Assumptions and caveats
A fundamental assumption in this work was that linear correlations among macroinvertebrates and climate represent mechanistic links that provide a basis for projection. There are parallels in terrestrial ecosystems, where climate-change effects are increasingly predicted from intercorrelations in space among climate, species distribution and population processes (e.g. Huntley et al., 2004; Thuiller et al., 2006). Such approaches are simple, and can generate hypotheses about processes testable at other scales. They also have the ability to inform management. Research into another long-term aquatic problem, acidification, illustrates these strengths: when combined with catchment-scale experiments, long-term data and process studies, empirical models increased understanding while guiding mitigation (Ormerod et al., 1988; Hindar & Wright, 2006). Similar outcomes are likely in climate change research.
Despite basing our projections on data collected over 25 years, uncertainties in extrapolating future climatic effects on streams are real. The future frequency or magnitude of extreme floods or droughts will differ from current regimes (Huntington, 2006). Additionally, change in stream temperatures over the next 20–50 years will be more sustained than the interannual variations assessed here. In addition to direct effects on existing species, such changes could promote invasion by exotic or lowland species to result in gains, as well as losses. Our ecological projections could, thus, be conservative. Conversely, however, projected rates of climatic adjustment among macroinvertebrates at Llyn Brianne might be considered typical because they reflect real long-term data.
A further uncertainty in projecting climate change effects at Llyn Brianne is that no variable explained more than 37% of the interannual variance among macroinvertebrates (Table 5; Fig. 4). Neither multiple predictors nor alternative procedures improved model fit. Unavoidable sources of nonclimatic variation included sampling effects and variations in stream chemistry. The first of these was reduced by examining trends averaged between paired streams, effectively doubling sampling effort. The second was minimized by examining trends separately in acidic and circumneutral streams. pH in acid forest, acid moorland and circumneutral streams at Llyn Brianne increased between 1981 and 2005, respectively, by 0.3, 0.8 and 0.4 units (S. J. Ormerod et al., unpublished data), but in no case did acid–base status explain variation additional to that explained by climate. Nevertheless, potential interactions between the NAO, climate change and acid–base variations may be important. Effects include short-term episodic events that would be undetected at the annual resolution of this study (Kowalik & Ormerod, 2006).
Effects of the NAO
Positive phases of the NAO at Llyn Brianne apparently reduced stability among macroinvertebrates across streams and habitats, as shown from a shorter data run (Bradley & Ormerod, 2001). There were also systematic variations in assemblage composition in acid moorland streams (see Fig. 3). The exact mechanisms are unclear, but thermal, hydrological and hydrochemical processes are possibilities. Available data indicate consistent NAO effects over larger areas of western Britain. For example, relationships between stream temperature and the NAO at Llyn Brianne (b=0.20–0.24; r2=0.35) were similar to both Plynlimon in central Wales (b=0.32–0.34x; n=7 years; r2=0.51–0.57) and Black Brows Beck in NW England (b=0.285x; n=29 years; r2=0.42) (Elliott et al., 2000; Briers et al., 2004). Effects also reach eastwards in central Europe (Webb & Nobilis, 2007). As well as affecting stability, temperature change linked to the NAO alters the development of salmonid fishes and insects, with potentially large consequences for phenology and assemblage composition (Elliott et al., 2000; Briers et al., 2004).
NAO effects on estimated winter discharge at Llyn Brianne were barely detectable over the duration of the study, although flows increase markedly elsewhere in NW Europe during positive phases (Bouwer et al., 2006). Over shorter timescales, links between discharge at adjacent Plynlimon sites and the NAO were clear (Bradley & Ormerod, 2001). The NAO also causes flow-dependent variations in the chemistry of base-poor streams on Plynlimon, reducing base-cations and increasing aluminium during positive phases (Ness et al., 2004). These effects are further complicated by increased sea-salt deposition on organic acids and H+ (Evans, 2005). In acid streams at Llyn Brianne, winter mean pH in acid moorland streams declines following sustained positive phases of the NAO on average by ca. 0.29 pH units, sufficient to offset almost 10 years of recovery from acidification at current rates (S. J. Ormerod et al., unpublished data). Five of the species least likely to occur during positive NAO phases were acid-sensitive taxa such as Wormaldia sp., Baetis muticus and O. sanmarkii. Such species are probably lost from acid-sensitive streams at Llyn Brianne due to low pH events (Kowalik & Ormerod, 2006).
