• aquatic insects;
  • climatic variation;
  • constancy;
  • disturbance;
  • stability


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
  • 1
     The North Atlantic Oscillation (NAO) affects marine ecosystems, lakes and some terrestrial ecosystems around the Atlantic rim, but evidence for effects on rivers is scarce. For 14 years, we sampled riverine macroinvertebrates in eight independent streams from the Llyn Brianne experimental catchments in central Wales. We assessed whether year-to-year persistence in rank abundance and species composition tracked the NAO.
  • 2
     Persistence was quasi-cyclical and varied synchronously across all streams, irrespective of chemistry and catchment land use. Invertebrate communities in all stream types or habitats switched significantly from being highly persistent during negative phases of the NAO (winter index < 1 = cold, dry winters) to unstable during positive phases (> 1 = mild, wet winters). These effects occurred in both rare and common taxa.
  • 3
     Acid episodes could not explain low persistence in positive NAO years because variations in acid-base status were not linked to the NAO. Moreover, fluctuations in persistence were apparent even in well-buffered streams.
  • 4
     Discharge in adjacent gauged catchments increased in positive NAO years by 15–18% but neither flow variability nor flow maxima were higher. Nor were variations in invertebrate persistence at Llyn Brianne directly correlated with discharge pattern. Discharge variations alone were therefore insufficient to explain links between persistence and the NAO, but we cannot exclude subtle effects due either to flow or temperature.
  • 5
     These data illustrate how the persistence of invertebrate communities varies through time in fluctuating environments. Positive phases of the NAO are accompanied by ecological instability in the Llyn Brianne streams, although the exact mechanisms are currently unclear. The effects of the NAO might confound or obscure other long-term change in rivers such as recovery from acidification or the effects of global warming.


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

The influences of medium-range climatic cycles on marine and continental ecosystems are increasingly well described (Dayton et al. 1992; Harrison 2000; Holmgren et al. 2001). By contrast with the El Niño Southern Oscillation (ENSO), however, there is less evidence for ecological patterns that reflect its Atlantic analogue – the North Atlantic Oscillation (NAO). In terrestrial ecosystems, the NAO has been linked to the phenology of birds or plants (Forchhammer, Post & Stenseth 1998; Post & Stenseth 1999; Przybylo, Sheldon & Merila 2000), the survival and demography of large mammals (Milner, Elston & Albon 1999) and the effects of northern snow cover on predator–prey interactions (Post et al. 1999a, 1999b). In marine systems, planktonic communities reflect year-to-year patterns in temperature, nutrient upwelling or fluctuating salinity driven by the NAO (Fromentin & Planque 1996; Lindahl et al. 1998; Nehring 1998; Planque & Taylor 1998; Belgrano, Lindahl & Hernroth 1999; Hagberg & Tunberg 2000; Hanninen, Vuorinen & Hjelt 2000). In lakes, the NAO affects plankton through temperature, wind-induced mixing, ice cover and altered food web dynamics (George & Harris 1985; Straile & Geller 1998; George & Hewitt 1999; George 2000; Straile 2000). Parallel and synchronous changes across large areas show that such effects are probably widespread (Straile & Adrian 2000). So far, however, and in contrast to lakes, few studies have assessed whether the NAO affects rivers and river organisms (Elliott, Hurley & Maberly 2000; Monteith, Evans & Reynolds 2000).

As the alternation of differences in atmospheric sea-level pressure (SLP) between the Azores and Stykkisholmur (Iceland), the NAO drives complex weather patterns over a cycle that is now roughly decadal and of increased intensity since the mid-1970s (Hurrell 1995; Hurrell & van Loon 1997). Hurrell’s (1995) NAO index describes these pressure fluctuations though time, negative values being accompanied by cold, dry and calm winters in NW Europe, whereas positive values are accompanied by milder winters, strong westerly winds and rainfall up to 30% above annual average (Hurrell 1995). With variations in precipitation and temperature so marked in the NAO, knock-on effects on river ecosystems might be expected. First, variations in river temperature follow closely those of air temperature (Mohseni & Stefan 1999). Secondly, for some regions rainfall variations in the NAO influence annual river discharge (Cullen & de Menocal 2000; Hanninen, Vuorinen & Hjelt 2000). Variations in discharge, in turn, can drive ecologically important changes in rivers such as floods, droughts and the dilution or mobilization of important ions (Reynolds, Emmett & Woods 1992; House & Warwick 1998; Jarvie, Whitton & Neal 1998; Bishop, Laudon & Kohler 2000). In the case of the ENSO, effects such as these influence river organisms (Puckridge, Walker & Costelloe 2000; Mol et al. 2000).

