1 Title
2 Effects of vegetation cutting on invertebrate communities of high conservation value
3 Calluna upland peatlands
4
5 Running title
6 Vegetation cutting on invertebrates of Calluna
7
8 Authors
9 Roy Sanderson (Corresponding author): Modelling, Evidence & Policy Research
10 Group, School of Natural and Environmental Sciences, Newcastle University,
11 Newcastle upon Tyne, NE1 7RU, UK. Email [email protected]
12 orcid.org/0000-0002-9580-4751
13 Samuel Newton: North York Moors National Parks Authority, The Old Vicarage,
14 Bondgate, Helmsley, York, North Yorkshire, YO62 5BP, UK. Email
16 Jen Selvidge: Royal Society for the Protection of Birds, RSPB Geltsdale, Stagside
17 Cottages, Brampton, Cumbria, CA8 2PN, UK. Email: [email protected]
18
19 Abstract
20 1. Upland moors and bogs in the UK are peatlands of high nature conservation
21 value, many designated under the European Union’s Habitat Directive, with
22 internationally important bird populations, which depend on abundant
23 invertebrate populations when breeding.
24 2. Moorland management in the UK traditionally employs controlled burning in
25 10-30 year rotations of heather, Calluna vulgaris, creating habitat mosaics of
1 26 different species composition and physical structure. This can increase overall
27 invertebrate biodiversity and abundance, for certain key invertebrate groups.
28 Burning has potential negative environmental effects, including peat erosion
29 and contamination of drinking-water supplies.
30 3. Rotational cutting of vegetation is now being trialled as an alternative to
31 burning, but its long-term effects on invertebrates are poorly understood. We
32 surveyed invertebrates on a 16-year chronosequence of rotational cutting on
33 an extensive area of dwarf shrub vegetation on upland peat soils in Northern
34 England.
35 4. Invertebrate Simpson diversity was greatest on intermediate-aged patches,
36 and along edges between cut and uncut areas. Older patches, cut between
37 2000 and 2008, were dominated by ants (Hymenoptera-Formicidae), plant-
38 feeding bugs (Hemiptera-Auchenorrhyncha) and parasitic wasps
39 (Hymenoptera-Parasitica). Patches cut more recently, between 2009 and
40 2016, had significantly lower invertebrate abundance, and were dominated by
41 predatory ground beetles (Coleoptera-Carabidae), ants and harvestmen
42 (Opiliones).
43 5. There were significant relationships between vegetation and invertebrate
44 community composition under both invertebrate sampling methods. We
45 recommend that rotational cutting is used as the primary means of
46 management, it should be undertaken in approximately 15-20 year rotations,
47 in irregularly-shaped mosaics, to maximise the benefits to invertebrates and
48 associated wildlife.
49
50 Keywords
2 51 Heather moorland, wet heath, cutting, invertebrates, diversity, vegetation mosaics
3 52 Introduction
53 Heather moorlands and bogs constitute a key habitat for upland wildlife in the UK,
54 with 13 International Union for Nature Conservation (IUCN) Red List bird species
55 (Eaton et al., 2009), and approximately 75% of the world’s heather moorland in the
56 UK (Backshall, 2001; Tallis et al., 1997). The vegetation is dominated by heather
57 (Calluna vulgaris L.), which is a dwarf ericaceous shrub adapted for acid soil
58 conditions, and which can regenerate rapidly either from seed or vegetatively after
59 fire (Gimingham, 1975). Important species of birds, from both a conservation and
60 economic perspective, include hen harrier (Circus cyanus L.), golden plover
61 (Pluvialis apricaria L.), peregrine (Falco peregrinus L.) and red grouse (Lagopus
62 lagopus scoticus Latham). Invertebrates, especially Diptera and Coleoptera are
63 important as key components of the diets of both adults and young of many
64 passerine and wader birds (Buchanan et al., 2006). Some invertebrates are of
65 conservation importance, such as the large heath butterfly (Coenonympha tullia,
66 Muller), a Biodiversity Action Plan species that has declined by approximately 50% in
67 range over the last 20 years (JNCC, 2010). The areas are primarily upland, 300m
68 above sea-level, and often grade into wet heath and peat bog in the north and west
69 of the UK where rainfall levels are higher (Webb et al., 2010). Both the area and
70 conservation value of these habitats have declined as a result of drains dug between
71 the 1950s and 1990s to improve productivity, primarily for sheep grazing (Holden et
72 al., 2007). In practice whilst drainage increased the nutritional value of Calluna it
73 often resulted in growth of unpalatable grasses, reducing the overall grazing value
74 (Coulson et al., 1990; Stewart & Lance, 1991), and grants to support this practice
75 ceased in the 1980s. Large areas are subject to management for sport shooting of
76 red grouse.
