THE CONCEPTUAL BASIS OF ECOLOGICAL-STATUS ASSESSMENTS USING DIATOMS Martyn Kelly, Lydia King and Bernadette Ní Chatháin ABSTRACT
Martyn Kelly (corresponding author; email:
[email protected]), Bowburn Consultancy, 11 Monteigne Drive, Bowburn, Durham, DH6 5QB, UK; Lydia King, Limnologie-PhykologieDiatomologie, Amberg 1, 88326 AulendorfZollenreute, Germany; Bernadette Ní Chatháin, Shannon River Basin District Project, Mulkear House, Newtown Centre, Annacotty, Co. Limerick, Ireland. Cite as follows: Kelly, Martyn, King, Lydia and Ní Chatháin, Bernadette 2009 The conceptual basis of ecological-status assessments using diatoms. Biology and Environment: Proceedings of the Royal Irish Academy 109B, 175–89. DOI: 10.3318/ BIOE.2009.109.3.175.
Most member states of the EU have chosen to use existing diatom-based metrics for assessment of ecological status, as required by the Water Framework Directive (WFD), rather than develop new methods. In this paper we assess the suitability of such methods in light of the requirements of the normative defi nitions of ecological status. In particular, we focus on the rationale for the placement of ecological-status boundaries. As the WFD defi nes ecological status in terms of the ‘structure and functioning’ of aquatic ecosystems, we interpret changes in the diatom assemblage in light of changes in entire phytobenthos. Whilst we believe that analysis of diatom assemblages is a sensible approach to developing a first generation of WFD-compatible tools, bearing in mind tight deadlines and limited budgets, the potential of non-diatoms should not be ignored when developing future methods.
INTRODUCTION The Water Framework Directive (WFD; 2000/60/EC; European Parliament and Council 2000) has stimulated a great deal of work in recent years as member states develop the methods required to fulfil its obligations. These include requirements to monitor a wider range of ecological properties than was formerly the case and to report the state of the environment in terms of ‘ecological status’, defined as ‘an expression of the quality of the structure and functioning of aquatic ecosystems’. Annexes of the WFD provide more detail about those parts of the biota, termed ‘biological quality elements’, that should be assessed and describe those characteristics of each biological quality element that need to be considered when assessing ecological status. One of the biological quality elements for rivers and lakes is ‘macrophytes and phytobenthos’. Most member states have chosen to evaluate macrophytes and phytobenthos separately and to use diatoms as proxies for the phytobenthos, mainly because a suite of well-developed methods already exists, underpinned by European standards (European Committee for Standardization 2003; 2004), and because relationships between diatoms and the chemical environment are well established (Kelly and Whitton 1995; Bennion et al. 1996; Battarbee et al. 1999). In this paper we argue that the WFD introduces new concepts to the legislation and management of freshwaters
and that, consequently, there is no a priori reason why methods developed to meet the needs of earlier legislation will meet all the needs of the WFD. Earlier methods relied on strong relationships between the diatom assemblage and environmental variables, and the resulting metrics provided a direct link with the pressures that needed to be managed. Most member states appear to have transferred this logic to their WFD monitoring systems; however, because diatoms are just one part of the phytobenthos, diatom-based metrics are subject to an additional layer of uncertainty: there is both an error term associated with the relationship between the diatom-based metric and the pressure and an error term, as yet unquantified, with the ‘true’ ecological status based on a holistic understanding of the phytobenthos. This may appear to be a point of semantics; however, there are practical consequences. Member states are expected to manage freshwater ecosystems in order to attain good ecological status, defi ned as ‘no more than a slight change in the composition and abundance of macrophytic and phytobenthic taxa compared to typespecific communities’ (European Parliament and Council 2000; Annex V). Such a state is desirable in order to ensure the sustainable use of water resources in the future. Therefore, there is a need for an ecologically based rationale for defi ning ecological-status classes for all biological quality elements, for knowing how changes in pressures
DOI: 10.3318/BIOE.2009.109.3.175 Biology and Environment: Proceedings of the Royal Irish Academy, Vol. 109B, No. 3, 175–189 (2009).