In this paper, we assess year-to-year changes among aquatic macroinvertebrates over 14 years in eight independent streams in western Britain in relation to the NAO. Following previous studies at the same sites and elsewhere, we focus on persistence as a measure of stability or environmental constancy (Townsend, Hildrew & Schofield 1987; Weatherley & Ormerod 1990a; Holomuzki & Biggs 2000). We defined persistence as a characteristic of the whole community revealed by the relative constancy in the rank abundance pattern and composition of assemblages through time (Holling 1973; Connell & Sousa 1983). Elsewhere, we have assessed long-term trends at the same sites in invertebrate abundance and richness (Bradley & Ormerod 2001).

study area

All the streams were in the experimental catchment area around Llyn Brianne reservoir, mid-Wales (52°8′N 3°45′W), described previously by Stoner, Gee & Wade (1984) and Weatherley & Ormerod (1987). The streams are all second or third order, with catchments at altitudes of 215–410 m and of 15–264 ha, consisting either of upland sheep-pasture or plantations of sitka spruce (Picea sitchensis Carriere) and lodgepole pine (Pinus contorta Douglas; Rutt, Weatherley & Ormerod 1989; Table 1). The underlying Ordovician and Silurian shales, mudstones and grits are base-poor with low buffering capacity. Overlying brown podzolic soils, stagnopodzols and peats provide limited buffering so that runoff is soft (mean total hardness 3·9–18·8 mg CaCO3 L−1) and in some cases acidified (mean pH 4·6–6·9; Rutt et al. 1989). The streams form groups classifiable as acid conifer (LI1, LI2, LI8), acid moorland (LI5, CI1, CI4) and circumneutral moorland (LI6, LI7), between which faunal communities differ substantially (Weatherley & Ormerod 1987; Table 1). Typically, the circircumneutral moorland streams are species-rich and characterized by ephemeropterans, trichopterans and plecopterans, whereas acid forest streams are species-poor and dominated by acid-tolerant plecopterans. Acid moorland streams are intermediate between these extremes. The catchments of three further study streams (L14, C12 and C15) were limed artificially over a decade ago, but these are not considered further here (Weatherley & Ormerod 1990a; Bradley & Ormerod 2001).

Table 1.  Catchment characteristics and chemistry for the study streams at Llyn Brianne. Chemical determinands are long-term annual means from 1985 to 1998 (adapted from Rutt, Weatherley & Ormerod 1989)
Site codeCatchment land useCatchment area (ha)Mean pHTotal hardness (mg CaCO3 L−1)Filterable aluminium (mg L−1)
LI1Acid stream in conifer forest (c. 40 years old)2644·9 7·00·40
LI2Acid stream in conifer forest (c. 40 years old)1944·9 6·90·43
LI8Acid stream in conifer forest (c. 25 years old)1125·4 7·90·23
LI5Acid stream in moorland 666·0 8·50·04
CI1Acid stream in moorland 155·2 3·90·10
CI4Acid stream in moorland 715·5 5·40·12
LI6Circumneutral stream in moorland 826·915·70·05
LI7Circumneutral stream in moorland 726·918·80·04

The climate at Llyn Brianne is temperate, with mean stream temperatures rarely outside the range of 0–20 °C, and mean annual rainfall c. 1900 mm (Weatherley & Ormerod 1990a, 1990b). Stream substrata vary between gravel (> 2–16 mm) and bedrock, with bryophytes the only submerged macroflora; in marginal areas, Juncus spp. and Sphagnum spp. form the most abundant vegetation (Rutt et al. 1989; Weatherley & Ormerod 1990a).