4 77
78 Heather moor and wet heath is a plagioclimax, as without human intervention over
79 long periods (many decades) parts would start to revert to upland woodlands,
80 depending on soil conditions and areas of existing woodland (Ratcliffe, 1984).
81 Conventional management of heather moor and wet heath prevents natural
82 succession to upland woodlands and has been carried out for centuries, initially for
83 grazing livestock (mainly sheep) and more recently for red grouse. The latter are of
84 economic importance, with approximately 1500 people employed in England alone,
85 although the net economic benefits are difficult to estimate (Bennett, 2016; Sotherton
86 et al., 2009). Rotational prescribed burning, sometimes referred to as ‘muirburn’,
87 creates a mosaic of habitats with heather at a range of different life-stages. Watt
88 (1947) and later (Gimingham, 1975) classified the growth of Calluna into different
89 “life-stages”, often referred to as “seedling” (<1 year); “pioneer” (2-8 years); “building”
90 (9-15 years); “mature” (16-19 years) and “over-mature” or “degenerate” (> 20 years).
91 In practice there is often variation in timing and duration of these classes, especially
92 as the last two life-stages are relatively similar and sometimes merged (Macdonald
93 et al., 1995). The UK Government recommends that burning should not take place
94 on rotations of less than 12 years (Defra, 2007), and infrequent burns (over 15
95 years) may be needed if the Calluna is growing more slowly (Defra, 2007). There is a
96 general presumption against burns on deep peat, although in some circumstances
97 Natural England can provide permission for individual landowners. Glaves et al.
98 (2013) report average burn 'return periods' on heather-dominated moors in England
99 at 26.5 years and 64 years on deep peat (over 50 cm depth; Defra, 2007).
100 Prescribed burning is restricted to between 1st October and 15th April (Defra, 2007)
101 when wetter conditions allow managed fires to be easily controlled. These ‘cool
5 102 burns’ primarily burn the surface vegetation and leaf litter and thus aim to minimise
103 the risk of damage to peat, bryophytes and other moisture-requiring plants growing
104 on the peat, and arthropods on them. Calluna is adapted to fire, with rapid vegetative
105 regeneration when burnt in the pioneer or building stages, or from seeds (Hobbs &
106 Gimingham, 1987).
107
108 Upland birds require a mosaic of habitats, as this increases both the total arthropod
109 biomass (Usher, 1992) and species richness (Cole et al., 2010). Invertebrate prey
110 are an essential component of the diet of many upland birds (Buchanan et al., 2006),
111 with waders such as plovers (Charadriinae) and curlew (Numenius arquata L.)
112 feeding on soil-dwelling arthropods, especially Diptera-Tipulidae. Prescribed burning
113 has been reported to increase invertebrate abundance and species richness
114 (McFerran et al., 1995; Usher, 1992) probably as a result of greater vegetation
115 structural complexity and number of species. Complex vegetation structure also has
116 direct benefits to several bird species through provision of shelter.
117
118 The use of prescribed burning is controversial (Davies et al., 2016; Glaves et al.,
119 2013; Tucker, 2003) as it has both positive and negative effects. Prescribed burning
120 can be useful to reduce the amount of “over-mature” or “degenerate” Calluna (sensu
121 Watt, 1947), i.e. stands of heather typically over 25 years old with a large quantity off
122 dead woody material relative to fresh growth. Carefully managed prescribed burning
123 is used to create firebreaks, and thus minimise the frequency and extent of
124 uncontrolled wildfires (Santana et al., 2016). However, if poorly managed, prescribed
125 burns may periodically run out of control, causing damage to deep peat: recovery
126 from deep burns takes longer, and if the seedbank is destroyed is also dependent on
6 127 adequate seed dispersal (Legg et al., 1992; Maltby et al., 1990). Whilst the exact
128 effects of this damage are difficult to quantify, they include loss of stored carbon,
129 exacerbated by excessive water loss through drains (Billett et al., 2006; Brown Lee
130 et al., 2016; Davies et al., 2016). Prescribed burning is tightly regulated on habitats
131 of high conservation value, for example Sites of Special Scientific Interest (SSSI),
132 peat bog, wet heathland with permission required from Natural England (Defra,
133 2007). Unfortunately, surveys using remote-sensed data indicate that the majority of
134 Special Protection Areas and Special Areas of Conservation in the UK have been
135 adversely affected by burning, including areas of deep peat (Douglas et al., 2015).