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Biology and Environment affect the structure and functioning of biological quality elements and for understanding how different biological quality elements interact within the freshwater environment. In this paper we attempt to reconcile the WFD’s legal framework with ecological theory and to develop a conceptual framework for good ecological status that is compatible with the existing use of diatom-based metrics. This framework will then be used to point out some limitations of current methods and to map out future research directions. Our core argument is that, although diatoms represent a good means for assessing changes in ecological status, setting status class boundaries requires a holistic understanding of the structure and functioning of the entire phytobenthos. HISTORICAL OVERVIEW The development of diatom-based pollution indices prior to 1980 is summarised in Prygiel et al. (1999). Whilst many indices and approaches had been developed, few were adopted by regulatory agencies as there was no legislative requirement for this type of monitoring and little perceived benefit in using diatoms, rather than invertebrates, for general water-quality assessment. The most significant development during the 1980s was that of the Indice de Polluosensibilité (IPS; Coste in CEMAGREF 1982), a metric, based on the weighted averaging equation of Zelinka and Marvan (1961), that claimed to provide integrated assessments of a range of water-quality variables, including organic pollution, eutrophication, salinity and toxic materials. This justification was based on strong relationships between the metric and the principal water-quality variables (Coste et al. 1991), but, whilst this index has been widely used, the lack of ability to discriminate between different types of pollution became a problem as legislation became more demanding in the 1990s. The Urban Waste Water Treatment Directive (UWWTD; 91/271/EEC; Council of the European Union 1991) required EU member states to identify and control discharges responsible for eutrophication. This legislation is significant, as regulators recognised a need to use primary producers, rather than invertebrates, for monitoring on a large scale. The development of the IPS highlighted the potential for using diatoms as monitors. Yet the IPS was itself unsuitable as it failed to discriminate between the general ‘saprobic’ effects of organic pollution and the effects of inorganic nutrients in isolation. Some
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work on this topic had already been done in Germany (Steinberg and Schiefele 1988), but the UWWTD provided an incentive to develop indices (Fig. 1)—the Trophic Diatom Index (TDI) in the UK (Kelly and Whitton 1995); an index, also called the Trophic Diatom Index, in Germany (Coring et al. 1999); and the Trophienindex (TI) in Austria (Rott et al. 1999). A separate metric for assessing organic-pollution (saprobic) effects, called the Saprobienindex (SI), was developed alongside the TI (Rott et al. 1997). At about the same time, the Indice Biologique Diatomique (IBD) was developed in France; this focused on organic and saline pollution effects but had a more sophisticated computational basis than the IPS, along with simplified taxonomy (Lenoir and Coste 1996). Whilst metrics such as the IPS tended to use diatoms to provide a complementary viewpoint of the effect of chemical pollutants, developments during the 1990s saw a gradual recognition of the ability of diatoms to tell their own story, with the chemistry increasingly becoming the supporting actor. For example, Kelly (2002) described changes in diatoms along the River Wear, north-east England, noting that the relationship between diatom assemblage composition and dissolved nutrients becomes increasingly uncoupled as the river receives more nutrient inputs. The implication was that nutrient reduction would not, of itself, achieve an ecological change unless concentrations could be reduced to a point where they start to affect community composition and biomass (see also Dodds et al. 1997; Bowes et al. 2007). This trend towards an ecological basis for assessment continued with the WFD (European Parliament and Council 2000), which required all waterbodies to be raised to a condition of good ecological status, defi ned primarily by the biological communities that they contained. Interestingly, despite the radical nature of this legislation, the response of most member states has been to adopt existing methods rather than develop new ones. This generally conservative approach reflects the high costs associated with developing new approaches, but it may also indicate a failure to grasp the true nature of the WFD. In particular, STAR, an EU-funded project that evaluated approaches to monitoring for the WFD, looked only at existing metrics and used correlations with a simplistic ‘nutrient/ organic gradient’ (Hering et al. 2006), thus missing many of the ecological and chemical nuances within this gradient. For an intercalibration exercise conducted between twelve member states in the central latitudes of Europe, five states
The conceptual basis of ecological-status assessments using diatoms UWWTD
WFD IBD
IPS
TDI (UK) TDI (D) S&S
TI & SI
1980
1990
WFD methods 2000
2010
Fig. 1—Schematic overview of the development of diatom-based metrics for evaluating water quality in Europe since 1980. IBD = Indice Biologique Diatomique; IPS = Indice de Polluosensibilité; S&S = Steinberg and Schiefele (1988); SI = Saprobienindex, TI = Trophienindex (Rott et al. 1999); TDI (D) = Trophic Diatom Index, German version; TDI (UK) = Trophic Diatom Index, UK version (Kelly and Whitton 1995); UWWTD = Urban Waste Water Treatment Directive; WFD = Water Framework Directive.
