Restoration and recovery from acidification in ... - Wiley Online Library

Journal of Applied Ecology 2009, 46, 164– 174

doi: 10.1111/j.1365-2664.2008.01587.x

Restoration and recovery from acidification in upland Welsh streams over 25 years

Blackwell Publishing Ltd

S. J. Ormerod* and Isabelle Durance Catchment Research Group, Cardiff School of Biosciences, Cardiff University, Cardiff CF10 3US, UK

Summary 1. Streams affected by acid deposition should now be recovering biologically, but long-term assessments are scarce. Here, we use the experimental catchments at Llyn Brianne (Wales, UK) to evaluate trends over 25 years (1981–2005) in the chemistry and macroinvertebrates of acid moorland and forest streams restored by liming relative to those responding ‘naturally’ to reduced deposition. 2. Mean H+ concentrations in acid moorland streams fell by c. 15–16 μquiv L−1 over the study, increasing mean winter pH by 0·8–1·3 units to pH 5·6–6·1. Liming moorland streams in 1987 and 1988 increased mean pH to 5·5–6·4, but differences from naturally recovering streams diminished over 12–18 years. 3. In limed and acid moorland streams, changes in invertebrate composition were consistent with recovery, and near-identical. Four acid-sensitive species, from a local pool of 29, increased significantly in abundance or occurrence, but effects were too small to increase similarity with circumneutral reference streams. 4. Mean H+ in acid forest streams declined by 8–15 μequiv L−1, but mean winter pH increased by only 0·4 units and remained too acid for invertebrate recovery (mean pH 4·8–5·2; Al > 0·3–0·6 mg L−1). One forest stream limed in 1987 and 1988 remained at mean pH > 5·9 thereafter, but there was no invertebrate response. 5. Climate affected recovery pattern. After accounting for time trends, wet winters increased acidity in moorland and forest streams sufficiently to offset 21–41% of the total 25-year decrease in H+ concentration. 6. Synthesis and applications. These data from one of the world’s longest running experiments on acidification confirm that upland British streams are recovering, but ecological effects are marginal and vary with land use. Conifer forest streams at Llyn Brianne remain too acid for sensitive invertebrates, while moorland streams are still at risk from acid events. In this example, liming had few long-term benefits compared with natural recovery, and we suggest that this should be a key, general criterion in evaluating the outcomes of ecological restoration. Key-words: acidification, acid rain, insects, invertebrates, Llyn Brianne, restoration, recovery, streams

Introduction Even during the peak of the acid rain problem, rainfall in Western Europe was only moderately acidified by comparison with other regions. However, large rainfall volumes in these relatively maritime locations multiplied S and N deposition onto base-poor rocks and soils to affect extensive areas (Reynolds et al. 1999). Wales is a well-known example, where half of the 24 000-km stream network was acidified to < pH 4 –5·7 either chronically or intermittently (Edwards, Stoner & Gee 1990). Some land-uses exacerbated this effect, in particular *Correspondence author. E-mail: [email protected]

exotic conifers planted at high altitudes (Fowler, Cape & Unsworth 1989; Ormerod, Donald & Brown 1989). In turn, acidification affected algae (Hirst et al. 2004), macrophytes (Ormerod, Wade & Gee 1987), micro- and macroinvertebrates (Weatherley & Ormerod 1987; Rundle & Ormerod 1991), salmonids (Milner & Varallo 1990), riverine birds (Ormerod et al. 1991), and functions such as decomposition (Merrix, Lewis & Ormerod et al. 2006). Similar changes have been detected frequently elsewhere. SO2 emissions in the UK have now declined to < 15% of their peak in the 1970s and 1980s, with sulphur deposition falling in many locations by at least 50%. Welsh data reflect these trends (Reynolds et al. 1999, 2004), and calculated

© 2008 The Authors. Journal compilation © 2008 British Ecological Society

Restoration and recovery from acidification 165 critical loads – the degree to which deposited acidity can be buffered – suggest that less than a third of Welsh freshwaters should now be impacted (Hall et al. 2004). Available data from streams confirm that chemical recovery is underway (Reynolds et al. 2004, Davies et al. 2005). Evidence for associated biological recovery is more tentative. Although reversal is apparent, only a proportion of lost taxa have been regained in recovering waters, while other acid-sensitive species occur only sporadically (Bradley & Ormerod 2002a; Tipping et al. 2002; Monteith et al. 2005). The best explanation is that continued acid episodes during rainstorms or snowmelt still exclude sensitive organisms even where mean pH has increased (Kowalik et al. 2007). Such shortfalls in chemical and biological recovery have prompted continuing investigation of local, symptomatic treatment of acidification using calcium carbonate (McKie, Petrin & Malmqvist 2006; McClurg et al. 2007). At the experimental catchments at Llyn Brianne in central Wales, for example, liming increased pH and calcium concentrations relative to acid reference streams while reducing aluminium (Bradley & Ormerod 2002a). In a broader ecological context, this response to impairment represents a prime example of river restoration (Palmer et al. 2005). One of the major needs in assessing natural recovery from acidification and evaluating the ecological effects of restorative management is for comparative long-term data. While monitoring has revealed broad trends (Davies et al. 2005; Monteith et al. 2005), long-term comparisons between streams restored experimentally by liming and those recovering naturally are scarcer. At Llyn Brianne, such data now span 25 years, collected initially during baseline surveys (Stoner, Gee & Wade 1984), and from the mid-1980s following catchmentscale manipulations (Bradley & Ormerod 2002a). Over this prolonged time-scale, the data allow also an assessment of how climatic variation between years has affected recovery either through effects on organisms (Bradley & Ormerod 2001; Durance & Ormerod 2007) or effects on acid-based status (Ness et al. 2004; Evans 2005; Eimers et al. 2007). Here, we evaluate recovery from acidification over a 25-year period at the Llyn Brianne experimental catchments (1981– 2005), judged from acid-base status and invertebrates. We compare trends in streams draining catchments limed in 1987 and 1988 with acid and circumneutral streams draining forest and moorland. In addition to comparing, for the first time, trends against the baseline provided by Stoner, Gee & Wade (1984), the work extends previous analyses forwards by 7 years to provide one of the world’s longest ever experiments on chemical and biological recovery from acidification (Bradley & Ormerod 2002a). We also evaluate how climatic variation has affected inter-annual pH variation. We tested these four specific hypotheses: 1. There should be chemical recovery from acidification at Llyn Brianne, specifically increasing pH, decreasing H+, and decreasing aluminium. 2. Inter-annual variations in discharge, mediated by climatic variability, affect chemical recovery (Ness et al. 2004; Eimers et al. 2007).