136 The effects of prescribed burning on carbon-balance and drinking water quality,
137 particularly dissolved and particulate organic carbon (DOC and POC), are equivocal
138 (Holden et al., 2012; Worrall et al., 2013; Yallop et al., 2006). Nevertheless, some UK
139 water companies have started to restrict prescribed burning in parts of their
140 catchments, and encourage blocking of drains (Evans et al., 2005). The loss of
141 carbon to the atmosphere through prescribed burning on upland peatlands has led to
142 research into afforestation schemes (Bunce et al., 2014), particularly of native
143 species (‘rewilding’), but the effects on carbon balance are complex. For example
144 whilst forests may take carbon from the atmosphere, this can be offset by carbon lost
145 as the peat is drained for forest plantations (Byrne, 2006).
146
147 Given the concerns outlined above about prescribed burning, alternative vegetation
148 management regimes are now being implemented on some heather moors and wet
149 heaths. The most practical is to cut the heather with machinery, usually mowers or
150 flails depending on the size of the heather. Cutting machinery is pulled by tractors,
151 often adapted with double-tyres to ensure that pressure on the ground is low and
7 152 thus minimise damage to underlying peat (Moors for the Future Partnership, 2017).
153 Of course, using cutting machinery powered by fossil fuels will also emit carbon, and
154 have potential climate implications, although no carbon budgets are yet available.
155 Management via cutting results in reduced risk of the underlying peat being exposed,
156 and hence lower DOC runoff and drinking-water discolouration. On young stands of
157 heather rapid regrowth after light burns is primarily vegetative (Kayll & Gimingham,
158 1965) but hot burns or older stands with more dead material suffer greater impacts
159 from fire, favouring regeneration from seed. In general, regrowth after cutting is
160 vegetative, and hence more rapid (Liepert et al., 1993). Cotton and Hale (1994)
161 reported similar vegetation regrowth between heather burning and flailing after 5 to
162 10 years, whilst Muñoz et al. (2012) recommended that cutting of wet Erica heaths in
163 North West Spain be included in standard vegetation management.
164
165 The invertebrate communities of areas subject to prescribed burning have been
166 widely studied (Coulson et al., 1990; Gardner, 1991; Usher, 1992). It is self-evident
167 that vegetation structure will change, and groups such as spiders (Araneae) respond
168 rapidly, as they are more abundant and species-rich in structurally complex upland
169 vegetation (Dennis, 2003). Highly mobile arthropods, including pollinators and flying
170 predatory and herbivorous arthropod species generally show greatest resilience to
171 fires, whilst saprophagous species in the leaf litter are most sensitive (Moretti et al.,
172 2006). However there have been relatively few studies into the long-term effects of
173 cutting on the upland invertebrates. Hancock et al. (2011) investigated Araneae,
174 Coleoptera, Lepidoptera larvae and Hymenoptera-Formicidae in capercaillie (Tetrao
175 urogallus L.) habitats dominated by Calluna and Pinus, to compare the effects of
176 cutting and burning. They reported little difference in arthropod abundance or
8 177 composition between the two treatments, although after 5 years both had greater
178 arthropod biomass than the untreated control areas. Upland vegetation regenerates
179 relatively slowly after either burning or cutting, therefore sites where such
180 management has been in place for over 10 to 15 years are required to understand
181 the changes that take place. Longitudinal studies are relatively rare in these habitats,
182 and the use of chronosequences (space for time) provides a useful alternative (e.g.
183 Clay et al., 2015). Buddle et al. (2006) compared insects in a chronosequence of
184 aspen woods in Canada subject to wildfires or cutting, and found recovery of
185 invertebrate biodiversity was more rapid in clear-cut stands, but biodiversity
186 gradually declined over time as the vegetation recovered in both scenarios.
187
188 Management in the form of cutting or burning creates edges or ecotones between
189 vegetation of different physical structure and community composition. There is
190 evidence from other systems that such edges can often harbour greater numbers
191 and diversity of invertebrates than the habitats on each side of the boundary. For
192 example, Downie et al. (1996) reported 72% increase in species richness, and 141%
193 increase in Simpson diversity amongst Araneae at the interface between pasture and
194 coniferous woodland. This type of effect has also been observed in Lepidoptera (van
195 Halder et al., 2011) but would appear to be less important for Coleoptera-
196 Curculionidae than Araneae (Horváth et al., 2002) in similar ecotones.
197
198 It is self-evident that both burning and cutting have major effects on vegetation
199 structure, and this affects many groups of invertebrates. The primary aim of the study
200 was to understand the effects of cutting management on invertebrate diversity,
201 across a chronosequence of upland plots cut over a 16-year period at a site in North
9 202 West England. We hypothesise that invertebrate abundance and composition will
203 depend on both vegetation structure and composition, and that invertebrate
204 abundance will differ on the adjoining edges of vegetation of different ages. The
205 research had the following aims:
206 1. Determine the invertebrate diversity and composition on upland plots, cut
207 between 2000 and 2015, using both pitfall traps and sweep nets to ensure
208 that invertebrates with a representative range of behavioural characteristics
209 are sampled.