Composite thermal and hydrochemical effects on stream organisms would support suggestions that the NAO predicts climatic effects synoptically on organisms more effectively than individual variables (Stenseth et al., 2002). Using the NAO as such a predictor at Llyn Brianne showed how future climatic change effects might be compounded during positive NAO phases by reduced interannual stability or altered assemblage composition (Fig. 6).
Effects of long-term climatic change
There were no long-term trends in discharge at the study sites, implying that directional climate change did not alter winter runoff during the study. This accords with current analysis of the duration required for climate change effects on discharge to become clear (Wilby, 2006). Nor did interannual variations in winter discharge affect macroinvertebrates. Discharge effects on stream macroinvertebrates elsewhere can be pronounced, but depend on refuge availability and bed sensitivity to disturbance (Townsend et al., 1997). Organisms are most affected in finer sediments, unlike the stony headwaters at Llyn Brianne. Other possible explanations for the lack of discharge effect are that (i) macroinvertebrates in these headwaters are resilient against naturally large flow variations; (ii) effects were masked by the effects of temperature or (iii) discharge parameterization (as winter Q5) was insufficient to capture ecological effects either at other times of the year (e.g. summer drought) or through flow effects that are threshold- or duration-dependent.
In contrast to discharge, long-term and directional variations in estimated stream temperature were distinct from the NAO, and associated with detectable ecological effects. Directional winter warming in these headwaters is apparently bigger than can be explained by the effects of progressive positive NAO amplification between the 1960s and 1990s (Hurrell et al., 2003).
Reliable long-term data on stream temperature are scarce, but the recent increase at Llyn Brianne is reflected elsewhere. In the upper Danube, temperatures have increased by 1.4–1.7°C since 1901, with much of this increase occurring form 1950–1970 onwards (Webb & Nobilis, 2007). Temperature in the upper Rhone increased by ca. 1.5°C between the late 1970s and 1990s (Daufresne et al., 2004), while spring temperature in the Girnock (NE Scotland) increased by 1.45°C between 1968 and 1997 (Langan et al., 2001). January temperatures in the River Pulham on Exmoor, within 120 km of Llyn Brianne, increased by 1.5°C between 1977 and 2004 (B. W. Webb, unpublished data). At Llyn Brianne, trends contrasted between forest and moorland streams, with the latter following air temperature most closely (Table 2). In larger streams, evaporative cooling, thermal mass and groundwater moderate temperature variations so that slopes linking air to water temperature are <1.0 (Caissie, 2006). In smaller glacially fed streams, meltwater effects also damp temperature variations with air temperature (Webb & Nobilis, 2007). In the small, cooler and dominantly rainfed moorland streams at Llyn Brianne, however, these effects are negligible, implying that sensitivity to atmospheric warming will be marked. By contrast, month-to-month and interannual temperature variations were damped in forest streams (Tables 2 and 3). Increasing riparian tree cover is an option for local mitigation of future heating effects on headwaters, a notion supported by extensive data (Caissie, 2006). Summer maxima might be reduced, but winter minima elevated.
The apparent effects of increasing stream temperature were twofold, with assemblage composition changing moderately while macroinvertebrate abundances during April declined. Both effects occurred only in circumneutral streams, where macroinvertebrate abundances and richness were greatest. One possibility is that acidification, in impacted streams, overrides climatic effects by simplifying assemblages and reducing richness (see Table 4; Bradley & Ormerod, 2002a; Kowalik & Ormerod, 2006). Warmer, wetter climates, in turn, might exacerbate acidification effects by altering N release, increasing organic acidity, diluting base-cations and offsetting recovery (e.g. Wright et al., 2006). Additionally, circumneutral streams are dominated numerically by Ephemeroptera, while acid streams are dominated by Plecoptera. These groups contrast in developmental sensitivity to temperature (Weatherley & Ormerod, 1990; Briers et al., 2004). Circumneutral streams contrast further from acidic streams in supporting salmonids and other predators such as riparian birds, with potential consequences for food-web function (Ormerod & Tyler, 1991).