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References


All streams were kick-sampled for invertebrates each April from 1985 to 1998, except 1991 (a year without funding), using a standard net (0·9 mm mesh; 230 × 255 mm frame). At each site a 2-min sample was taken from mid-channel riffles, and a 1-min sample taken from stream margins by a combination of kicking and sweeping upstream in marginal vegetation (Rutt, Weatherley & Ormerod 1989). This sampling strategy was derived following extensive quantitative sampling throughout several years and reflects communities for the entire winter–spring periods; paired samples in margins and riffles record in excess of 75–85% of the species present in any stream at a given time (Weatherley, Rutt & Ormerod 1989 and unpublished data). Samples were preserved in 70% alcohol on site prior to sorting and identification in the laboratory to the lowest feasible taxonomic level. This involved species except for Oligochaeta, Chironomidae, Simuliidae, Tipulidae, Dixidae, and the early instars of other groups, which were identified to the generic or family level. When there was any doubt about the correct identification of a species we combined all possible species at the generic or family level for the purposes of analysis.


Aluminium (mg/L), calcium (mg/L), total hardness (mg/L) and pH were monitored weekly or fortnightly from 6 January 1984 to 24 November 1998 as measures of acid-base status. Weatherley & Ormerod (1987) detail the analytical procedures involved. The data provided mean, maximum, minimum and the range for each determinand in each stream, either for 14 whole years, for summer (April–September) or for winter (October–March). We made this seasonal separation as the climatic influence of the NAO is highly seasonal (Hurrell 1995). Annual values were calculated from April to the end of March the following year, thereby describing conditions antecedent to invertebrate sampling in early April. Continuous hydrological data were not available for the Llyn Brianne catchments, but measures of discharge (cumecs) were made continuously in two directly adjacent streams, the Afon Hafren and Afon Gwy, at the Centre for Ecology and Hydrology Plynlimon experimental site. These streams are second and third order, at similar altitudes to the Llyn Brianne streams and are less than 40 km directly to the north. Intercalibration of sites across Wales indicates that generally similar temporal patterns of discharge occur in streams throughout this region (Stevens et al. 1997). Therefore, it can be expected that the Hafren and Gwy will show similar year-to-year discharge regimes to the Llyn Brianne streams. These data were arranged seasonally exactly as the chemical data.

Data for the North Atlantic Oscillation (NAO) index were obtained online from sources also used elsewhere (Hurrell 1995; Monteith et al. 2000). The NAO index is calculated from the difference of normalized sea level pressures (SLP) between the Azores and Iceland (Hurrell 1995; Belgrano, Lindahl & Hernroth 1999). The NAO indices were arranged as yearly means and winter (December–March) means because the NAO influences precipitation and temperature largely during the winter (Hurrell 1995; Fromentin & Planque 1996; Forchhammer, Post & Stenseth 1998; Lindahl et al. 1998; Planque & Taylor 1998; Belgrano et al. 1999).


Following standard methods (Townsend et al. 1987; Weatherley & Ormerod 1990a), invertebrate persistence was measured within streams in two ways. First, to reflect patterns in relative abundance between pairs of successive years, we determined Spearman’s rank correlation coefficients between the abundances of all species in the two communities being assessed. Values range from −1 indicating low persistence to +1 indicating high persistence. Secondly, persistence in species composition was calculated as the Jaccard’s coefficient of similarity (J; see Hellawell 1978):

inline image

where a= number of taxa in community A (first year in any pair), b= number of taxa in community B (the second year in any pair) and j= number of taxa found in both communities. Values range from 0 (no similarity, low persistence) to 1 (identical similarity, high persistence). In both cases (Jaccard and Spearman’s ranks), values were determined for the total stream community and also for riffle (R) and margin (M) samples separately. We calculated persistence for the 15 most abundant taxa overall, and for the remaining ‘rare’ taxa in each community to assess whether the latter were particularly sensitive to change. In addition to pairwise comparisons between successive years within streams, we determined persistence measurements for year pairs of gradually larger and larger intervals to assess any long-term drift (i.e. year pairings of 1, 2, 3 and up to 14 years apart), but none was found.

Measures of persistence were related using Pearson product-moment correlation to the NAO, to chemical variables and to discharge in the antecedant winter (see Townsend et al. 1987; Weatherley & Ormerod 1987). We also compared measures of persistence between years with positive (> 1) and negative (< 1) NAO winter indices using one-way analysis of variance (anova). Prior to these analyses, chemical variables and taxon abundances were log-transformed and Spearman’s rank correlation coefficients were z-transformed to normalize variances (Weatherley & Ormerod 1990a).