210 2. Compare invertebrate abundance within vegetation plots, and along the
211 boundary between plots of different ages, to assess evidence of an ‘edge
212 effect’ between habitats.
213
214 Methods
215 Study area
216 Sampling was undertaken at the Royal Society for the Protection of Birds (RSPB)
217 Geltsdale nature reserve in Cumbria, UK, within the North Pennine Moors Special
218 Protection Area or SPA (54° 54' 51.48" N, 2° 38' 34.8" E). SPA’s are designated for
219 bird protection under the European Union’s Birds Directive (2009/147/EC). The
220 reserve is approximately 5000 ha in area and contains a range of habitats, including
221 deciduous woodland and rough grasslands at lower altitude, through to blanket bogs
222 and heather moorland at higher altitude. The latter are dominated by Calluna
223 vulgaris-Eriophorum vaginatum blanket bog (National Vegetation Classification type
224 M19; Rodwell, 1991) with smaller areas of Calluna vulgaris-Vaccinium myrtillus
225 heath (NVC H12). The M19 community is typically found in the UK above 300m
226 altitude, on peat soils up to 2 m thick, pH 4, whilst H12 occurs at similar altitudes, but
10 227 generally slightly more freer-draining, shallower soils; invertebrate samples were
228 collected from areas dominated by M19. RSPB Geltsdale is of particular
229 conservation importance due to its birds, such as hen harrier, barn owl (Tyto alba
230 Scopoli), buzzard (Buteo buteo L.) and black grouse (Tetreo tetrix L.). Drains had
231 been cut into the blanket bog to reduce waterlogging, as was typical in the UK in the
232 first part of the 20th Century, especially towards the south of the reserve; however
233 over the last 15-20 years RSPB has been engaged on a policy of blocking drains, to
234 encourage more natural vegetation. Whilst historically large areas of what is now the
235 reserve were subject to prescribed burning management (with the inherent risks
236 mentioned in the Introduction), since 1998 RSPB has been cutting the vegetation
237 with the aid of a double-wheeled tractor, pulling a 3 m wide flail mower between
238 October and February. Cutting is done in strips approximately 30 m wide by 125 m
239 long, generally across contours, and the cut vegetation removed (it can be used to
240 create heather bales to block drains). A mosaic of patches with vegetation cut in
241 different years now exists at the northern end of Geltsdale reserve, and we took
242 advantage of this to create a ‘chronosequence’ of samples, from patches cut for the
243 first time in 2000, 2004, 2008, 2012 and 2015. An additional sixth patch (described
244 as ‘2012-recut’) was first cut in 2000 and then recut in 2012, and was included to
245 assess any regeneration effects over longer time periods.
246
247 Invertebrate and vegetation sampling
248 Invertebrates were sampled between June and August 2016 using both pitfall traps
249 and sweep nets. Pitfall traps consisted of polypropylene pots, approximately 8 cm
250 diameter × 10 cm depth, placed into holes dug into the soil with an auger, such that
251 their tops were level with the soil surface. They were filled with water to about 5 cm
11 252 depth, plus a small amount of detergent to break the surface tension and reduce the
253 chances of invertebrates escaping. In each cut area, 3 sets of pitfall traps were
254 placed along the centre line of a cut vegetation strip, with 30 m between each of the
255 3 sets. At each of these 3 sampling points, 4 pitfall traps were used, placed
256 approximately 1m apart; the positions of 3 sets of pitfalls were marked by white
257 canes for ease of relocation (Patterson et al., 2018). Pitfalls were placed out for three
258 sampling periods of 7-8 days duration: 4th July to 13th July 2016; 20th to 27th July
259 2016; 27th July to 3rd August 2016. Sweep net sampling was undertaken using a
260 canvas sweep net (40 cm diameter) doing 10 sweeps at a steady walking pace
261 within 5 m to 10 m of each of the 3 sets of pitfall traps (New, 1998). Sweep net
262 samples were collected on 30th June 2016, 20th July 2016 and 27th July 2016. Both
263 pitfall and sweep samples were transferred to plastic bags in the field, and kept
264 refrigerated prior to identification and storage in 70% industrial methylated spirit in
265 the laboratory. All invertebrates were identified to family taxonomic level, although
266 this was not possible for some groups such as micro-Lepidoptera, Hymenoptera-
267 Parasitica and immatures. Whilst invertebrates were not identified down to subfamily
268 or species level, the use of families has been shown to be a valid measure of
269 invertebrate biodiversity and for comparing community composition (New, 1998;
270 Oliver & Beattie, 1996; Patterson et al., 2018). An additional small-scale survey was
271 undertaken across the boundaries between plots cut on different dates, to
272 investigate any ‘edge effects’ on invertebrate abundance. Four pitfall traps were
273 placed in the centre of the 2000-cut plot, the centre of the 2012-cut plot, and on the
274 edge between the two plots, during the same sampling periods as the main survey.
