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Low-gradient rivers flowing through the agricultural and urban landscapes of north-west Europe have long been subjected to intensive management (Purseglove 1988; Moss 1998; Rackham 2000). Probably more than 95% of lowland river channels in south-east England and Denmark have been modified to enhance land drainage, river navigation and flood prevention (Iversen et al. 1993; Brookes 1995). As a result, many have highly simplified and uniform channels, unnaturally steep banks and little dynamic connectivity with their flood plains.
River modification accelerated in the twentieth century, largely associated with the intensification of agriculture, when many rivers were straightened, deepened and widened to facilitate catchment drainage and to prevent local flooding (McCarthy 1985; Brookes 1988). Instream gravel deposits and most instream woody debris were often dredged from such rivers, further reducing their physical heterogeneity (Swales 1989; Brookes 1988). The characteristic longitudinal and lateral sediment deposition pattern of actively meandering channels was then replaced by a more uniform and diffuse deposition of finer material in constrained channels. The physical complexity of natural marginal and riparian habitats was also usually greatly simplified. Water quality changed to reflect a greater input of nutrients and organic material from more-intensively managed catchments (Sweeting 1996; Riis & Sand-Jensen 2001).
The physical changes associated with river engineering are widely reported to reduce diversity, abundance and biomass among fish, invertebrate and macrophytes, particularly those species associated with coarse substrata, shallow water, high velocity and complex riparian or marginal habitats (Swales & O’Hara 1980; McCarthy 1985; Boon 1988; Swales 1989; Hey 1996). Such effects are increasingly seen as unacceptable in Western Europe and North America and attempts to rehabilitate heavily modified rivers are now being made, where they do not conflict with flood or drainage imperatives (Swales 1989; Boon 1988; Gardiner & Cole 1992; Brookes 1996).
River rehabilitation can be either passive, allowing natural hydraulic forces slowly to re-shape rivers, or active, applying specific measures to modify channel form and structure more rapidly (Gordon, McMahon & Finlayson 1992; Hey 1996). Most recent river rehabilitation projects in the UK have involved small-scale active methods, largely adopted from rehabilitation schemes developed for small salmonid streams in North America since the 1920s (Tarzwell 1932; Swales 1989; Brookes 1996). A variety of instream structures have been developed for these physically dynamic, high-gradient, gravel streams to mitigate the effects of channel engineering. Two of the commonest measures include artificial gravel riffles, intended to mimic natural dynamic riffles, and large boulders and lateral flow deflectors placed on the stream bed to increase channel sinuosity and flow heterogeneity (Tarzwell 1932; Wesche 1985; Brookes, Knight & Shields 1996). Such measures have now been widely adopted for lower-gradient, larger streams and rivers in the UK to enhance populations of lithophilous fish and other biota (Swales & O’Hara 1980; Swales 1982; Swales 1983; McCarthy 1985; Spillett, Armstrong & Magrath 1985; Swales 1989; Hey 1990; Brookes 1992; Brookes et al. 1996).
So far, few lowland river restoration projects have incorporated quantitative analysis of the impacts on fish and invertebrates, allowing little assessment of their true ecological impact or value (Friberg et al. 1994; Brookes 1996; Friberg et al. 1998). Where there have been ecological impact assessments, they have usually been conducted within single river systems and at small spatial scales, offering little insight into their general applicability. In view of the growing interest and expenditure on river rehabilitation, there is therefore an urgent need for a general, quantitative assessment of such practices, both in terms of species richness and ecosystem processes (Muotka & Laasonen 2002).
Here, we sought to quantify the biological impact of two common types of instream rehabilitation measure, artificial riffles and flow deflectors, installed over the last 10–15 years in heavily engineered, low-gradient rivers in central and eastern England. We adopted a paired sampling design, replicated across different rivers, comparing a rehabilitated stretch with a nearby, unrehabilitated reference stretch. This appears to be the first truly replicated assessment of such rehabilitation measures for any British river and we asked whether there were any general changes associated with artificial riffles and flow deflectors across a sample of schemes. We deal only with data on macroinvertebrates, having previously described impacts on fish (Pretty et al. 2003).
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The restoration of engineered channels in Europe has a much shorter history than in North America, but there has been a similar focus on instream habitat improvement (Brookes 1992; Iversen et al. 1993). Artificial rehabilitation structures have now been installed in many lowland rivers in the UK with the objective of enhancing (mainly non-salmonid) fish and macroinvertebrate communities (Driver 1997). Although stream restoration measures have long been the subject of evaluation and appraisal in North America (De Jalón 1995; Gore, Crawford & Addison 1998; White 2002), their success in Europe cannot readily be assessed because biological evaluation is only rarely undertaken (Iversen et al. 1993; Brookes et al. 1996; Harper, Ebrahimnezhad & Cot 1998). As a replicated spatial study on the biological impact of rehabilitation structures on macroinvertebrates in a number of UK lowland rivers, our contribution therefore aims to fill an important gap.