used the IPS or IBD, two used the TDI and two used a combination of the TI and SI. North-west Spain used a multimetric that combined several existing indices with two regionally specific metrics, whilst Flanders and the Netherlands developed new metrics from scratch (Kelly et al. 2008c); although the Netherlands subsequently changed to the IPS (van Dam et al. 2008). THE NORMATIVE DEFINITIONS— WHAT THEY SAY AND WHAT THEY MEAN The WFD provides a general def inition of ecological status as an ‘expression of the structure and functioning of aquatic ecosystems’ (European Parliament and Council 2000; Article 2), and provides further detail in Annex V. For rivers and lakes, member states are expected to monitor four biological quality elements: phytoplankton, macrophytes and phytobenthos, benthic invertebrates and fish. For macrophytes and phytobenthos, the WFD defines ecological status in terms of the composition and abundance of the f lora and of the presence of undesirable disturbances and bacterial tufts. For composition and abundance, the key distinction is between a slight change (good status) and a moderate change (moderate status), leaving member states with the task of defining ‘slight’ and ‘moderate’ in a meaningful manner. ‘Undesirable disturbances’ are not
defined further in the WFD but are now understood to mean ‘a direct or indirect anthropogenic impact on an aquatic ecosystem that appreciably degrades the health or threatens the sustainable human use of that ecosystem’ (ECOSTAT 2005). These should be absent at ‘good status’, but there is no mention of their occurrence at moderate, poor or bad status. Finally, ‘bacterial tufts’—not defined but interpreted as ‘sewage fungus’ and related heterotrophic growths—are absent at good status but ‘may interfere with or displace the phytobenthic community’ at moderate status. Most member states regard ‘macrophytes and phytobenthos’ as primarily an indicator of the extent of cultural eutrophication, and it is perhaps useful to compare the normative defi nitions with current understandings of eutrophication (Table 1). There is a fairly straightforward link between the ‘undesirable disturbances’ of the WFD and the excessive growths of fi lamentous algae, such as Cladophora, and associated sideeffects, such as oxygen depletion. It is, however, harder to reconcile such undesirable disturbances with the ‘moderate changes in composition and abundance’ expected at moderate status. For this reason, we suggest that at moderate status there may be a possibility of undesirable disturbances when there are prolonged periods of low flow, for example, but that such conditions would not typically be encountered in a river whose flora corresponds to a ‘moderate change’ in either composition or abundance from the undisturbed
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Biology and Environment Table 1—A comparison between a definition of eutrophication (Environment Agency 2000) and WFD definitions of good and moderate status for macrophytes and phytobenthos in rivers. Eutrophication
Normative definition for macrophytes and phytobenthos Good status
The enrichment of water by nutrients … … stimulating an array of symptomatic changes … … including increased production of algae and/or higher plants … … which can adversely affect the diversity of the biological system, the quality of the water and the uses to which the water may be put.
No more than slight changes in composition …
Composition differs moderately from the type-specific community … … and abundance of macrophytic and … Moderate changes in the phytobenthic taxa … average macrophytic and the average phytobenthic abundance are evident … … Such changes do not indicate any accelerated growth of phytobenthos or higher forms of plant life resulting in undesirable disturbances to the balance of organisms … … The phytobenthic community … The phytobenthic community is not adversely affected by bacterial tufts may be interfered with and, in some areas, displaced by bacterial and coats. tufts and coats.
condition. Within moderate status we would expect a gradually increasing likelihood of ‘undesirable disturbance’ with distance from the good/moderate boundary. By poor or bad status, we suggest that undesirable disturbances will be typical of the river, not just during periods of extended low flow. If avoidance of undesirable disturbances is a valid management target, there is a problem if the main method adopted throughout Europe for assessing phytobenthos uses a group that is rarely, in rivers at least, directly responsible for undesirable disturbances. In a number of cases it should be noted that diatom-based methods are performed in parallel to assessments of the macrophyte community and that fi lamentous algae such as Cladophora are included in many of these. The requirement to assess composition presents fewer challenges as the weighted averaging methods described in the previous section provide a sensitive measure of the taxonomic changes along the gradient. There are a few data which show that changes in composition of diatoms reflect changes in other algae (Kelly 2006; Kelly et al. 2008b), and a few member states choose to assess non-diatoms as well as diatoms (Rott et al. 1999; Schaumburg et al. 2004). What composition metrics do not provide, however, is any insight into deciding where the boundaries
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Moderate status
should be placed, and countries have adopted a variety of ways of defi ning a slight and a moderate change, from simply dividing the metric scale into equal divisions to methods that are more strongly rooted in ecological theory (Kelly et al. 2008a). This topic will be considered in more detail in the next section. Measuring abundance is also problematic. Measurements of phytobenthic biomass are time-consuming and subject to a variety of sources of spatial and temporal uncertainty (Pan et al. 1999). Overall, disturbance plays an important part in determining the biomass measured on any given occasion (Cardinale et al. 2006); as a result, relationships between biomass and nutrient concentrations are rarely strong. On the other hand, measurements of abundance provide another link between the WFD and defi nitions of cultural eutrophication. Also, in waterbodies where eutrophication is severe, the excessive biomass of fi lamentous algae is obvious to the naked eye. In other words, there may be more practical sense in conducting simple visual assessments of fi lamentous algae rather than complicated and time-consuming measurements. The rationale for assessment of bacterial tufts is similar to that for assessment of undesirable disturbances: they are not measured directly by diatom metrics, and whilst it should be pos-
The conceptual basis of ecological-status assessments using diatoms sible to establish values of the chosen metric that correspond to a high probability of such growths, it makes more sense for the presence of these to be noted at the same time as samples are collected.