3. Recovering streams should have gained acid-sensitive species, but to a limited extent because of continued acid episodes (Kowalik & Ormerod 2006; Kowalik et al. 2007). 4. Liming should have accelerated both chemical and biological recovery.

Methods STUDY AREA

The Llyn Brianne catchments cover c. 300 km2 of upland Wales (215 – 410 m above sea level) in the upper Afon Tywi (52°8′ N 3°45′ W; Fig. 1) and have been extensively described (Weatherley & Ormerod 1987; Edwards, Stoner & Gee 1990). The 14 basins of 15–264 ha used here contain permanent second- to third-order streams that rise either in rough, sheep–grazed moorland (CI1–CI6; LI5–LI7) or plantations of Sitka spruce Picea sitchensis Carr. with lodgepole pine Pinus contorta Doug. (LI1–LI4; LI8). Base-poor Ordovician and Silurian rocks combine with brown podzolic soils, stagnopdzols and peats to produce soft-water runoff (mean total hardness 3·9– 7·9 mg CaCO3 L−1) sensitive to acid deposition. Some streams (LI6 and LI7) are buffered by calcite veins at 15–19 mg CaCO3 L−1. The catchments of three streams (CI2, CI5, LI4) were limed in 1987 and 1988 either entirely over their catchments or in hydrological source areas using powdered calcium carbonate (see Bradley & Ormerod 2002a). Some reductions of forest cover have occurred consistent with normal logging operations, with LI2 and LI3 most affected. Climate has varied over the study: winter climatic warming of 1·4– 1·7 °C has affected organisms in circumneutral streams (Durance & Ormerod 2007); positive phases of the North Atlantic Oscillation have increased temperature and discharge, reducing inter-annual similarity among invertebrates in all streams without affecting species composition (Bradley & Ormerod 2001; Durance & Ormerod 2007).

STREAM INVERTEBRATES

Stream macroinvertebrates were first sampled at Llyn Brianne in 1981–1982 by Stoner, Gee & Wade (1984) and from 1985–2005 (except 1991) by Cardiff University using identical, quality-appraised methods (Bradley & Ormerod 2002b). All samples were collected in April using standardized kick-samples of 3-min total duration aggregated between riffles (2 min) and marginal habitats (1 min) using a hand-net (0·9 mm mesh; 230 × 255 mm). Animals were identified where practicable to species (i.e. except for Diptera (family) and Oligochaeta (class)), and abundances per sample recorded for each. All nomenclature conforms to the CEH Coded Checklist of Animals Occurring in Fresh Water in the British Isles, Version 5 (November 2007). For some analyses, we used individual invertebrate records for all 14 streams, for which data spanned 8–22 years per location over the 25-year study. For invertebrate trend analysis, however, we used replicate stream pairs with the longest and most complete sampling runs containing 20–22 years’ data: the acid forest streams, LI1 and LI2 (range of annual mean pH 4·9–5·4); the acid moorland streams CI1 and CI4 (pH 5·2–6·1); the limed moorland streams CI2 and CI5 (pH 5·5–6·7); and the circumneutral moorland streams LI6 and LI7 (pH > 6·9). In these cases, abundances were averaged and assemblages pooled between replicate stream pairs to increase the probability of reflecting rarer, acid-sensitive taxa in analysis. Sampling validation suggested that equivalent effort collects over 90% of all but the rarest

© 2008 The Authors. Journal compilation © 2008 British Ecological Society, Journal of Applied Ecology, 46, 164– 174

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S. J. Ormerod & I. Durance

Fig. 1. The study catchments at Llyn Brianne, with geographical position in Wales shown relative to sites in the Welsh Acid Waters Survey.

species thereby minimizing sampling effects (Bradley & Ormerod 2002b). Invertebrate data were also available from samples collected using identical methods on upland streams during the regional Welsh Acid Waters Survey (WAWS; see inset map, Fig. 1) in April 1984 and in 1995. We used these to assess variations in the acid-base tolerances of different invertebrate species (see below). The sites common to both WAWS surveys (n = 74) were spread through central and north Wales on base-poor streams in independent sub-catchments, at similar altitudes to Llyn Brianne, and draining similar soils (Wade, Ormerod & Gee 1989).