275 Equivalent sweep net samples were also collected at these three locations, using the
12 276 same methods as for the main survey. Note that the overall design is unreplicated,
277 as we did not have multiple plots of each age since cutting to compare.
278
279 A vegetation survey was undertaken on 3rd August 2016. Vegetation structure at the
280 centre of each set of 4 pitfall traps was measured using a metre rule positioned
281 vertically, and the number vegetation leaves / twigs etc. touching each 20 cm interval
282 on the metre rule recorded. The percentage cover of all vascular plants and
283 bryophytes was estimated by eye using individual 1 m2 quadrats at each of the
284 sampling points (i.e. adjacent to each of the 3 sets of pitfall traps, across the 6 cut
285 areas, giving 18 quadrats in total).
286
287 Data analysis
288 All statistical analyses were undertaken in R (R Core Team, 2015). Linear models
289 were used to test the effects of year (and any recut effect) on taxon richness,
290 Simpson’s diversity (1/D) and abundance. Both linear and quadratic terms were fitted
291 in the models to explore changes over time, and quantile-quantile plots used to
292 confirm that model assumptions had been satisfied. Multivariate analyses used the R
293 ‘vegan’ package (Oksanen et al., 2013). Non-metric multidimensional scaling
294 (NMDS) was used to compare the community composition of the invertebrate and
295 vegetation samples, using the ‘metaNMDS’ function to create a set of random starts
296 (Oksanen et al., 2013). Analyses were undertaken on the combined pitfall and sweep
297 data (Eyre & Leifert, 2011; Moretti et al., 2006), but data and plots on the pitfall and
298 sweep net samples are shown in the Supplementary Information. The relationship
299 between invertebrate and vegetation community composition was tested through
300 Procrustes rotation (Peres-Neto & Jackson, 2001) which produces Procrustes
13 301 correlation statistic which is equivalent to the conventional univariate statistic, but
302 with the advantage of analysing all axes simultaneously.
303
304 Results
305 Vegetation
306 The vegetation was dominated by the dwarf shrub Calluna vulgaris, which was the
307 single most common species, both in terms of its frequency within quadrats, and
308 cover abundance. Mean cover of Calluna was 88.3% on the oldest areas (cut in
309 2000), dropping to 68.3% in areas cut in 2008, and only 10.0% in areas cut in 2015,
310 one year prior to the survey. As might be expected, the opposite pattern was
311 observed for the second most common ground cover, leaf litter and brash, being the
312 most dominant cover in those areas cut in the preceding three years (25.3% to
313 85.0%), but less than 4% on the oldest areas. The sedge Eriophorum vaginatum L.
314 occurred in most quadrats, whilst bryophytes, especially Pleurozium schreberi, Mitt.
315 were widespread, often forming a mat (over 40% cover) underneath the canopy of
316 taller vegetation. Both the species richness and Simpson diversity of the vegetation
317 were highest in the intermediate-aged plots, at a maximum in 2008 (i.e. cut 8 years
318 before the survey in 2016), with lower diversity for both the oldest and most recently
319 cut plots (Fig. 1). The diversity of the 2012-recut plot was not significantly different to
320 samples from the plot first cut in 2012 (Table 1). This pattern is also apparent in the
321 results of the NMDS ordination (Fig. 2), where the 2012-recut and 2012 plots had
322 similar positions on NMDS Axis 1 (Fig. 2a), although they separated on NMDS Axis
323 2, primarily as a result of larger amounts of Erica and Eriophorum species in the
324 2012-recut compared to the 2012 plot (Fig. 2b). The vegetation structure showed
325 clear changes with time with the tallest vegetation being that last cut in 2000, up to
14 326 60 cm tall in some quadrats. Plots cut since 2012 were dominated by vegetation less
327 than 20 cm tall, with virtually identical vegetation structure profiles for the 2012 and
328 2012-recut plots (Fig. 3).
329
330 Invertebrates
331 The most abundant invertebrates overall were Hymenoptera-Formicidae (ants),
332 Coleoptera-Carabidae (ground beetles), Opiliones (harvestmen), Hemiptera-
333 Cicadellidae (leaf hoppers) and Hymenoptera-Parasitica (mainly Braconidae)
334 (Supplementary Tables S1 to S3). There were differences in the most common
335 invertebrates collected by the two sampling methods, with pitfall trap samples
336 dominated by epigeal species such as Hymenoptera-Formicidae and Coleoptera-
337 Carabidae, whilst in sweep nets abundant groups included those active in the
338 vegetation canopy such as Hymenoptera-Parastica, Hemiptera-Cicadellidae and
339 Diptera-Acalyptratae.
340
341 Invertebrate abundance, taxon richness and Simpson diversity were highest in
342 intermediate-aged plots, last cut in 2004 or 2008, across both sampling methods.