WHAT DO DIATOM METRICS TELL US? The previous sections have demonstrated two apparently conf licting points: first, that most member states are using diatoms for assessing ecological status; second, that diatom metrics partially fulfil only one of the requirements of the normative definitions—that for assessing taxonomic composition. This is not, however, the whole story, as most member states use diatoms in tandem with macrophyte assessment systems, which may provide stronger evidence of excessive biomass of filamentous algae and of undesirable disturbances. Similarly, some member states assess bacterial tufts separately. There is, in other words, a good case for diatom metrics as part of an overall multimetric assessment system for ‘macrophytes and phytobenthos’, but a considerably weaker case for any member state that chooses to use diatoms as their sole tool for assessing this biological quality element. Setting meaningful boundaries between status classes is another challenge that member states have had to overcome, with the good/ moderate boundary proving to be problematic because of the need to distinguish between a slight and a moderate change—a distinction that is as much etymological as it is ecological. Both of these problems could be addressed if there was a robust conceptual framework through which the large quantity of diatom data that has been collected in recent years could be interpreted. Some progress on this has been made, both in terms of the functional ecology of stream periphyton assemblages (e.g. Biggs et al. 1998) and in terms of the processes governing stream eutrophication (Hilton et al. 2006). The remainder of this review is going to place this theoretical work into the context of the WFD and of the diatom monitoring that is taking place around Europe to fulfi l national obligations. A CONCEPTUAL FRAMEWORK FOR ECOLOGICAL STATUS AS APPLIED TO DIATOMS The current practice for sampling and analysing benthic diatom samples from rivers is now standardised around Europe (Kelly et al. 1998;
European Committee for Standardization 2003; 2004), and involves removing the biofilm from a number of replicates of a single habitat (typically cobbles), followed by digestion in one or more oxidising agents, preparation of a permanent slide and analysis in order to determine the composition and relative abundance of each taxon. Recent advances in diatom taxonomy (e.g. Krammer 1997a; 1997b; Lange-Bertalot 2001; Mann et al. 2008) have resulted in a steep increase in the number of species recorded from Europe, and there is a prevailing, though largely unspoken, dogma that identifying all of these will lead to more-sensitive indices than were possible in the past. A typical diatom sample will contain 30–40 taxa, or more. However, when the taxa are arranged in rank order, they typically display a log-normal distribution, with a few taxa present in considerable numbers and a long ‘tail’ of less common species (Fig. 2) (Patrick 1973; van Dam 1982). Contrary to the prevailing dogma that focuses on fi ne-level taxonomy of the entire diatom assemblage, we suggest that an understanding of ecological status should focus on the broader ‘association’ of phytobenthos and that this will in turn provide a more direct link to the WFD’s defi nition of ecological status as a measure of the structure and functioning of aquatic ecosystems. In order to illustrate this point, we ask readers to visualise the transition between a forest and grassland. Phytosociologists would recognise each of these as distinct associations, recognise transition zones between the two with distinct associations of their own and describe each in terms of the dominant and subdominant taxa that are present. Ecologists could also point to differences in the way that the two ecosystems cycled C and inorganic nutrients, show how the plant taxa in each association are interrelated with each other and demonstrate how this affected the types of consumers that were able to thrive in each. They could, in short, provide an assessment of the way in which the two associations were structured and of how they both functioned. We can push this analogy a little further because the climax vegetation across most of central and northern Europe is forest, and most grassland in this area is the result of human modification of the landscape. If we considered ‘forest’ to be the reference state (to use WFD terminology), then ‘deforestation’ might be an appropriate measure of the amount of human disturbance, analogous to metrics such as the IPS. However, we would also need to defi ne a point on this gradient of deforestation when the ‘forest’ is no longer a ‘forest’. Of course, 179
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Relative abundance
1
0.1
0.01
0.001
0.0001 1
6
11
16
21
26
31
36
41
46
51
56
61
66
Rank order of taxa Fig. 2—Relative abundance of diatom taxa found in the River Ribble, downstream of Clitheroe sewage-treatment works, 17 September 2004. Only four out of the 69 taxa recorded at the site constitute more than 10% of the total count each, whilst 58 taxa comprise < 1% of the total each. The River Ribble at Clitheroe is not at good status, but this rank-abundance pattern is typical of benthic diatom samples from both rivers and lakes.