CHEMISTRY

As with invertebrates, chemical data were first collected from the streams in 1981–1982 (Stoner, Gee & Wade 1984). Water samples have been collected thereafter at weekly–monthly intervals using the same standardized procedures, although the exact length of each data run varies. LI1, LI2, LI8 and LI6 have been sampled since 1981 to give c. 600–1300 individual samples per stream; LI4, CI1, CI2, CI4 and CI5 annually since 1984 (c. 500–1100 samples per stream); and CI3, CI6; LI3, LI5, LI7 more sporadically but for at least 8 years from 1984 onwards (see Weatherley & Ormerod 1987; Bradley & Ormerod 2002a). Briefly, pH was measured on samples returned to the laboratory by combination glass electrode, while Ca and Al were measured using atomic absorption spectrophotometry after acidification and filtration at 0·45 μm. Chemical samples during WAWS were collected weekly (1984) or monthly (1995) using identical methods. In view of their biological importance, we calculated mean calcium concentrations, minimum and mean pH, and mean and maximum aluminium concentrations annually, but also paid particular attention to trends during the winter period (October to March) prior to invertebrate sampling (n = 10–26 samples per site per year). Seasonal variations in rainfall mean that conditions are most acid at this time of the year, and explain most variations in invertebrate assemblages between sites (Wade, Ormerod & Gee 1989).

CLIMATIC DATA

Hydrological data were not collected regularly at Llyn Brianne, but discharge was measured continuously in streams of identical order and similar altitude by the Centre for Ecology and Hydrology on Plynlimon, 37 km to the north (Plynlimon flume; 52°28·14′ N, 3°41·16′ W). Inter-annual trends here track those at Llyn Brianne, validated by comparison with discharge at a second gauged site on the Afon Cothi (51°51·37′ N, 4°11·00′ W), also in the Tywi system (r = 0·90, n = 289 monthly mean values, P < 0·0001). We used the average daily flows in cubic metres per second respectively for the ‘winter’ (October–March inclusive) and ‘summer’ (April–September inclusive) in each year. To assess any effects of the North Atlantic Oscillation (NAO) on inter-annual variations in runoff acidity, we used the winter index (December–March inclusive; provided by the Climate Analysis Section, NCAR, Boulder, CO, USA; Hurrell 1995), calculated from the difference in sea surface pressure between the Azores and Iceland. Positive values are associated with mild, wet winters in north-west Europe and negative values with cold, dry winters (Hurrell 1995).

DATA ANALYSIS

To test hypothesis 1, trends through time at Llyn Brianne in mean pH, minimum pH, calcium and aluminium concentrations were assessed using regression, with quadratic terms where appropriate. We performed this trend analysis on pH data since these are normally distributed and best met the requirements for regression, while proton concentrations in Llyn Brianne streams are log-normally distributed (see Stoner, Gee & Wade 1984). Analyses using H+ gave conclusions that were effectively unchanged. Moreover, we converted the resulting trends to H+ concentrations to compare across streams of contrasting pH. To test hypothesis 2 with respect to climatic effects on inter-annual variation in runoff acidity, we used summer or winter discharge, or the annual winter index of the NAO, as additional predictors, and assessed whether these variables explained

© 2008 The Authors. Journal compilation © 2008 British Ecological Society, Journal of Applied Ecology, 46, 164–174

Restoration and recovery from acidification 167 any residual variation between years. We focussed this analysis on H+ to allow equal assessment of effects across all parts of the pH range at different times and in different stream types. To test hypotheses 3 and 4, we assessed trends in overall species composition and among individual species expected to be acidsensitive according to the Welsh Acid Waters Survey (WAWS). For each taxon, we assessed the winter mean and minimum pH at occupied WAWS sites in each of the survey years. Taxa were next ranked in order of the increasing mean pH at which they occurred. Finally, we used Ward’s (1963) method to classify taxa into four categories of acid sensitivity (I = acid-tolerant to IV = acid-sensitive) on the basis of both the mean and minimum pH for all the sites they occupied. At Llyn Brianne, we used Detrended Correspondence Analysis (DCA) to ordinate all samples from all years and streams simultaneously using CANOCO 4·5, with rare species down-weighted and abundances log-transformed. As one of the most straightforward and widely used ordination methods, DCA uses reciprocal averaging to order samples objectively according to the frequency of co-occurrence among their constituent taxa (Van Der Maarel 1969). Sample scores reflect turnover in composition along orthogonal axes, such that 4 SD ≡ 100% change on any one axis. Scores can then be related quantitatively to sample attributes or conditions, for example, the pH under which each has been collected, and for this purpose, we used Pearson correlations after checking assumptions. We preferred unconstrained over constrained ordination (i.e. canonical correspondence analysis) in view of the greater flexibility allowed in subsequent data analysis. To assess trends through time, we averaged DCA scores among adjacent pairs of streams (acid forest LI1 + LI2; acid moorland streams CI1 + CI4; limed streams CI2 + CI5; circumneutral moorland streams LI6 + LI7) and measured the linear correlation between mean DCA score and sampling year. To assess whether any recovery led to assemblages that converged with those in circumneutral reference streams, we calculated similarity indices in each year between the acid moorland, limed and acid forest pairs and LI6 + LI7. Our expectation was that biological recovery from acidification in limed or acid moorland streams would increase similarity with circumneutral streams. We used Jaccard coefficients (J = c/(a + b − c)) in which a is richness in sample a, b is the richness in sample b, and c is the number of taxa in common to both, with positive values indicating similar assemblages (Bradley & Ormerod 2001). We used correlation to assess trends through time in the abundance and richness of taxa in each sensitivity group, derived from WAWS, to reveal systematic gains or losses of taxa characteristic of different pH classes. We also examined changes at the species level, using correlation to assess trends in abundance and log-likelihood ratio tests to assess changes in the frequency of species occurrence between the first and last 10 years of data collection.