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

patterns in persistence

Taxon abundance, taxon richness and measures of persistence varied from year to year in all streams. Mean total invertebrate abundance in any year ranged from 64 (±18 SE individuals per sample) to 822 ± 99, while mean richness across streams ranged from 8 ± 1 to 26 ± 2. Average persistence in rank abundance of the total faunal community for each stream between successive years ranged from 0·70 ± 0·10, indicating high persistence, to 0·02 ± 0·40, indicating low persistence. Similarly, Jaccard’s coefficients ranged from 0·74 ± 0·08, indicating similar taxonomic composition between years, to 0·27 ± 0·08, indicating low similarity.

These variations were not random, but instead appeared to be quasi-cyclical (Fig. 1). Variations occurred synchronously across all streams so that persistence in rank abundance (all r > 0·7, P < 0·01) and Jaccard’s coefficients (all r > 0·9, P < 0·01) were significantly intercorrelated between the three stream groups (acid forest, acid moorland, circumneutral moorland). Variations also occurred synchronously for abundant and rare taxa (Fig. 2), and for both stream habitats sampled (Fig. 3). Together, these data illustrate that large magnitude variations in persistence were a whole-stream and whole-community feature of the long-term trends at Llyn Brianne over the study period, and characterized all stream types.


Figure 1. Year-to-year trends in community persistence in streams at the Llyn Brianne experimental catchments between 1985 and 1998. The values are means for each stream group (see Table 1) and show (a) Spearman’s rank correlation on rank abundance across species (r) and (b) similarity in taxonomic composition (J). The dashed lines represent missing data interpolated for 1991.

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Figure 2. Year-to-year trends in community persistence in streams at the Llyn Brianne experimental catchments between 1985 and 1998 for abundant taxa (the 15 most abundant species overall) and rare taxa (the remainder). The values are means for all streams ± SE. Other conventions as in Fig. 1.

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Figure 3. Year-to-year trends in community persistence in streams at the Llyn Brianne experimental catchments between 1985 and 1998 for riffles and marginal habitats. The values are means for all streams ± SE. Other conventions as in Fig. 1.

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relationships with the nao

The NAO characteristically fluctuates between extreme positive and negative values (Hurrell & van Loon 1997; George & Hewitt 1999), and such fluctuations were apparent during the 14 years of this study. Moreover, concordance between the phases of the NAO and patterns of invertebrate persistence were striking (Fig. 4). For invertebrate communities in all stream types, all habitats, and of either rare or common occurrence, index values were significantly higher following negative phases of the winter NAO (< 1) than following positive NAO values (> 1) (Table 2). In the case of rare taxa in all streams, and all taxa in circumneutral streams, stability and persistence declined linearly with increasing values of the NAO winter index (Table 3).


Figure 4. Year-to-year trends in community persistence in streams at the Llyn Brianne experimental catchments between 1985 and 1998 as shown by Spearman’s rank correlation on rank abundance across species (▴). The values are means for all streams ± SE and the dashed lines represent missing data interpolated for 1991. Also shown is the NAO winter index (inline image after Hurrell 1995).

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Table 2.  Variations between values of persistence or stability among invertebrates at Llyn Brianne in years between 1984 and 1998 characterized by either positive (NAO > 1) or negative (NAO < 1) values of the NAO winter index
 NAO < 1NAO > 1F1,9
  • ** 

    P < 0.01;

  • *** 

    P < 0.001.