any felling to create clearings for agriculture or settlement will result in local changes in ‘structure’ and ‘functioning’. A forest with clearings is still, in common parlance, a ‘forest’. But at what point will the term ‘forest’ become redundant for a particular area? This is a relevant question to ask with respect to the aquatic environment, too, as it introduces the idea that monitoring ecological status is as much about recognising a ‘state’ as it is about measuring biotic parameters. Because phytobenthos specialists, in particular, have to use destructive sampling methods, leading to sub-sub-sub-samples of the original biota arranged randomly on a slide and then magnified hundreds of times, it is easy to resort to a ‘nameand-count’ approach, with the resulting taxa list summarised by a simple metric, rather than to focus on properties and characteristics that may give us more profound insights into ecological status. Some preliminary thoughts on this subject are outlined in Yallop and Kelly (2006). In the years that followed this publication, we have gathered more data and developed our ideas (Kelly et al. 2008a). The study by Kelly et al. (2008a) included a screening process to identify ‘reference sites’: sites free from significant, short-term anthropogenic influences. Truly ‘pristine’ sites are absent from much of north-west Europe, and these sites probably equate to ‘least-disturbed condition’, following the terminology of Stoddard et al. (2006). The diatom assemblage at these sites showed typical log-normal distributions, with a few taxa abundant regardless of the type of stream, and others of varying abundance depending particularly upon the alkalinity of the stream. Observations of live samples from these 180
streams over a number of years (unpublished studies) have shown that non-diatom algae also form a significant part of the microphytobenthos at many of these sites—principally Cyanobacteria (both with and without heterocysts), Chlorophyta and Rhodophyta. Typically, these non-diatoms were less common than the diatoms (up to about one-third in rivers and just over a half of the total in lakes), but there were samples in which they were more abundant than the diatoms. Though most studies have focused on spatial differences in diatom assemblages, and have related these to changes in the chemical environment, there are a number of recent papers showing that there is also a considerable amount of within-site variability (Kelly 2002; Springe et al. 2006; Kelly et al. 2009). Part of this may reflect natural micro-successional processes (Biggs et al. 1998), and it is important to recognise that such systems are inherently variable and, indeed, that such variability is a fundamental characteristic, imparting resilience to the ecosystem (Moss 2008). Part of this can be interpreted in terms of intra- and interspecific competition, with a shift in species composition during the course of the succession towards those species that are better able to thrive in situations where there is intense competition for scarce resources. Using detailed studies in one near-pristine lake, Wastwater in the UK (King 2000), we were able to reorder the species lists into visualisations of the biota at three different stages (Fig. 3). Achnanthidium spp. were common throughout the succession, but other diatom species, such as Tabellaria, Gomphonema acuminatum and Cymbella affinis, were more common at later stages. In this particular example, non-diatoms were relatively scarce, but this was
The conceptual basis of ecological-status assessments using diatoms
Fig. 3—Diagrammatic representation of a diatom-dominated epilithic biof ilm at three stages of development. Based on data from benthic assemblages in Wastwater from King (2000). Scale bar: 10µm. (a) After one week, early colonists such as Achnanthidium minutissimum (i), Gomphonema parvulum (ii), coccoid cyanobacteria (iii) and narrow filaments of Oscillatoria (iv) occupy space on the upper surface of littoral rocks. (b) After three weeks, these algae cover most of the available space on the rock, and density-dependent factors start to have a significant inf luence on the community ((v) coccoid Chlorophyta). (c) At six weeks, the structure of the community has changed, with long-stalked taxa such as Gomphonema acuminatum (vi) and Cymbella (vii) rising above the layer of Achnanthidium, whilst f ilamentous taxa such as Tabellaria flocculosa (viii) grow entangled with the attached algae.
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Biology and Environment not always the case. An important point, which has a bearing on our understanding of good ecological status, is that at the reference state most of the diatoms were either attached directly to the substratum or ‘tangled’ in the upper layers of the mature biofi lm. Motile diatoms, such as Navicula and Nitzschia, were relatively scarce. Whilst there were a few representatives of both of these genera (e.g. Navicula angusta, Nitzschia disputata), which are found at high status, these were rarely found in any great numbers. Our experience is that such assemblages are the most commonly encountered across a wide range of streams and rivers that qualify as reference sites. In soft-water sites (e.g. < 2mg l−1 Ca) the proportion of Eunotia spp. relative to Achnanthidium increased, and Achnanthes oblongella and Tabellaria flocculosa also tended to be more common, whilst at sites with hard water there was an increase in the relative abundance of Cymbella spp. (e.g. C. affinis, C. delicatula, C. microcephala) and Amphora pediculus. We are, however, cautious about ascribing too much importance to the abundance of A. pediculus at reference conditions, due to the problems of fi nding sufficient lowland, highalkalinity sites at or near to reference conditions. Overall, despite the many problems associated with species-level taxonomy of Achnanthidium (Potapova and Hamilton 2007) and Fragilaria capucina (Hürlimann and Schanz 1991), we believe that it is possible to defi ne the reference states for phytobenthos as assemblages where Achnanthidium, in particular, tends to be abundant. This, we suggest, forms the matrix within which other taxa grow. During the early part of successions, this taxon is often overwhelmingly dominant, but as resources become limiting in the biofi lm, longer-stalked taxa, such as Gomphonema acuminatum, and loosely attached or tangled taxa, such as Fragilaria capucina and Tabellaria flocculosa, are favoured; although Achnanthidium is also found at these later stages. The role of spates to ‘reset’ the biofi lm to a pioneer state is supported by the literature, although some caveats need to be applied. Indeed, floods that did not overturn the substratum have actually led to increased algal biomass (Stevenson 1990), perhaps because invertebrate grazers were removed. Blenkinsopp and Lock (1994) noted that Cocconeis placentula remained attached to substrata even after spates, though both they and Stevenson (1990) noted an increased proportion of broken diatom frustules after high-flow events. On the other hand, Kelly et al. (1995) and Ní Chatháin and Harrington (2007) both noted more motile diatom taxa on boulders