Consideration of thermal effects on macroinvertebrate numbers requires caution since good abundance estimates depend notoriously on large samples (Needham & Usinger, 1956). The variations with temperature identified here reflected 25 years of samples in paired streams, and so are difficult to dismiss as sampling error. Moreover, trends through time were consistent with the effects of increasing temperature, albeit at P<0.1. Reductions in density also parallel those detected during experimental stream warming (Hogg & Williams, 1996) or following periods of extreme weather (Mouthon & Daufresne, 2006). However, recent quantitative sampling of macroinvertebrates in LI6 suggests that year-round densities (geometric mean 2067 m−2; 2002/2003) are not dissimilar from the 1980s (geometric means 629–1374 m−2; 1985–1988) (Weatherley et al., 1989; Kowalik & Ormerod, 2006; S. J. Ormerod et al., unpublished data). Reductions in abundance over the 25 years of the study therefore appear to be restricted to the spring period represented by April sampling.
Candidate mechanisms likely to reduce spring macroinvertebrate abundance at higher temperatures include alterations in emergence phenology or energy flow. Directional climate change over the period 1981–2005 has altered detectably the emergence phenology of amphibians in adjacent ponds (Chadwick et al., 2006), and effects on stream macroinvertebrates are likely (Briers et al., 2004). For example, a 1.7°C rise in mean winter temperatures between ca. 3.1 and 4.8°C over 25 years at Llyn Brianne could increase the average specific growth-rate of overwintering Baetis rhodani, the most abundant insect in circumneutral streams, from 0.78% to 1.02% day−1 (Elliott et al., 1988). Critically, any nymphs already at 1 mm or larger by 1 October in the preceding year would then attain sufficient size (ca. 9 mm) to emerge as adults before 1 April. Effects on abundance would depend on complex interactions among these losses through emergence, as well as any thermal effects on the hatching, development, survival and detectability of the subsequent spring–summer cohort (Elliott et al., 1988). Thermally mediated energetic effects in streams are also complex, but include increased predation by fish as temperatures increase (Kishi et al., 2005), as well as increased loss rates of litter to decomposition (e.g. Lepori et al., 2005). Both can reduce macroinvertebrate numbers demonstrably. Experimental evidence from Llyn Brianne shows how shredder numbers depend on litter supplies (Dobson et al., 1995).
Variations with temperature in the macroinvertebrate composition of circumneutral streams mostly affected less common species that occurred under warm or cool extremes (Fig. 5). One exception was D. annulatus, a relatively abundant trichopteran that was also sensitive to elevated temperature in a German stream (Wagner, 2005). The occurrence of B. niger and H. instabilis at higher temperatures is consistent with their more typical downstream distribution in warmer waters and implies some potential for invasion from other basins (Hildrew & Edington, 1979; Masters et al., 2007). However, larger numbers of commoner ‘core’ species had wider temperature amplitudes, and persisted through interannual variations. This pattern might be expected in a relatively high-latitude location such as western Britain, where many species have larges latitudinal and thermal ranges than at lower latitudes (Addo-Bediako et al., 2000).
When projected under future climates, trends with temperature in the abundance and composition of macroinvertebrates at Llyn Brianne were substantial. While many core species would persist even if gains reached 3°C, a net loss of four to 10 taxa might occur through local extinction. This is equivalent to 10–25% of typical mean richness, or 5–12% of the species pool in circumneutral streams. Outcomes from such modelling exercises can be variable (e.g. Thuiller et al., 2005), but these values from Llyn Brianne are remarkably similar to the potential loss of 6–11% of 1200 plant species from a theoretical portfolio of European protected areas over the next 50 years (Araújo et al., 2004). Reduced richness in river macroinvertebrates at increasing temperature also follows trends detected elsewhere (Mouthon & Daufrense, 2006). Such losses would have major conservation significance, particularly if obligate cooler-water species were lost from their limited, higher altitude range (Daufresne et al., 2004). Reduced macroinvertebrate abundance in streams will also affect energy transfer to predators during critical periods of annual reproduction (e.g. Ormerod & Tyler, 1991), while potential functional consequences associated with energy processing are being assessed (I. Durance et al., unpublished data).
Overall, these data suggest that the ecological consequence of climate change for upland streams could be far reaching, with effects greatest in the most species-rich locations. Our strongest recommendations are, therefore, for (i) an increased commitment to the longer-term assessment of climate change effects on headwater organisms and the processes affecting them and (ii) increased research into measures for adaptation and mitigation with specific focus on stream ecosystems.