343 Values were lowest in plots cut within four years of the survey, i.e. 2012 first cut, the
344 2012-recut, and 2015 (Fig. 4). Plots cut on or before 2008 were dominated by
345 Formicidae and Auchenorrhyncha (mainly Cicadellidae and Cercopidae; Fig. 5a),
346 whilst areas cut since 2008 had much lower overall invertebrate abundance, and
347 were dominated by Carabidae. These older plots were characterised by a range of
348 invertebrate taxa including Diptera-Tenthredinidae, Coleoptera-Elateridae, Araneae-
349 Gnaphosidae, Hemiptera-Aphididae (Fig. 5b). These differences are reflected in the
350 NMDS results (Fig. 6) with the three most recently cut plots having high NMDS Axis
15 351 1 scores whilst the three older plots had low NMDS Axis 1 scores (Fig. 6a). The older
352 plots also showed less between-sample variation in invertebrate community
353 composition, as indicated by the smaller polygons, compared to the more recently
354 cut plots (Fig. 6a). There was very little difference between those plots first cut in
355 2012, or the 2012-recut areas, for any of the invertebrate metrics (Table 2). The
356 boundary-effect study provided weak evidence of higher invertebrate abundance
357 along the boundary between the vegetation cut in 2012 and the oldest heather,
358 although this was marginally non-significant (F=2.92, p=0.084). Procrustes analysis
359 indicated a significant association between the vegetation and invertebrate
360 community composition (Procrustes correlation 0.488, P=0.022), with the strongest
361 association between the vegetation and invertebrates in the oldest patches (Fig. 7).
362
363 Discussion
364 Both the invertebrate and vegetation results indicate that greatest diversity and
365 abundance of species occur in patches last cut about 8 to 12 years ago. Areas cut
366 more than 15 years ago (their exact ages are uncertain), or within the last 5 years,
367 had poorer faunas and floras. To maintain these intermediate-aged patches
368 proactive management of the vegetation is required to produce a mosaic of habitats
369 of different ages, probably over a 10 to 15 year timescale to make it practical to
370 cover large areas of upland hillside. Here the proactive management used was
371 cutting, although the typical management is prescribed burning. It is important to be
372 cautious when interpreting these types of invertebrate data due to biases in sampling
373 method but the broad patterns of the results were similar for both sampling methods
374 (see Supplementary Material).
375
16 376 Invertebrates are an essential resource for upland birds, especially chicks, and whilst
377 the results imply that areas cut 8 to 12 years ago had highest numbers and diversity
378 of invertebrates, it would be wrong to automatically infer that this was the best
379 habitat for breeding birds. Indeed, some invertebrate species known to be important
380 in bird diets, such as macro-Lepidoptera, Oligochaeta and Carabidae (Buchanan et
381 al., 2006), were more characteristic of the most recently cut plots. Conversely, whilst
382 the oldest plots might be classed as being dominated by ‘over-mature’ Calluna
383 vulgaris, with lower vegetation diversity and invertebrate abundance, they
384 nevertheless provide additional cover for birds into which to escape predators. For
385 example, edges between contrasting habitat types might be particularly important for
386 fledgling birds, as they appear to provide higher food resources. Furthermore,
387 precocial chicks that can be exposed to predators in short vegetation could take
388 cover in the adjacent mature Calluna-dominated habitat. This type of pattern has
389 been suggested for golden plover chicks which appear to have higher survival rates
390 in patchworks of Calluna, soft rush, or wet heath (Whittingham et al., 2001) and
391 potential predatory birds such as merlin (Falco columbarius L.) and hen harriers.
392
393 To some extent the results appear to accord with the ‘intermediate disturbance
394 hypothesis’ (IDH; Connell, 1978), with higher vegetation and invertebrate diversities
395 in moderate-aged plots (8-12 years since last cut) rather than recently cut or over-
396 mature Calluna, a pattern that has been previously observed on burnt and cut
397 heathlands (Usher, 1992). However, in reality, the patterns probably reflect the
398 changes in vegetation structure and species composition that arise from the cutting
399 treatments, and should not be viewed as providing direct support for IDH (Fox,
400 2013), especially the greater number of species of plants in intermediate-aged plots
17 401 (Fig. 1a). The oldest plots were dominated by Calluna which often comprised over
402 80% of the vascular plant cover. Furthermore, our study was not undertaken over a
403 15-year period from 2000 to 2015, but rather in a single year, using the plots as a
404 chronosequence. Chronosequences have weaknesses (Johnson & Miyanishi, 2008)
405 in that they may not represent the actual colonisation and extinction processes that
406 have occurred over time at a site, and assume a common site environmental history.