Results CHEMICAL TRENDS

Mean pH increased significantly through time in all stream types at Llyn Brianne when judged on both annual and winter data (Table 1, Fig. 2), but changes were more rapid in acid moorland (0·03–0·05 pH units year−1) than circumneutral moorland (0·017 units year−1) or acid forest (0·014 – 0·012 units year−1). These trends reflect the position of each stream on the logarithmic pH scale, and in reality reductions in winter H+ concentrations were broadly similar in acid moorland and acid forest streams over the 25-year study (14·6–15·9 vs. 8·0–

Fig. 2. Trends in calcium concentrations (a, b, c) and pH (d, e, f ) of experimental stream pairs at Llyn Brianne, central Wales, between the 1980s and 2005. Shaded symbols denote streams limed in 1987/88 in either acid moorland ( ) or conifer forest ( ). Open symbols are unlimed reference streams in acid moorland () or conifer forest (). All values are annual means ± SD. Trends in aluminium concentrations (g, h, i) can be seen in Supporting Material, Fig. S1.


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Fig. 3. Trends in minimum pH at two acid moorland (CI4; CI1) two acid forest (LI1; LI8) sites at Llyn Brianne over the period 1981–2005. Curves were fitted using quadratic regression, with r2 of 0·59–0·80. Table 1. Trends in pH over the period 1981–2005 in unmanipulated streams at Llyn Brianne, Wales, modelled by linear regressions (y = a + bx) for mean values or quadratic regression for minima (see Fig. 3). Fits were calculated from 21–25 years of data (*P < 0·05; **P < 0·01; ***P < 0·001)

Stream

b (± SD)

100 · R2

Fitted mean (minimum) pH, 2005

Estimated decrease in mean H+ (μquiv L−1) (1981–2005)

LI6 (circumneutral moorland) CI4 (acid moorland) CI1 (acid moorland) LI1 (acid forest) LI8 (acid forest)

0·0167 (0·006) 0·0333 (0·007) 0·0529 (0·007) 0·0117 (0·005) 0·0142 (0·004)

27·5* 53·1*** 80·7*** 20·3* 35·1**

6·9 (6·6) 5·6 (5·2) 6·1 (4·7) 4·8 (4·5) 5·2 (4·9)

0·2 14·6 15·9 15·2 8·0

Table 2. Trends over the period 1981–2005 in dissolved aluminium concentrations in unmanipulated streams at Llyn Brianne modelled by linear regressions ( y = a + bx). Other conventions as in Table 1

Stream

B

100 · r 2

Mean (and maximum) concentrations, 2002–2005 mg L–1

LI6 (circumneutral moorland) CI4 (acid moorland) CI1 (acid moorland) LI1(acid forest) LI8 (acid forest)

–0·0013 (0·0004) –0·003 (0·0007) –0·004 (0·0005) –0·002 (0·003) –0·0004 (0·001)

37·1%** 48·6%*** 74·8%*** 1·9% 0·2%

0·046 (0·072) 0·087 (0·112) 0·031 (0·128) 0·590 (0·713) 0·340 (0·471)

15·2 μequiv L−1). pH minima also increased near-linearly in all acid streams, with trends in this case best described by quadratic regression (Fig. 3). Estimated reductions between 1981 and 2005 in average H+ concentrations during events were 31–118 μequiv L−1 (Table 1), but pH values still fell during episodes to < pH 4·7–5·1 in untreated moorland streams even at the end of the study period. Despite clear indications of chemical recovery, forest streams at Llyn Brianne were still acid by 2005, with means of

4·8–5·2 and minima 4·5 – 4·9. Here, also, aluminium concentrations still exceeded 0·34–0·59 mg L−1 on average by the end of the data run, and had not declined. This contrasts with significant reductions at both acid and circumneutral moorland sites to 0·046–0·087 mg L−1 (Table 2). In the limed streams CI5, CI2 and LI4, calcium concentrations were elevated relative to adjacent reference streams immediately following liming in 1987 and 1988, although this effect was least durable in CI2 and restricted to c. 7−9 years


Restoration and recovery from acidification 169 Table 3. Inter-annual differences between dry and wet years in H+ concentrations (μequiv L−1) after accounting for trends through time in two acid forest streams (LI8, LI1) and one acid moorland stream (CI4) at Llyn Brianne over the period 1981–2005. The values are residual mean H+ concentrations in winter (with SD; n) for years with discharge greater or less than the winter median (0·763 m3 s−1) or antecedent summer median (0·309 m3 s−1) Site

Dry years (SD; n)

Wet years (SD; n) F (d.f.)