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compared to cobbles and related this to the frequency with which each was rolled by floods. These observations correspond with those of Passy (2007), who noted fewer erect, fi lamentous, branched, chain-forming, tube-forming, stalked, colonial, centric or motile diatom taxa at high current velocities. Van Dam and Mertens (1995), by contrast, associated high proportions of fi lamentous diatoms, especially Aulacoseira crenulata and A. alpigena, with streams in the Netherlands with low current velocity and constant, deep groundwater discharge. Finally, Biggs (2000) recorded a positive relationship between benthic chlorophyll concentration and the number of days since the most recent flood disturbance. Factors such as these create some of the spatial and temporal variation that is inherent in diatom metrics and that needs to be appreciated when using diatoms for monitoring (Kelly 2002; Kelly et al. 2009). In addition to this characteristic (albeit dynamic) structure, there are a number of physiological characteristics associated with phytobenthos assemblages at high status. Many of the taxa associated with such sites are, to use the terminology of Grime (1979), ‘stress-tolerant’ and include taxa with adaptations to environments with low concentrations of available nutrients. For example, nitrogen-fi xing cyanobacteria are frequent, as are taxa that are known to be able to use extracellular phosphatases to acquire P from organic compounds, such as Chaetophorales (Gibson and Whitton 1987), Batrachospermum (Whitton 1988) and Didymosphenia (Ellwood and Whitton 2007). The current orthodoxy of the River Continuum Concept (Vannote et al. 1980; Allan 1996) considers headwaters, in particular, to be net heterotrophic, fuelled by abundant inputs of plant litter and other debris. However, much of the C in such sources needs to be processed by invertebrates before it becomes available to micro-organisms. In temperate streams the balance will shift when forest cover is removed (Mulholland et al. 2001), allowing greater primary productivity, and this in turn will stimulate greater heterotrophic activity (Dodds 2006). The net productivity of epilithon alone appears to not have been so widely studied. Lock et al. (1984) show both chlorophyll and adenosine triphosphate concentrations on submerged rocks to be highly variable, so generalisations may be difficult. Nonetheless, sites subject to organic and/or nutrient pressures have a completely different assemblage of diatoms. Whilst some individuals of the taxa associated with reference conditions may be present, there is typically a distinct
The conceptual basis of ecological-status assessments using diatoms increase in motile diatoms, particularly Navicula sensu lato and Nitzschia sensu lato, along with taxa such as Rhoicosphenia abbreviata and Cocconeis pediculus that are commonly epiphytic on fi lamentous algae. The precise mechanisms that govern these changes are not yet known, and what follows is a tentative hypothesis, based on our own experiences of UK and Irish rivers. As resource supply increases, ‘competitive’ (sensu Grime 1979) algae such as Cladophora glomerata are favoured over the stress-tolerant algae that are common at reference conditions. Cladophora attaches to the substratum via basal holdfasts, which form a dense, intertwined network over the rock surface. These in turn trap sediment particles, whilst an increase in dissolved organic matter will lead to an increase in heterotrophic organisms such as bacteria and hyphomycetes. A diatom such as Achnanthidium that attaches directly to the substratum will now be at a competitive disadvantage over those, such
as Navicula and Nitzschia, in particular, that are able to move through the dense networks of organic and inorganic material that form the biofi lm. Similarly, the fi lamentous algae forming the ‘canopy’ will themselves provide a substratum for epiphytic diatoms, along with other algae (e.g. Chamaesiphon incrustans). This type of assemblage is visualised in Fig. 4, although the precise combination of diatoms that are found will reflect adaptations to different levels of pressure. For example, several representatives of Navicula sensu stricto are common when nutrient levels are elevated but organic pollution is quite low, whilst genera such as Fistulifera and Mayamaea, both formerly Navicula sensu lato, are associated with high concentrations of organic pollution. Species of Nitzschia are, by contrast, spread along the gradient, though they are rarely abundant at high or good status. N. fonticola is associated with moderate levels of pressure, whilst N. palea and N. communis
Fig. 4—Diagrammatic representation of biofilm architecture at moderate status. Increased nutrient concentrations lead to growth of competitive filamentous algae such as Cladophora glomerata (Cg). Greater availability of organic matter from anthropogenic sources and from recycling within the biofilm favours heterotrophic organisms such as Sphaerotilus natans (Sn). Greater autotrophic and heterotrophic production leads to a thicker biofilm that in turn traps mineral particles. These conditions no longer favour algae that live attached directly to the substrata; instead, motile taxa such as Navicula tripunctata (Nt) and Nitzschia fonticola (Nf ) thrive along with epiphytic taxa such as Chamaesiphon incrustans (Ci), Cocconeis placentula (Cp) and Rhoicosphenia abbreviata (Ra), as well as taxa such as Diatoma vulgare (Dv) that are loosely attached or tangled amongst other organisms. Ciliate protozoa such as Vorticella sp (Vs) may also be present. gv = girdle view; vv = valve view. Scale bar: 10µm.