Jaccard coefficient
 All streams
  All taxa0·64 0·3825·18***
  Abundant taxa0·89 0·6322·46***
  Rare taxa0·38 0·1423·69***
  Riffles only0·62 0·3623·42***
  Margins only0·54 0·3412·37**
  Circumneutral streams0·64 0·3834·42***
  Acid moorland streams0·57 0·3510·88**
  Acid forest streams0·64 0·3419·64**
 Spearman’s rank
  All streams
  All taxa0·61 0·2426·16***
  Abundant taxa0·56 0·3412·37**
  Rare taxa0·02−0·4220·73***
  Riffles only0·60 0·2328·73***
  Margins only0·40−0·0427·24**
  Circumneutral streams0·75 0·3138·50***
  Acid moorland streams0·58 0·1011·89**
  Acid forest streams0·68 0·1515·17**
Table 3.  Pearson product-moment correlations between persistence, stability and the NAO winter index for all streams for individual stream types at Llyn Brianne between 1984 and 1998 (*P ≤ 0·05)
 Persistence (all taxa)Stability (all taxa)Persistence (rare taxa)Stability (rare taxa)
All streams−0·56−0·56−0·64*−0·62*
Acid forest streams−0·58−0·53−0·64*−0·59
Acid moorland streams−0·52−0·49−0·39−0·30
Circumneutral moorland streams−0·63*−0·63*−0·61*−0·66*

Direct measures of total invertebrate abundance or taxon richness were unrelated to the NAO. However, four species declined significantly with the NAO winter index (Nemurella picteti Klapalek, r = −0·793, P < 0·01; Elmis aenea Muller r = −0·575, P < 0·05; Hydropsyche siltalai Dohler r = −0·830, P < 0·01; Paraleptophlebia submarginata Stephens r = −0·944, P < 0·001) whereas Chloroperla tripunctata Scopoli increased (r = 0·834, P = 0·001).

Possible mechanisms through which the NAO might affect invertebrate stability at Llyn Brianne were unclear from the data collected. The annual average discharge of the Hafren (r = 0·645, P < 0·05) and Gwy (r = 0·56, P < 0·05), and the average winter discharge of the Hafren (r = 0·57, P < 0·05) increased during positive phases of the NAO winter index by 15–18%. However, discharge ranges, discharge variability (as coefficients of variation) and discharge maxima were not significantly different in positive NAO years from other years. More directly, there were no significant correlations between discharge in the Gwy or Hafren and any measure of persistence in the Llyn Brianne streams. Nor were there any consistent relationships between any measures of acid-base status (pH, calcium or aluminium concentrations) and either the NAO or measures of persistence.


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

Several studies have shown that environmental factors act across catchments to influence the persistence of river invertebrates (Townsend et al. 1987; McElravy, Lamberti & Resh 1989). Weatherley & Ormerod (1990a) observed such supra-catchment effects at the sites in this study from 6 years’ of data, but at the time no explanation was available. This longer run of 14 years’ data shows not only that synchronous fluctuations in persistence have continued, but also that they are consistent across contrasting catchments, taxa and stream habitats in the Lynn Brianne experimental area. On the basis of correlative evidence, the North Atlantic Oscillation provides a region-wide effect of the type required to explain the observed pattern: low year-to-year persistence in rank abundance and composition in all catchments was associated with positive phases of the NAO winter index while high persistence was associated with negative phases. Although this study is not yet long enough to link stream invertebrate communities unequivocally to the NAO through several cycles (cf George & Harris 1985; Straile & Geller 1998; George & Hewitt 1999; George 2000; Straile & Adrian 2000; Straile 2000), these data are the first to show that a link might exist. So far, evidence that the ENSO affects river organisms is from fish rather than invertebrates, and implicates droughts rather than floods (Puckridge, Walker & Costelloe 2000; Mol et al. 2000). At these Welsh sites, persistence among invertebrates was lower in wet phases of the NAO than in dry phases.

Fluctuations in persistence at Llyn Brianne involved two different ecological effects. Spearman’s rank coefficients showed that relative abundances across species were more constant during cooler, drier phases of the NAO, but dissimilar during warmer and wetter phases. In turn, Jaccard’s coefficients showed that high and low persistence in rank abundance were accompanied, respectively, by high and low constancy in species composition. These effects might well be linked – for example if shifts in relative abundance were sufficiently large to reduce some species to undectable levels. We can discount sampling error as an explanation for these trends, which in other long-term studies can explain apparently significant community variation between census years (Arnott et al. 1999). First, sampling error would be randomly distributed across samples, years and streams, rather than being responsible for highly synchronous variations across streams and habitats. Secondly, sampling error among aquatic invertebrates is usually greatest among rare taxa and complex habitats, where sampling is most difficult (Parsons & Norris 1996; Cao, Larsen & Thorne 2001). In our case, these effects were apparent because communities from structurally complex marginal habitats and rare species had the lowest overall persistence (Weatherley & Ormerod 1990a). Nevertheless, year-to-year variation in persistence among rare taxa and margins echoed exactly persistence in common taxa and riffles. We conclude that changes in persistence through time at Llyn Brianne have been real.