P

Winter discharge LI8 −1·7 (2·8; 12) LI1 −2·8 (6·5; 11) CI4 −1·5 (1·8; 11)

1·6 (2·5; 13) 2·6 (7·7; 12) 1·6 (4·4; 10)

9·59 (1,23) 3·31 (1,21) 4·73 (1,19)

0·005 0·08 0·04

Summer discharge LI8 0·06 (3·4; 12) LI1 2·0 (9·2; 10) CI4 1·8 (4·80; 8)

−0·05 (2·9; 13) −1·5 (5·8; 13) −1·1 (2·04; 13)

0·01 (1,23) 1·26 (1,21) 3·65 (1,19)

NS NS 0·07

post-treatment (Fig. 2). pH increased following liming in all three treated streams relative to adjacent reference streams, although pH differences between the two limed and acid moorland streams declined towards zero over the following 12–18 years (C2 vs. C1: r = 0·90, n = 18 years post-liming, P < 0·01; C5 vs. C4 r = 0·72, n = 18 years post-liming, P < 0·01). Liming effects on aluminium were more variable among streams, reducing concentrations in LI4 and CI5, but not CI2 (Fig. 2). Climate influenced variations in runoff acidity judged from residuals around the regressions of mean H+ against study year. In particular, wetter-than-average winters increased

residual H+ concentrations on average by 3·1–5·4 μquiv L−1 relative to dry winters, with this effect detectable in both forest and acid moorland streams (Table 3). At rates of annual change over the last 25 years, this effect was sufficient to offset 21%, 35% and 41% of the 25-year recovery, respectively, in CI4, LI1 and LI8, or the equivalent of c. 5–10 years of decline in H+ concentrations. By contrast, drier-than-average summers had a weak and barely detectable tendency to increased residual H+ concentrations on average by 2·9 μquiv L−1only in the moorland stream CI4 relative to wet summers. There was no detectable effect of the NAO.

ACID SENSITIVITY AMONG INVERTEBRATES

Regional data from acid-sensitive streams across Wales in 1984 and 1995 revealed variations in the mean and minimum pH at sites occupied by invertebrate taxa that also occurred at Llyn Brianne (Fig. 4a). Species at sites with intermediate mean pH (range 5·7–6·4) were exposed to greater episodic variations than at chronically acid or circumnuetral sites (Fig. 4b). Ward’s classification identified four groups respectively of taxa typical at chronically acid sites (Group I: mean pH < 5·7; minimum pH < 5·1); acid-tolerant taxa typical at fluctuating acid sites (Group II: mean pH 5·7–6·1; minimum pH 4·5–5·3); moderately acid-sensitive taxa sometimes able to tolerate reduced minima (Group III mean pH 6·1– 6·4; minimum pH 4·9–5·7); and acid-sensitive taxa typical of circumneutral streams (Group IV mean pH 6·4 – 6·7; pH minima > 5·7) (Table 4).

Table 4. Invertebrate taxa collected at streams in upland Wales in 1984 and 1995 at sites characterized by different mean and minimum pH as classified using the method of Ward (1963) Group I (mean pH < 5·7; minimum pH < 5·1)

Group II (mean pH 5·7–6·1; minimum pH 4·5–5·3)

Group III (mean pH 6·1–6·4; minimum pH 4·9–5·7)

Group IV (mean pH 6·4 – 6·7; minimum pH > 5·7)

Coenagrionidae Velia caprai Leptophlebia sp. Nemurella picteti Platambus maculatus Leuctra fusca Hydrophilidae Diura bicaudata Oxyethira sp. Cordulegaster boltonii Chaetopteryx villosa Rhyacophila munda

Leuctra hippopus Nemoura spp. Polycentropus flavomaculatus Plectrocnemia conspersa Simuliidae Oulimnius tuberculatus Phagocata vitta Siphonoperla torrentium Amphinemura sulcicollis Chironomidae Protonemura spp. Leuctra nigra Rhyacophila dorsalis Oligochaeta Leuctra inermis Plectrocnemia geniculata Isoperla grammatica Drusus annulatus Limnius volckmari Brachyptera risi Tipulidae Gyrinus substriatus Hydropsyche siltalai Chloroperla tripunctata Elmis aenea Esolus parallelepipedus

Oreodytes sanmarkii Tabanidae Crenobia alpina Sialis spp. Perlodes microcephalus Pisidium sp. Ceratopogonidae Lepidostoma hirtum Hydraena gracilis Helichus substriatus Sericostoma personatum Diplectrona felix Rhithrogena semicolorata Ecdyonurus spp. Electrongena lateralis Hydropsyche instabilis Philopotamus montanus

Silo pallipes Ancylus fluviatilis Baetis rhodani agg. Caenis sp. Wormaldia sp. Perla bipunctata Glossossoma conformis Metalype fragilis Odontocerum albicorne Alainites muticus Serratella ignita Agapetus fuscipes Limnebius truncatellus Crunoecia irrorata Lymnaeidae Dixidae Paraleptophlebia submarginata. Helophorus spp.


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Fig. 4. Characterizing invertebrate taxa on pH tolerance using data from the Welsh Acid Waters Survey. (a) The mean pH (± 1SD) of streams from which each invertebrate taxon was collected, with pH minima for each represented by dashes (b) Difference between mean pH and minimum pH for sites occupied by each taxon. See Table 4 for the invertebrate groups derived using the classification method of Ward (1963).