Headwater streams typically have greater abundance and diversity of benthic macroinvertebrates in natural riffles than in pools and other habitats (Rabeni & Minshall 1977; Logan & Brooker 1983; Allan 1995). This is generally attributed to the effect of current velocity and substratum particle size (Cummins & Lauff 1969; Hynes 1970; Allan 1975; Williams & Mundie 1978; Hildrew, Townsend & Henderson 1980; Erman & Erman 1984). The higher velocities and larger, more stable substratum particles of riffles offer more profitable foraging sites for algal grazers and filter feeders (Williams & Moore 1986; Allan 1995), while larger interstitial pore sizes can increase retention of particulate organic food and act as refugia from adverse flow conditions and/or predators (Gee 1982; Culp, Walde & Davies 1983). Much river rehabilitation in the UK has thus aimed to recreate distinct habitats, such as riffles and pools, in the belief that this will restore the physical heterogeneity of the original system and increase biodiversity within the channel. Such ‘functional’ habitats have been proposed to represent the building blocks of river rehabilitation, and, as such, should become the prime focus of river managers (Harper, Smith & Barham 1992; Harper & Everard 1998; Newson et al. 1998). The riffles and flow deflectors installed in the lowland rivers discussed here can be viewed as an attempt to recreate such functional habitats, and might have been expected to have strong ecological effects.
Our results suggest, however, that the impact on macroinvertebrate assemblages was only modest for artificial riffles and barely detectable for flow deflectors, particularly when viewed at a stretch-wide scale, when taking all habitats, including macrophytes, into account. Taxon richness and abundance was greatest in marginal and instream macrophytes, respectively, rather than in the rehabilitated benthos. Riffles and, particularly, flow deflectors added few invertebrate taxa not found in other habitats (3–6 for artificial riffles and 1–7 for flow deflectors) so that the main effect of artificial riffles was to increase the relative abundance of benthic rheophilic taxa already common elsewhere in the channel. Furthermore, several rehabilitation schemes failed to have any positive impact of stretch-wide diversity. Similarly, the effect of the same structures on fish populations in the same rivers was weak (Pretty et al. 2003). What might explain this limited ecological response to rehabilitation in these rivers?
Within a physical context, the structures may have been limited in several ways. First, they perhaps failed to provide habitat that was sufficiently different from non-rehabilitated habitats elsewhere in the river. Contrary to the expectation of the functional habitat approach – that distinct habitats represent ‘building blocks’ which can effectively be added into a system to create greater physical and biological diversity (Harper et al. 1992) – other factors might override the relationship between invertebrate taxa and particular habitats, particularly given the plasticity of response to habitat by many, although not all, lotic macroinvertebrates (Minshall & Minshall 1977; Jenkins, Wade & Pugh 1984; Barmuta 1989; Palmer & O’Keefe 1991; Bournaud, Tachet, Berly & Cellot 1998; Wright & Symes 1999; Harrison 2000). The apparent changes in physical habitat created by artificial riffles and, particularly, flow deflectors may lie within the normal physical tolerances of taxa living in the unrehabilitated channels.
Secondly, riffles may have only a weak functional role for benthic invertebrates in higher-order lowland rivers. In contrast to hydraulically active upland streams, flow refugia in lowland rivers are likely to be less important to benthic invertebrates, while aggregations of particulate organic matter (POM) will be less limiting (Allan 1995). In addition, both functions may be performed by the abundant instream and marginal macrophytes in lowland systems (Jacobsen & Sand-Jensen 1992).
Thirdly, habitat provided by rehabilitation structures may occur only rarely in lowland rivers. Riffles and pools reflect hydraulic forces acting on the stream bed, with their physical characteristics and longitudinal periodicity resulting from complex interactions between large-scale hydraulic forces and the dominant sediment size (Gordon et al. 1992). A strong active riffle-pool bedform is characteristic of high-gradient rivers of high stream power where there is an adequate coarse sediment supply. They would be less common in low-gradient, finer-sediment rivers of low stream power (Brookes & Sear 1996), possibly occurring at sporadic intervals in local high gradient stretches. The role of artificial coarse gravel riffles in replacing lost ‘natural’ habitat in these latter systems would thus be lessened.
Furthermore, the morphological features of unaltered rivers in their natural state are not static. For example, coarse sediment particles move downstream from one riffle to another during periods of high flow, while finer sediment is constantly deposited and scoured at all points along a river under more common hydraulic events (Gordon et al. 1992). It is likely that this sedimentary dynamism, mediated by hydraulic disturbance, is essential in the maintenance of natural benthic faunas (Hildrew & Giller 1994). Indeed, rheophilic invertebrates are more abundant on natural dynamic riffles that actively recruit new gravel (Rice, Greenwood & Joyce 2001). Artificial riffles in low-gradient systems, with very weak movement of larger particles, are unlikely to act like natural riffles, and will be further compromised by the deposition and infilling of fine sediments. This was certainly characteristic of many of the riffle schemes in our study.