Few studies have assessed persistence among river invertebrates, particularly over timescales similar to this study. Although none have linked patterns with climatic cycles, those available provide valuable insight into the relationship between persistence and environmental variation. In general, persistence is greatest where environmental conditions are relatively constant (Robinson, Minshall & Royer 2000) and where taxa are adapted to the prevailing environmental regime (Miller & Golladay 1996). For example, in the United Kingdom, Townsend, Hildrew & Schofield (1987) showed that persistence was greatest where streams had low discharge, constantly low summer temperatures and pH regimes that were acid and stable. By contrast, persistence is least where environmental conditions fluctuate or are characterized by pulse, press or ramped disturbances (Meffe & Minkley 1987; Lake 2000). Those disturbances known to influence persistence include changes in catchment character – for example the replacement of seminatural forest by agricultural development (Brewin, Buckton & Ormerod 2000) – or particular catchment-scale events such as pesticide use (Hutchens, Chung & Wallace 1998) and fires (Richards & Minshall 1992). Changes in flow conditions also affect persistence, for example where regimes shift from long-term stable to short-term unstable due to freezing conditions or floods (Matthaei, Uhlinger & Frutiger 1997; Bradt et al. 1999). Flow effects like this occur at a variety of scales from the whole stream down to the individual patch (Death 1996; Matthaei & Townsend 2000). On all of this evidence, variations in persistence at Llyn Brianne between different climatic phases of the NAO would be consistent with a link between persistence and environmental variability.

With the NAO influencing western European rainfall, and discharge variation in turn affecting the persistence of invertebrates, it might be expected that the effects of flow would be central to our results. No discharge data were available directly from the study sites, although average discharge in rivers just 40 km away was indeed larger in positive NAO years by up to 18%. However, extreme flows were no larger, and discharge was no more variable in positive NAO years than in negative NAO years. Thus, on our evidence low persistence cannot have reflected the effects of pronounced floods (Lake 2000). More importantly, there were no direct correlations between discharge pattern and indices of persistence in the Llyn Brianne streams. Therefore, direct discharge effects cannot alone be sufficient to explain varying invertebrate persistence. We cannot discount subtle effects since the influences on flow on aquatic invertebrates are many and varied (Hart & Finelli 1999; Lancaster 1999; Holomuzki & Biggs 2000; Doisy & Rabeni 2001). Those species which increased or decreased at Llyn Brianne under different phases of the NAO provide clear examples: local density and net-spining activity in the caddis Hydropsyche siltalai reflects current velocity (Statzner & Bretschko 1998); Paraleoptophlebia submarginata is a marginal specialist characteristically occurring in areas of low hydraulic stress (Mobes-Hansen & Waringer 1998); Elmis aenea belongs to a group with well-characterized preferences on current velocity (Dietrich & Waringer 1999), and Nemurella picteti shows clear microdistributional responses to changing flow conditions (Lancaster & Hildrew 1993). We suggest that further assessment of discharge and flow pattern might well explain how invertebrate communities respond to the NAO.