INVERTEBRATE TRENDS AT LLYN BRIANNE

In the ordination of all 260 samples from all available stream– year combinations (i.e. 14 streams × 8–22 years per stream), three axes together explained 28·4% of variance in invertebrate assemblage composition (Fig. 5a). Scores on the first axis, which explained 16·5% of the variance in taxonomic composition, increased highly significantly with mean pH in the sampling year and stream in which the sample was collected (F12,202 = 11·07, P < 0·001 after accounting for ‘stream’ effects) and we assumed that increase in axis 1 scores represented a positive response to increasing pH. There was no formally significant trend in DCA axis 1 scores through time in circumneutral streams (LI6/LI7: r = 0·40, n = 21, P < 0·07), while axis 1 scores in replicate acid forest streams actually declined (LI1/LI2: r = –0·44, n = 21, P < 0·04). Nor was there any trend in axis 1 scores in the limed forest stream LI4 (r = 0·0, n = 22; data not plotted). In contrast, axis 1 scores in the replicate acid moorland streams (CI1/CI4: r = 0·58, n = 19, P < 0·007), and limed streams (CI2/CI5: r = 0·69, n = 19, P < 0·001) increased significantly (Fig. 5b). Part of this trend was explained by reductions at limed sites in the abundance or richness of some acid-tolerant Group I and II taxa (Table 5) including Plectrocnemia conspersa (r = –0·51, P < 0·05), Polycentropus flavomaculatus (r = –0·49, P < 0·05), Protonemura spp. (r = –0·60, P < 0·01), Rhyacophila dorsalis (r = –0·53, P < 0·05) and Simuliidae (r = –0·59, P < 0·01). Moderately sensitive

Fig. 5. The ordination of invertebrates in 14 streams at Llyn Brianne over the period 1981–2005, and the trends in assemblages revealed by the resulting major axis. In the top panel, the positions occupied by each stream through time are indicated by ellipses, while the lower panel plots the changing position of replicate pairs of circumneutral streams (LI6/LI7 ), acid moorland streams (CI4, CI1 ), limed moorland streams (CI5; CI2 ), and acid forest streams (LI1, LI2 ).

taxa from Group III also declined in aggregate abundance at acid moorland and limed sites with the strongest contribution from the loss of Crenobia alpina. However, limed and acid moorland sites were the only locations to gain acid-sensitive species (Group IV), and their abundance in acid moorland streams also increased (Table 5). Three different taxa contributed significantly to this pattern. In replicate acid moorland streams, Serratella ignita increased in abundance, while Wormaldia sp. increased in abundance and frequency of occurrence (log-likelihood test G = 8·32, P < 0·01). In limed streams, Alainites muticus increased in abundance and frequency (G = 4·19, P < 0·05), while Wormaldia sp. increased in frequency (G = 4·19, P < 0·05). Baetis rhodani doubled in frequency of occurrence at both limed and acid moorland sites, while Limnebius truncatellus, Paraleptophlebia submarginata and Dixidae also tended to increase, but in none of these cases could chance effects be discounted (Table 6). One moderately


Restoration and recovery from acidification 171 Table 5. Trends (correlation coefficients) in the abundance and richness of groups of invertebrates of contrasting acid-sensitivity in different stream types at Llyn Brianne, 1991–2005 (*P < 0·05, **P < 0·01, ***P < 0·001; n = 20–22 years)

Stream type Abundance Acid forest Acid moorland Limed Circumneutral Richness Acid forest Acid moorland Limed Circumneutral

Group I (most acid tolerant)

Group IV (least acid Group II Group III tolerant)

−0·14 −0·15 −0·43 −0·50*

−0·60** −0·11 −0·47* −0·47*

−0·30 −0·51* −0·57** 0·11

0·20 0·69*** 0·00 −0·09

−0·58** −0·43 −0·57** −0·36

−0·36 −0·27 −0·33 −0·27

−0·34 −0·36 −0·12 −0·03

−0·02 0·50* 0·58** 0·21

acid-sensitive species from Group III, Diplectrona felix, increased significantly at limed sites in both abundance (r = 0·61, P < 0·01) and frequency of occurrence (log-likelihood test G = 5·55, P < 0·05). Despite the trends among individual species and overall assemblage composition, Jaccard coefficients revealed no significant convergence between limed (r = –0·29, n = 19, NS) or acid moorland streams (r = –0·15, n = 20; NS) and circumneutral streams. This reflected the relatively modest proportion (four out of 29 species) of the total pool of species from naturally circumneutral streams at Llyn Brianne that have been gained by limed or acid moorland streams (Fig. 6).

Fig. 6. The overlap in taxonomic composition in different stream types at Llyn Brianne observed over 25 years. The values are the numbers of taxa in each segment, and the similarity (Jaccard index) between the stream types are J = 0·57 for circumneutral/acid moorland; J = 0·53 for acid moorland/acid forest, and J = 0·27 for circumneutral/acid forests.

