feeding dynamics of brown trout, Salmo trutta L., in a small nursery stream. Aquacult. Res. 21, 107–124. Langer, E.R., Steward, G.a., Kimberley, M.O., 2008. Vegetation structure, composition and effect of pine plantation harvesting on riparian buffers in New Zealand. For. Ecol. Manage. 256, 949–957. Little, D., Collins, K., Cross, J., 2008. Native Riparian Woodlands – A Guide to Identification, Design, Establishment and Management. In, Native Woodland Scheme Information Note No. 4. Woodlands of Ireland. Dublin. Lucas, R., Davis, J., 1961. Relationships between ph values of organic soils and availabilities of 12 nutrients. Soil Sci. 92, 177–182.

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Martínez, A., Larrañaga, A., Pérez, J., Descals, E., Basaguren, A., Pozo, J., 2013. Effects of pine plantations on structural and functional attributes of forested streams. For. Ecol. Manage. 310, 147–155. MCPEE Liaison Unit Warsaw, 2007. State of Europe’s Forests 2007: The MCPFE Report on Sustainable Forest Management in Europe. In: Warsaw, Poland, United Nations Economic Commission for Europe & Food and Agriculture Organization of the United Nations. Minshall, G.W., Petersen, R.C., Cummins, K.W., Bott, T.L., Sedell, J.R., Cushing, C.E., Vannote, R.L., Monographs, E., 1983. Interbiome comparison of stream ecosystem dynamics. Ecol. Monogr. 53, 1–25. Naiman, R.J., Décamps, H., 1997. The ecology of interfaces: riparian zones. Annu. Rev. Ecol. Syst. 28, 621–658. Nakano, S., Miyasaka, H., Kuhara, N., 1999. Terrestrial-aquatic linkages: riparian arthropod inputs alter trophic cascades in a stream food web. Ecology 80, 2435– 2441. Patrick, W.H., Wyatt, R., 1964. Soil nitrogen loss as a result of alternate submergence and drying. Soil Sci. Soc. Am. J. 28, 647–653. Perrin, P.M., Roche, J.R., Barron, S.J., 2009. Scoping study and pilot survey for a national survey and conservation assessment of upland habitats and vegetation in Ireland. Piedallu, C., Gégout, J.C., Perez, V., Lebourgeois, F., 2013. Soil water balance performs better than climatic water variables in tree species distribution modelling. Global Ecol. Biogeogr. 22, 470–482. Podani, J., 2006. Braun-Blanquet’s legacy and data analysis in vegetation science. J. Veg. Sci. 17, 113–117. Quinn, J.M., Boothroyd, I.K.G., Smith, B.J., 2004. Riparian buffers mitigate effects of pine plantation logging on New Zealand streams: 2. Invertebrate communities. For. Ecol. Manage. 191, 129–146. Rackham, O., 1986. The History of the Countryside. JM Dent, London. Richardson, J.S., Naiman, R.J., Bisson, P.A., 2012. How did fixed-width buffers become standard practice for protecting freshwaters and their riparian areas from forest harvest practices? Freshwat. Sci. 31, 232–238.

Riipinen, M.P., Fleituch, T., Hladyz, S., Woodward, G., Giller, P., Dobson, M., 2010. Invertebrate community structure and ecosystem functioning in European conifer plantation streams. Freshwat. Biol. 55, 346–359. Roem, W.J., Berendse, F., 2000. Soil acidity and nutrient supply ratio as possible factors determining changes in plant species diversity in grassland and heathland communities. Biol. Conserv. 92, 151–161. Ryan, D.K., Kelly-Quinn, M., 2014. Effects of riparian canopy cover on salmonid diet and prey selectivity in low nutrient streams. J. Fish Biol. 353, 1–16. Ryan, D.K., Yearsley, J.M., Kelly-Quinn, M., 2013. Quantifying the effect of seminatural riparian cover on stream temperatures: implications for salmonid habitat management. Fish. Manage. Ecol. 20, 494–507. Shannon, C.E., Weaver, W., 1963. The Mathematical Theory of Communication Urbana, 125. Silvertown, J., Araya, Y., Gowing, D., 2015. Hydrological niches in terrestrial plant communities: a review. J. Ecol. 103, 93–108. Sykes, J.M., Lowe, V.P.W., Briggs, D.R., 1989. Some effects of afforestation on the flora and fauna of an upland sheepwalk during 12 years after planting. J. Appl. Ecol. 26, 299–320. Tank, J.L., Rosi-Marshall, E.J., Griffiths, N.A., Entrekin, S.A., Stephen, M.L., 2010. A review of allochthonous organic matter dynamics and metabolism in streams. J. N. Am. Benthol. Soc. 29, 118–146. Vannote, R.L., Minshall, G.W., Cummins, K.W., Sedell, J.R., Cushing, C.E., 1980. The river continuum concept. Can. J. Fish. Aquat. Sci. 37, 130–137. Wallace, H.L., Good, J.E.G., 1995. Effects of afforestation on upland plant communities and implications for vegetation management. For. Ecol. Manage. 79, 29–46. Wallace, H.L., Good, J.E.G., Williams, T.G., 1992. The effects of afforestation on upland plant communities: an application of the British National Vegetation Classification. J. Appl. Ecol. 29, 180–194. Zhang, K., Dang, H., Tan, S., Wang, Z., Zhang, Q., 2010. Vegetation community and soil characteristics of abandoned agricultural land and pine plantation in the Qinling Mountains, China. For. Ecol. Manage. 259, 2036–2047.