Lastly, the rivers in which riffles (and riffle-dwelling taxa) occur more naturally are generally characterized by cool, clean, well-oxygenated water. Such conditions would be less usual in higher order, low-gradient systems, particularly those flowing through urban or intensively agricultural catchments. In several of the rivers we investigated, poor water quality resulting from high nutrient inputs is likely to have determined invertebrate community composition more than the lack of suitable physical habitat.
The physical context of small-scale rehabilitation schemes could mitigate against their ecological effectiveness. However, their ecological context, at a landscape scale, might also be important (Wiens 2002). The functional habitat approach may be too simplistic to predict a specific ecological response to small-scale rehabilitation measures (Harper et al. 1995): simply observing that particular habitats have partially distinct faunas does not imply that those faunas would be recreated by providing habitat in a small-scale restoration scheme. Local sets of species do not always interact with each other or their environment in the same way, frustrating site-specific predictions (Lawton 2000). Further, most macroinvertebrates have complex life cycles in which different life stages use different parts of the aquatic and riparian environment. This includes the use of marginal vegetation by newly hatched aquatic larvae, which then disperse into the mid-channel benthos (Cellot, Bournaud & Tachet 1984; Harrison 2000), while aerial adults sometimes use terrestrial bankside vegetation (Harrison et al. 2000; Harrison & Harris 2002). Simple structural features such as artificial riffles and flow deflectors are unlikely to satisfy all these requirements.
An ubiquitous pattern in large-scale ecology is the relationship between local and regional species richness, in which local species richness is constrained by the regional species pool (Lawton 2000). In the context of river schemes, the taxon richness of the rehabilitated reach was more a function of taxon richness in the wider, unimproved river, than of the physical structure itself. Such a result could argue for the primacy of larger scale regional processes in determining the local community in lotic ecosystems (Heino, Muotka & Paavola 2003). Thus, placing patches of gravel in a landscape where suitable species are scarce is unlikely to have a major, local effect. The metapopulation structure and dispersal of aquatic invertebrates is still poorly understood (Wilcock, Hildrew & Nicholls 2001; Berendonk & Bonsall 2002) but such should be considered in river rehabilitation.
The influences of large-scale factors, such as geology and land use on lotic invertebrate communities is well known (e.g., Hildrew & Giller 1994; Sponseller, Benfield & Valett 2001). This suggests the need for more extensive rehabilitation measures at the whole catchment scale, using the natural dynamics of the river to fashion instream and riparian habitats naturally, rather than artificially to create local patches of habitat (Sear 1994; Muhar, Schmutz & Jungwirth 1995; Brookes & Sear 1996; Poole, Frissell & Ralph 1997; Harper et al. 1999). Whole catchment restoration is often impractical in intensively managed lowlands, as the need for flood control and land drainage usually overrides any attempt to ‘restore’ whole river catchments to their natural state. Often, however, features at smaller scales in the catchment hierarchy, such as reach-scale riparian land use are important (Harrison 2000; Sponseller et al. 2001; Harrison & Harris 2002). Marginal and riparian habitats are rich in terrestrial and semi-aquatic species and have an important impact on the structure, functioning and diversity of river ecosystems (Hynes 1975; Sweeney 1993; Harrison 2000; Harrison & Harris 2002).
The future of lowland river rehabilitation will differ from past practices. Ideally, it should involve minimal engineering and should rely on natural fluvial dynamics as far as practicable. Rivers should be given as much lateral space as possible, reinstating floodplains, side-channels and marginal habitats where feasible. This study, in particular, has demonstrated the importance of marginal, emergent macrophytes for invertebrate diversity in managed rivers. Rehabilitation of riparian habitats, by the simple act of fencing them from cattle, can have significant benefits for invertebrate communities in taxon-rich lowland streams (Harrison & Harris 2002).
The study has also highlighted the need for much more clearly defined goals and expectations in river rehabilitation, both in the targets (such as the kinds of species, processes or habitat that is desired) and in the context within which projects are set. A macrophyte-rich, low gradient, downstream reach of a river, for example, would require a very different set of rehabilitation principles compared to a higher-order, high gradient, more erosive system. Even within ‘lowland’ rivers, the physical nature of streams may differ widely, largely as a function of catchment geology. Rivers of similar discharge and gradient flowing through limestone, sandy, clay and sandstone catchments may be expected to have markedly different physical, chemical and biological properties. Equally, rehabilitation in rivers flowing through intensively managed catchments may be constrained by excessive nutrient inputs, precluding the reintroduction of target species that need good water quality. Cost effective rehabilitation should thus be tailored both to the natural physical, chemical and biological conditions of the river and to the anthropogenic disturbance that determines the actual state of the river.