407 However, given that the study plots at Geltsdale were relatively close to each other,
408 on the same soil type and had identical original vegetation (based on surveys done
409 prior to cutting), it is reasonable to assume that the use of a chronosequence was
410 valid (Gill, 1977). Studies using chronosequences on heathlands have been explored
411 in detail, including succession following fires (Hobbs & Gimingham, 1987). An
412 advantage of chronosequences, particularly for invertebrate studies, is that temporal
413 year-to-year variations in weather conditions are smoothed out, and it is possible to
414 undertake the research with more limited resources in a single season (New, 1998).
415
416 The main drivers of the invertebrate community diversity and composition appeared
417 to be both the vegetation structure and composition, which accords with previous
418 studies. The oldest plots, last cut in 2000, were dominated by Calluna vulgaris, which
419 at times formed almost 100% of the vegetation cover by area. Mature heather
420 generally has a lower number of invertebrate species (Gardner et al., 1997; Usher,
421 1992), although it may contain high abundance of specialists such as heather weevil
422 (Micrelus ericae). Likewise, the most recently cut plots (2015) were dominated by
423 litter and brash from the previous year’s cut, with a limited number of grasses,
424 mosses and herbs growing back in small patches. This open habitat will also be
425 exposed to greater extremes of temperature, rainfall and other environmental
18 426 stresses. The intermediate-aged plots contained a mosaic with a wider range of
427 vegetation species and growth forms, which in turn supported greater invertebrate
428 diversity. The vegetation was most dense in the 2004 plots, although these lacked
429 the tallest height categories (Fig. 3), with invertebrate richness and diversity highest
430 in the 2004 and 2008 plots. Vegetation structure is known to have strong effects on
431 many groups of invertebrates, especially Araneae, in both the UK and internationally
432 (e.g. Dennis, 2003; Haysom & Coulson, 1998; Nittérus & Gunnarsson, 2006; Perner
433 et al., 2005).
434
435 Some caution is needed in interpretation of our results, particularly as soil
436 invertebrates were not sampled. Collembola, mites, and the immature stages of
437 several insects such as Diptera-Tipulidae are abundant in these habitats (Buchanan
438 et al., 2006). Likewise, each of the two sampling methods deployed has their own
439 biases, such that their effectiveness in trapping invertebrates is affected by
440 vegetation structure. Pitfall traps are biased towards fast moving surface-active
441 invertebrates, and collect more in relatively open areas (New, 1998) whilst sweep
442 nets can be challenging to standardise between samples, and are less effective at
443 collecting invertebrates that reside low in the vegetation canopy, in short foliage, or
444 woody vegetation which slows down the motion of the net. Overall invertebrate
445 abundance was highest in the pitfall traps for areas cut in 2008 (Supplementary
446 Information Fig S1c), and highest for the sweep samples in the oldest (2000) areas
447 (Supplementary Information Fig S2c), although changes in Simpson diversity were
448 broadly similar (Supplementary Information Fig S1b and S2b). Insects feeding on
449 Calluna, especially Hemiptera-Auchenorrhyncha and Coleoptera-Curculionidae,
450 were most abundant on the oldest patches (Supplementary Information, Tables S1 to
19 451 S3) although no obvious temporal change was observed amongst epigeal predators
452 of Coleoptera-Carabidae. Only one survey of the vegetation was undertaken, and
453 therefore temporal change in the vegetation in relation to the invertebrates was not
454 analysed. However, the vegetation on these sites is relatively slow-growing, and this
455 type of approach using multiple invertebrate samples in a season, but only one
456 vegetation sample, is not uncommon (Sanderson et al., 1995).
457
458 In conclusion, our results indicate that cutting can be used to create a mosaic of
459 habitats on heather moorland and bogs, which can support both a high taxonomic
460 diversity and abundance of invertebrates, which will benefit other species, especially
461 birds. For practical reasons our study did not include a direct comparison of
462 chronosequences of burnt and cut areas at the same site, and so it is not possible to
463 evaluate which creates the highest invertebrate diversity or abundance. Swengel
464 (2001) in an international review of insect responses to fire notes a severe lack of
465 quality studies comparing fire with cutting, and the resultant gap in knowledge.