Discussion The magnitude of chemical recovery from acidification across Europe and North America is increasingly being established (Stoddard et al. 1999; Evans et al. 2001; Davies et al. 2005; Eshleman et al. 2008). In the UK Acid Waters Monitoring Network (AWMN), H+ concentrations declined significantly during 1988–2002 at around one-third of the 22 sites typically

Table 6. Trends in the abundance (correlation) and frequency of occurrence (log-likelihood tests) of selected acid-sensitive invertebrates in different stream types at Llyn Brianne, 1985–2005 (*P < 0·05, **P < 0·01, ***P < 0·001; n = 20–22 years)

Acid moorland streams Alainites muticus Baetis rhodani Serratella ignita Paraleptophlebia submarginata Wormaldia sp. Agapetus fuscipes Helophorus spp. Limnebius truncatellus Dixidae Limed moorland streams Alainites muticus Baetis rhodani Serratella ignita Paraleptophlebia submarginata Wormaldia sp. Agapetus fuscipes Diplectrona felix Silo pallipes Helophorus spp. Limnebius truncatellus Dixidae

Trends in abundance

Frequency of occurrence (1985–1995)

Frequency of occurrence (1996–2005)

0·15 0·04 0·48* 0·01 0·68*** −0·34 −0·32 0·18 −0·12

1 3 0 1 0 1 4 0 1

2 6 2 1 6** 0 0 1 0

0·47* −0·04 −0·04 0·18 0·39 0·07 0·61** −0·34 −0·12 0·31 0·39

0 3 1 0 0 0 0 1 1 1 1

3** 6 0 2 3** 1 3** 0 0 4 4


172


by c. 0·2–0·6 μquiv L−1 year−1, near identical to the reductions of 0·32–0·64 μquiv L−1 year−1 recorded over 25 years at Llyn Brianne (Davies et al. 2005). Also consistent with Llyn Brianne, aluminium concentrations have fallen at almost half of all the AWMN sites. While nitrate contributes to runoff acidity locally (Curtis et al. 2005), the principal driver of recovery in the AWMN has been a reduction in non-marine sulphate, with concentrations in runoff falling significantly at 80% of the sites in direct response to declining deposition (Cooper 2005; Curtis et al. 2005). Records for non-marine sulphate at Llyn Brianne are incomplete, but values in adjacent streams on Plynlimon have declined by around 1·2– 1·5 μquiv L−1 year−1 to reduce concentrations by almost 50% from their former peak (Davies et al. 2005). At Llyn Brianne, nitrate has not increased in runoff (S. J. Ormerod unpublished data), while sulphate is the major acid anion (Kowalik et al. 2007). Large increases in concentrations of dissolved organic carbon characterize some of the AWMN, with the potential to increase organic contributions to acidity while complexing aluminium (Evans, Monteith & Cooper 2005). However, this effect would tend to offset pH increase. Moreover, at Llyn Brianne, organic contributions to acidity and episodicity are small (Kowalik et al. 2007). The best explanation for increasing mean and minimum pH, reductions in H+ and decreases in aluminium concentration at unmanipulated sites in this study, therefore, is a reduction in sulphur deposition, and the data support hypothesis 1 unequivocally. Unlike wider monitoring networks, the Llyn Brianne experiments allow an assessment of the influence of land-use on recovery in catchments within close proximity. Reductions in H+ concentration in acid forest were of approximately similar magnitude (8–15 μequiv L−1) to those in moorland, but the lower initial pH in forest streams means that pH increase during recovery has so far been slower. Differences in acidity and aluminium concentrations between forest and moorland streams detected by Stoner, Gee & Wade (1984) were thus still apparent at the end of our study period. Conifer trees at high altitude and in moderately polluted climates like Llyn Brianne increase the capture of acidity and acid anions (Fowler, Cape & Unsworth 1989; Hornung et al. 1990), and deposition is unlikely to have diminished more rapidly here than in adjacent moorland (Harriman et al. 2003). Interestingly, rates of reduction in H+ have been almost twice as rapid in LI1 (15 μequiv L−1), with its mature canopy throughout the experiment, as in LI8 (8 μequiv L−1), where the forest canopy has developed from c.10–12 years old in 1981 to 35–37 years old by the end of the current study. One explanation would be that canopy development has offset at least part of the reduction in acid inputs – the opposite of forest clearance effects observed in Scotland (Harriman et al. 2003). Superimposed over long-term recovery from acidification, chemical conditions vary in base-poor streams during rainstorms, snowmelt or other hydrological events. As well as strong acids in deposition, natural processes such as basecation dilution, sea-salt addition, the release of organic acids, and the (re-)oxidation of sulphur in organic soils can all reduce pH transiently (Wigington et al. 1996; Laudon &

Bishop 1999). As a consequence, episodic acidification still occurs widely even in streams where average pH is increasing. This situation is reflected at Llyn Brianne, where pH minima have increased markedly through time, but pH still fell during events in 2005 to at least 4·7–5·1, with aluminium concentrations reaching over 0·1 mg L−1. Here, as in other sensitive areas of Wales, episodic acidification still reflects the titration effects of strong acid anions from deposition, that is, sulphate and, to a lesser extent, nitrate (Kowalik et al. 2007). One emerging question in the assessment of recovery from acidification is whether, as strong acids in deposition decline, other drivers of episodic acidification will restrict ecological recovery in the many systems where soil base saturation has been reduced by decades of acid deposition. Such drivers include not only the natural processes outlined above, but also climatic variation. While climate-change effects on precipitation are still difficult to discern from background variability (Wilby 2006), future projections for Western European precipitation are for an overall increase in winter by 10–30% and a decline in summer by 10–20% over the next 20–40 years (United Kingdom Climate Impacts Programme http://www.ukcip.org.uk/). In turn, projections from hydrochemical models and available data suggest strong climatic interactions with recovery from acidification in Wales and other locations (Evans 2005; Evans et al. 2008). Our observations, that high winter or low summer discharge can be large enough in individual years to offset 20–41% of the 25-year recovery in H+ concentrations in some streams, are therefore interesting. These effects support hypothesis 2, although with some qualification, since effects were not formally significant in every case. The exact mechanisms at Llyn Brianne are unclear, but on Plynlimon, wet winters are characterized by decreased base-cation concentration through dilution and increasing organic acidity and aluminium during associated events (Ness et al. 2004). On similar soils in northern England, summer drought resulted in the subsequent release of sulphate oxidized from inorganic and organic sources in peats, leading to reduced pH as the soil replaced atmospheric deposition as the main sulphate source (Clark et al. 2005). Similar effects are well-established elsewhere (Eimers et al. 2007). Data on invertebrate responses to long-term recovery from acidification are scarce. At some locations characterized by pH increase, such as the southern English Broadstone stream (Woodward, Jones & Hildrew 2002), acid-tolerant taxa have declined, and this is matched closely in our data by trends in the acid-tolerant Plectrocnemia conspersa, Polycentropus flavomaculatus and Protonemura spp. Elsewhere, in the UK AWMN, there have been increases in abundance or occurrence typically of 1–3 invertebrate species per site at 11 of the 22 locations monitored (Monteith et al. 2005). Again, there were strong similarities at Llyn Briane, where four acidsensitive species increased significantly in abundance or frequency of occurrence in moorland streams irrespective of whether or not they had been limed. Although clearly consistent with recovery, the involvement of such a small proportion of the possible local pool of 29 acid-sensitive species