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Biology and Environment tend to be associated with higher levels of anthropogenic pressure. As for high and good status, this characteristic structure is linked to the way that the biofi lm functions. For example, competitive algae such as Cladophora are efficient at using available P sources, have high nutrient-uptake rates, are fi rmly attached to the substratum via heterotrichous holdfasts and are resistant to grazing (Dodds and Gudder 1992; Biggs et al. 1998). These taxa form a canopy under which a mixture of auto- and heterotrophic organisms live, and as the level of nutrient/organic pressure increases, the balance will shift towards heterotrophy rather than autotrophy. Some of the algae found within these biofi lms are facultative heterotrophs, such as the Euglenophyta (van den Hoek et al. 1995), but there are also records of both C and N heterotrophy within the diatoms (Hellebust and Lewin 1977; Tuchmann 1996). We have sketched this conceptual framework in broad terms, and we know that there will be exceptions where local environmental features create different ‘reference conditions’ to those described here and where pressures lead to different trajectories being followed than a straightforward nutrient/organic gradient. Our main point is that diatomists have focused too much on the metrics that measure generalised pressure gradients and have not spent enough time relating these to particular ecological states. This framework provides an ecologically based rationale for the placement of status class boundaries (Kelly et al. 2008a) and in turn for the development of regulatory standards that will protect this state. Similar principles can be applied to other pressure gradients (e.g. acidification) to enable a holistic approach to environmental assessment.
A PARADIGM SHIFT IN EUROPEAN PHYTOBENTHOS MONITORING? Most EU member states have chosen to base assessments of phytobenthos on diatom metrics, mostly developed before the WFD was envisaged (see Kelly et al. 2008c). All of these metrics rely on species-level identification of diatoms, and most ignore non-diatom algae. Yet the framework outlined above relies more on defining a broad assemblage of phytobenthos and ignores finerscale variations, and it is possible that this can be done without recourse to fine-level taxonomy. In order to test this hypothesis, we calculated two indices from the UK data set used in Kelly et al. (2008a): the IPS, which requires specieslevel identification; and the Indice Diatomique Générique (IDG; Rumeau and Coste 1988), which uses genus-level identification. There was a good correlation between these two indices (r2 = 0.62) and also good correlations between both of these indices and log soluble reactive phosphorus (r = −0.64 for IPS and 0.56 for IDG). For log nitrate-N, there was actually a stronger correlation with the IDG (r = −0.54) than with the IPS (r = −0.43)—a feature also observed by Kelly et al. (1995). A cluster of samples fell below the main trend, and closer inspection showed these were dominated by Amphora pediculus, which the IPS treated as more sensitive than the way the genus was treated in the IDG. By adjusting the IDG so that Amphora pediculus was given the same score as in the IPS, the r2 value increased to 0.71 (Fig. 5). In other words, about two-thirds of the variation in the species-level index could be explained by the genus. Water-quality managers, however, are interested primarily in whether or not a site achieves
IDG (adjusted)
20
16
12
8
4 4
8
12
16
20
IPS Fig. 5—Relationship between the IPS, a species-level index, and the IDG, a genus-level index, for UK river sites. The vertical line represents the good/moderate boundary adopted by Wallonia (southern Belgium), and the horizontal line is the same boundary, computed from the regression equation, for the genus-level index.
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The conceptual basis of ecological-status assessments using diatoms good status. In order to look at the efficacy of species- versus genus-level metrics for this purpose, we superimposed the good/moderate boundary used by Wallonia (southern Belgium) over Fig. 5. This was a convenient model to use as Belgium is just 50km from southern England and has a straightforward classification system, without a typology. The equivalent boundary for the IDG was computed from the regression equation. Samples that fall into the bottom left and top right quadrants of this graph give the same answer (i.e. both ‘moderate or less’ or both ‘good or better’), whilst those in the top left and lower right quadrants give different answers to this key question. Interestingly, 91% of all samples achieved the same classification, regardless of whether species- or genus-level identification was used. What is the reason for this? One possible argument is that many of the traits that influence structure and function are associated with a genus, family or order, rather than with a species. These traits include habit—whether a diatom is motile, prostrate on a substratum, erect or stalked (Round et al. 1990; Cox 1996)—but also ecophysiological characteristics. For example, N fi xation is restricted to one family of freshwater diatoms—Rhopalodiaceae (Round et al. 1990)— and a few orders of cyanobacteria (van den Hoek et al. 1995), whilst extracellular phosphatases are similarly restricted to a few orders across different phyla (Whitton 1988). The ability to use genera in this way also reflects the outcome of taxonomic research that has led to the revision of generic concepts (e.g. Williams and Round 1986; 1987; Krammer 1997a; 1997b), as genera in use now should be more natural groupings than they were twenty years ago. If a number of species share genes that control their morphology, it is reasonable to assume that they also share genes controlling aspects of their physiology and, consequently, that some consistency in ecological responses is to be expected. Palynologists routinely make profound insights about vegetation history and environmental change (e.g. Huntley and Webb 1988) based on pollen grains that are typically resolvable only to family or genus. This is not to argue that species-level identification has no value. Indeed, there is evidence for a range of environmental tolerances within some genera (e.g. Potapova and Hamilton 2007; Poulícˇ ková et al. 2008). Yet, conversely, the small size of many diatoms, coupled with the lack of clarity of the descriptions in the literature, means that such taxa are also often a source of variability in inter-laboratory comparisons (Kelly et al.