Variations in discharge not only have direct effects on stream organisms, but also indirect effects through changes in stream chemistry. As Townsend et al. (1987) noted, persistence is greater in stable chemical environments. At soft-water sites such as those in this study, acid episodes are potentially important responses to flow that result from increased base-cation dilution, and from increased titration effects due to anions mobilized from catchment soils (Bishop et al. 2000). There is clear experimental evidence that acid episodes in some Llyn Brianne streams affect stream invertebrates (Ormerod et al. 1987) and might still offset biological recovery from acidification (Bradley & Ormerod 2001). However, several lines of evidence show that fluctuations in invertebrate persistence due to the NAO were not related to episodicity. First, some of the most pronounced variations in persistence occurred in circumneutral moorland streams – where buffering is greatest, and pH never falls below 6–6·5. Acid episodes therefore do not offer the regionwide effect necessary to explain our data. Secondly, no measure of acid-base status across years in any stream type was linked to the NAO. Thirdly, measures of persistence were not correlated with either pH, aluminium or calcium concentration. Other aspects of stream chemistry vary with discharge in ways that might be important ecologically, for example because interactions between droughts, floods and temperature affect the mobilization of nitrogen and phosphorus from catchment soils (Reynolds, Emmett & Woods 1992; House & Warwick 1998; Jarvie, Whitton & Neal 1998). Across the 22 lakes and streams in United Kingdom Acid Waters Monitoring Network, negative phases of the NAO through 11 years have been associated with increased nitrate concentrations in surface waters (Monteith et al. 2000). In turn, low temperature at the surfaces of catchment soils appear to result in increased nitrate loss into runoff, perhaps because plant or microbial retention is lower (Monteith et al. 2000; B. Reynolds, unpublished data). At present we have no data on whether such effects on nutrient release are widespread, nor on what their influence on stream productivity or invertebrate stability might be.

In addition to rainfall, NAO cycles are also reflected in temperature variations due to air movements from different sources, and because cloud cover or atmospheric dust affect radiation budgets (Moulin et al. 1997; Forchhammer et al. 1998). Such temperature variations affect upland British rivers enough to affect the emergence times of salmonid fish (Elliott et al. 2000). Thermal regimes, in turn, affect invertebrate persistence (Townsend, Hildrew & Schofield 1987). At Llyn Brianne, variations in stream temperatures between months, years and catchments reflect air temperature and insolation, with variations sufficient to affect the emergence periods of some stream insects (Weatherley & Ormerod 1990b). Effects are subtle, however, and there is so far no evidence that they affect communities. Among the species responding to the NAO at Llyn Brianne, some such as Chloroperla tripunctata have clear thermal tolerance, but in this instance increased abundance during warm phases would be contrary to expectation in this cold-water species (Elliott 1988). Nevertheless, increased winter temperatures affect the timing of insect lifecycles and oviposition success of adults (Chen & Folt 1996), so that warm phases of the NAO could translate into population performance and hence low measures of persistence between years. As with examination of flow responses to the NAO, we suggest that temperature effects should figure in the search for processes linking variations in river invertebrate communities to the NAO.


These results provide further evidence that ecosystems in northern and western Europe are affected by fluctuations in the NAO and its associated climatic effects. In addition to lakes and stream fishes, our data show that the NAO probably affects stream invertebrates. In keeping with the well-known challenges of testing hypotheses about large-scale and long-term effects on whole ecosystems, our data are of necessity correlative and require further investigation of the processes involved (Manel, Buckton & Ormerod 2000; Ormerod & Watkinson 2000). Nevertheless, apparent effects of the NAO occurred in replicate streams of contrasting chemistry and catchment land-use and might therefore be widespread.

Pronounced variations in invertebrate persistence that follow the NAO have both fundamental and applied importance. In fundamental terms, these data not only confirm previous ideas that persistence in invertebrate communities reflects environmental variability, but also show that persistence can vary through time within the same river system: constancy is not a fixed property of a given location. In applied terms, our data confirm Weatherley & Ormerod’s (1990a) view that variations in aquatic communities risk confounding or obscuring the effects of other long-term trends such as recovery from acidification, eutrophication or the effects of climatic change (Chen & Folt 1996; Lancaster et al. 1996; Lawlor et al. 1998; Monteith, Evans & Reynolds 2000). We recommend that researchers designing long-term monitoring programmes consider such effects carefully.


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References

We are grateful to all current and past members of the Catchment Research Group who were involved in compiling this data set funded variously by the Environment Agency, Welsh Assembly, Forestry Commission and the Department of Environment, Food and Rural Affairs, which also funded this analysis. The Environment Agency, Wales conducted chemical sampling. We wish to thank particularly Mark Bishop and Simon Wyke for providing the chemical data, and Brian Reynolds of CEH Bangor and David Biggin of CEH Wallingford for allowing access to the discharge data from Plynlimon. Brian Reynolds, Dave Raffaelli and two anonymous referees also provided helpful comments on the manuscript.


  1. Top of page
  2. Summary
  3. Introduction
  4. Methods
  5. Results
  6. Discussion
  7. Acknowledgements
  8. References
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