Restoration and recovery from acidification 173 implies that responses to increasing pH are still somehow limited (Fig. 6). It might be argued that naturally circumneutral streams are not suitable references for the recovery of acidsensitive streams because they have greater calcium concentrations (range of annual means 2·3–5·0 in LI6 vs. 0·8–1·1 mg Ca L−1 in CI4). However, overlap in species composition is otherwise large (Fig. 6). Moreover, the adults of acid-sensitive species arrive often at recovering streams, and even colonize them briefly, so that inter-basin dispersal and habitat selection by adults can be ruled out as limits on recovery (Bradley & Ormerod 2002a; Masters et al. 2007). A more likely explanation, supported by available data, is that episodic conditions in recovering streams at Llyn Brianne are still sufficient to eliminate many sensitive taxa even where mean stream pH has increased. Experiments here confirm that increased acidity (to < pH 5) and metal concentrations (e.g. Al > 100 –200 μg L−1) at high-flow cause rapid mortality in sensitive mayflies, while substantially reducing their density (Kowalik & Ormerod 2006). Across upland Britain generally, invertebrate assemblages differ significantly between streams where episodes are caused by strong acid anions as opposed to base-cation dilution (Kowalik et al. 2007). On these grounds, the long-term data from Llyn Brianne support hypothesis 3. More rapid biological recovery could soon occur if the severity of episodes continues to diminish as indicated in our data from moorland streams (Fig. 3), but conditions remain finely balanced. One of the clearest results to emerge from Llyn Brianne, here and in previous assessments, concerns the limited effects of liming as a stream restoration measure. Although treatment in three streams significantly increased pH and calcium concentrations while reducing aluminium (Bradley & Ormerod 2002a), the more prolonged record provided here shows that this effect in two moorland streams has diminished over 12– 18 years. The reasons are stream-specific: in CI2, Ca loss has been more rapid, while in CI5, pH difference from the adjacent CI4 has been reduced by ‘naturally’ increasing pH in the latter (Fig. 2e). Bradley & Ormerod (2002b) had already detected the recurrence of acid episodes at limed sites, and illustrated why this restricted the effectiveness of liming in restoring acid-sensitive assemblages. Eighteen years of postliming data now show how rates of invertebrate recovery caused by liming cannot be distinguished statistically from naturally recovering streams despite a modest early acceleration (Fig. 5). Nor can the array or rate of species gained. On these grounds, our data falsify hypothesis 4. This result is not atypical for streams limed to combat acidification, and implies that liming will not always bring expected benefits (McKie, Petrin & Malmqvist 2006; McClurg et al. 2007). In judging the success of river restoration, Palmer et al. (2005) suggested several criteria, but none involved favourable comparison with natural rates of recovery – in this case caused by emission control over larger spatial extents. We suggest that this comparison is fundamental to assessing restoration outcomes: many decisions on whether to invest in ecological restoration could depend on projecting the expected gains of local restoration against the gains of either taking no action or acting at larger spatial extents. Long-term data on which

to base such projections in ecological restoration are still extremely scarce.

Acknowledgements Many colleagues have aided data collection at Llyn Brianne over the last 25 years during projects funded by the Welsh Assembly Government, Defra, Natural Environment Research Council, Freshwater Biological Association, Forestry Commission and The Environment Agency, which carried out the chemical analysis. Dr Isabelle Durance is funded by the Daphne Jackson Trust and Natural Environment Research Council. We thank two anonymous reviewers for comments on the manuscript.

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Supporting Information Additional Supporting information may be found in the online version of this article: Fig. S1. Figure 2 g, h, i. Trends in the aluminium concentrations of experimental stream pairs at Llyn Brianne, central Wales, between the 1980s and 2005. Shaded symbols denote streams limed in 1987 and 1988 in either acid moorland ( ) or conifer forest ( ). Open symbols are unlimed reference streams in acid moorland () or conifer forest (). All values are annual means ± SD. Please note: Wiley-Blackwell are not responsible for the content or functionality of any supporting materials supplied by the authors. Any queries (other than missing material) should be directed to the corresponding author for the article.