2002; Prygiel et al. 2002; Kahlert et al. 2009). It seems clear that some Achnanthidium spp. (e.g. A. eutrophilum, A. saprophilum) are genuinely tolerant to elevated nutrients and/or organic pollution (Potapova and Hamilton 2007). Given the framework outlined above, the proliferation of ‘tolerant’ Achnanthidium taxa must indicate an unusual set of environmental conditions, yet such nuances are lost when the data are reduced to a few simple metrics. And there is a fi nal intriguing paradox: does a shift from A. minutissimum to A. eutrophilum actually indicate a change in the ‘structure’ and ‘function’ of the biofi lm? Unfortunately, for this diatom, as for so many, the subtleties of its autecology have yet to be investigated. In other words, we will not be able to make full use of developments in fi ne-scale diatom taxonomy until we make similar improvements in our understanding of the ecology of these species. To be blunt, diatomists are good at describing diatom species ecology in terms of a few variables that are easy to measure but are, in the process, missing many nuances of the interactions between physical and chemical environments, of the importance of the speciation of nutrients and of the effects of short-term variability in the chemical environment. Yet, at the same time, we are not making full use of phytobenthic taxa that are often visible alongside the diatom assemblages that we sample, and whose ecology is known in some detail (Livingstone and Whitton 1984; Gibson and Whitton 1987; Lindstrøm et al. 2004). Incorporating such information into assessment systems whilst retaining their costeffectiveness is one of the intriguing challenges for the next decade. CONCLUSIONS Given short timescales and limited budgets, it is no great surprise that many EU states have resorted to tried-and-tested methods rather than developed new methods from scratch. Such methods were designed to measure pressures rather than ‘ecological status’ as such. Part of the objective of this paper has been to try to provide some ecological underpinning that justifies the use of such metrics for ecological-status assessment. Certainly, where the nutrient/organic gradient is concerned, we believe that such simple metrics probably do provide an efficient means of summarising changes along the gradient, and could form the foundation for a suite of metrics that encompass other pressures such as acidification and toxic pollution.
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Biology and Environment However, the problem with the use of simple metrics for measuring ecological status is that they give no indication of where to place ecological-status class boundaries. The reason for this, we suggest, is that the studies have focused too much on fi ne details of diatom taxonomy. The outcome should perhaps be described as ‘taxonomic status’ rather than ‘ecological status’ because it fails to incorporate higher-level thinking about how the diatoms contribute to the overall structure and function of the phytobenthos. By focusing on genus-level identification in the previous section, we are not suggesting that this, alone, will be sufficient. We wanted to point out that the considerable effort that diatomists routinely put into species- and varietal-level identification may lead to more data but that it does not necessarily lead to more information from the point of view of the end-user. Understanding ‘ecological status’ means focusing on the ‘big picture’, and would the effort that goes into keying out rare and uncommon diatoms not be better spent checking the identities of the more conspicuous macroalgae that grow alongside the diatoms? Why rely on diatoms to confi rm that a river is in ‘bad status’ when rocks are covered with sewage fungus? In short, how best should the fi nite amount of time available to a single sample be divided up in order to extract the maximum amount of information from it? The fi rst generation of tools developed for the WFD probably qualify—just—as ‘fit-forpurpose’. There has not been the paradigm shift in the way that applied aquatic biologists think about monitoring that some (e.g. Moss 2008) believe is needed, but the problems facing the regulatory agencies and tool developers were substantial, and it is wrong to suggest that progress has not been made. The challenge is whether we can now start to build more ecological thinking into a second generation of tools? The situation is, we believe, similar to the state of navigation prior to the scientific revolutions ushered in by Copernicus and Galileo. In those days, despite the erroneous belief that the sun revolved around the earth, there was a cumbersome yet effective method of stellar navigation, devised by Ptolomy and developed by others over many centuries (Koestler 1959). The key point is that sailors generally got to where they wanted to go. After Copernicus and Galileo, navigation was placed on a fi rmer conceptual basis. Navigation subsequently involved fewer complicated calculations, and, because the underlying theory was sound, the discipline was able to develop.
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And so it is for diatoms and the WFD. The use of pressure metrics for assessing ecological status probably gives the right answer most of the time. But we see little scope for real progress until ecological-status assessment is placed on a fi rmer theoretical foundation.
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