466 Maximum arthropod diversity was found at about 8 to 12 years, although the patterns
467 over time were less consistent for abundance. It is likely that approximately 15 to 20-
468 year rotational cut cycle would be optimum to maintain both invertebrate abundance
469 and diversity, although ideally the chronosequence should extend to 20 or 25 years
470 to confirm. A rotational cut cycle of 15 to 20 years is shorter than the 20 to 25 years
471 typical in prescribed burning on blanket bog (Defra, 2007), and may reflect faster
472 regeneration of heather via vegetative regrowth after cutting (Liepert et al., 1993),
473 rather than from seed after burning, as well as local site characteristics. Of course,
474 shorter rotation times for cutting in comparison to burning will increase management
475 costs, against the benefits of reduced potential environmental risks. Our findings also
20 476 highlight the importance of boundaries between habitat patches of different ages in
477 increasing invertebrate biodiversity, suggesting that mosaics should be cut with
478 irregular shapes, to maximise the beneficial effects of edges.
479
480 Acknowledgements
481 SN was funded by Newcastle University. We also wish to Samantha Tranter
482 (Pennine Dales AONB) for advice on this project.
483
484 References
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24 663 Figure legends
664 Fig. 1. Change in a) vegetation species richness, b) vegetation Simpson diversity
665 and c) vegetation cover in relation to cutting year. All plots surveyed in 2016.
666
667 Fig. 2. NMDS ordination of vegetation; a) sample points with the year of last cutting
668 (2000R is the recut strip; all plots surveyed in 2016) and b) the individual species.
669 NMDS stress=0.145; NMDS r-squared=0.979.
670
671 Fig. 3. Patterns in vegetation structure, as measured by numbers of hits of plants, in
672 response to year of cutting. All plots surveyed in 2016.
673
674 Fig. 4. Change in a) invertebrate taxonomic richness, b) invertebrate Simpson
675 diversity and c) overall abundance in relation to cutting year. All plots surveyed in
676 2016.
677
678 Fig. 5. Top-ranked invertebrate taxa in areas a) cut on or before 2008 and b)
679 between 2009 and 2015.
680
681 Fig. 6. NMDS ordination of invertebrates; a) sample points with the year of last
682 cutting (2000R is the recut strip; all plots surveyed in 2016) and b) the individual
683 invertebrate taxa. NMDS stress=0.168; r-squared=0.972.
684
685 Fig. 7. Mean residuals from Procrustes analysis to compare fit of invertebrate and
686 vegetation community composition. Smaller residuals indicate a stronger relationship
687 between vegetation and invertebrate community composition.
25 688 Table 1. Vegetation species richness and Simpson diversity: summary of linear
689 models on effects of time since last cut, and recut or first cut.
Variable Species richness Simpson diversity Vegetation cover
First cut / recut F1,14=0.501; P=0.491 F1,14=1.130; P=0.306 F1,14=0.696; P=0.418
Year F1,14=8.035; P=0.013 F1,14=5.485; P=0.034 F1,14=1.376; P=0.260
Year (quadratic) F1,14=8.037; P=0.013 F1,14=5.484; P=0.035 F1,14=1.381; P=0.259
690
691
692
693 Table 2. Invertebrate species richness and Simpson diversity: summary of linear
694 models on effects of time since last cut, and recut or first cut.
Variable Species richness Simpson diversity Total abundance
First cut / recut F1,14=0.986; P=0.338 F1,14=0.380; P=0.548 F1,14=1.855; P=0.195
Year F1,14=5.458; P=0.035 F1,14=9.164; P=0.009 F1,14=0.712; P=0.413
Year (quadratic) F1,14=5.473; P=0.035 F1,14=9.180; P=0.009 F1,14=0.716; P=0.411
695
696
26 697
698 Fig. 1. Change in a) vegetation species richness, b) vegetation Simpson diversity 699 and c) vegetation cover in relation to cutting year. All plots surveyed in 2016.
27 700
701 Fig. 2. NMDS ordination of vegetation; a) sample points with the year of last cutting
702 (2000R is the recut strip; all plots surveyed in 2016) and b) the individual species.
703 NMDS stress=0.145; NMDS r-squared=0.979.
704
28 705
706 Fig. 3. Patterns in vegetation structure, as measured by numbers of hits of plants, in
707 response to year of cutting. All plots surveyed in 2016.
708
29 709
710 Fig. 4. Change in a) invertebrate taxonomic richness, b) invertebrate Simpson
711 diversity and c) overall abundance in relation to cutting year. All plots surveyed in
712 2016.
30 713
714 Fig. 5. Top-ranked invertebrate taxa in areas a) cut on or before 2008 and b)
715 between 2009 and 2015.
716
31 717
718 Fig. 6. NMDS ordination of invertebrates; a) sample points with the year of last
719 cutting (2000R is the recut strip; all plots surveyed in 2016) and b) the individual
720 invertebrate taxa. NMDS stress=0.168; r-squared=0.972.
32 721
722 Fig. 7. Mean residuals from Procrustes analysis to compare fit of invertebrate and
723 vegetation community composition. Smaller residuals indicate a stronger relationship
724 between vegetation and invertebrate community composition.
33