ICES CM 2003/Session J-02
CLASSIFICATION TOOLS FOR MARINE ECOLOGICAL QUALITY ASSESSMENT: THE USEFULNESS OF MACROBENTHIC COMMUNITIES IN AN AREA AFFECTED BY A SUBMARINE OUTFALL Ángel Borja, J. Franco and I. Muxika AZTI Foundation Department of Oceanography and Marine Environment Herrera Kaia, Portualdea, z/g 20110 – Pasaia (Spain) Tel: +34-943-004800; fax: +34-943-004801 e-mail:
[email protected]
Abstract In December 2000, the Water Framework Directive (WFD) entered into force. The WFD establishes that the ecological status of water masses must be assessed. In order to undertake this requirements, several biological elements must be considered; amongst others, benthic invertebrates. To comply with the requirements of the WFD, some classification tools for biological elements have been developed within Europe. Benthic macroinvertebrates are the most traditionally used biological indicators of ecosystem health in the marine environment. AZTI has developed a tool (AMBI: www.azti.es), which provides assistance in determining the impacts and quality status of soft-bottom marine benthic communities; it is being utilised broadly along European coasts. In this communication, different metrics for benthic macroinvertebrates in coastal areas are compared, including the AMBI, richness, diversity, biomass and abundance. They are also combined in a biological quality index. Such comparison is undertaken by considering the results obtained from a monitoring programme in a coastal area affected by a submarine outfall, on the Basque coast (N. Spain). The investigation includes several surveys undertaken before and after the construction of the outfall. These results will be discussed within the framework of the WFD requirements. Keywords: biotic coefficient, impact evaluation, soft-bottom benthos, Water Framework Directive, submarine outfall
Introduction The European Water Framework Directive (WFD 2000/60/EC) develops the concept of Ecological Quality Status (EQS) for the assessment of the quality of water masses and, by 2006, European Member States will be required to establish ecological quality objectives. The assessment of the status will be based upon the composition and abundance of different biological elements of the ecosystem (e.g. phytoplankton, benthos, fish), as well as the
BORJA, FRANCO AND MUXIKA
physico-chemical and hydromorphological indicators in relation to reference sites. The WFD sets the objective to prevent deterioration in the status of all Community waters (i.e. both surface- and ground-waters, including coastal waters, throughout the EU) and to ensure the achievement of their good status by 2015. Similarly, other international initiatives and agreements are following the same objectives. The Oslo & Paris Convention (1992) has promoted and is in the process of adopting Ecological Quality Objectives (EQO), for the implementation of its strategies to combat eutrophication and for the promotion of an ecosystem approach to environmental assessment. The European Environment Agency has been advocating the development of indicators of change and, on a larger scale, the Earth Summits at Rio de Janeiro (1992) and Johannesburg (2002) require quantitative indicators as a tool in the modelling, monitoring and management of aquatic systems. For the marine environment, it is well accepted that some biological components are more valuable than others as integrative environmental indicators and the WFD places a significant emphasis on sedentary benthic communities inhabiting the beds of estuaries and coastal areas. Ecologically-based classification schemes for the WFD will be based upon five classes (High, Good, Moderate, Poor and Bad), with the main aim being to achieve at least ‘good’ ecological status for all water bodies. According to the WFD, the surface water bodies are classified in rivers, lakes, transitional waters or coastal waters (apart from artificial and heavily modified surface water bodies). Likewise, the application of the WFD to transitional and coastal waters requires a significant amount of effort, to ensure effective implementation, as outlined by the CIS Working Group 2.4 (Coast) created by the Common Implementation Strategy. Further to their central role in marine ecosystem functioning, the benthic invertebrates are a well-established target in evaluations of environmental quality status. Various studies have demonstrated that the macrobenthos responds relatively rapidly to anthropogenic and natural stress (Pearson and Rosenberg, 1978; Dauer, 1993) and macrobenthic animals: (i) are relatively sedentary i.e. cannot avoid deteriorating water/sediment quality conditions; (ii) have relatively long life-spans (thus, indicate and integrate water/sediment quality conditions, over time); (iii) consist of different species that exhibit different tolerances to stress; and (iv) have an important role in cycling nutrients and materials, between the underlying sediments and the overlying water column (Hily, 1984; Dauer, 1993). Several authors have reviewed the use of biotic indices (Washington, 1984; Codling and Ashley, 1992). Many authors (e.g. Washington, 1984) accept that a biotic index is unlikely to be universally applicable, as organisms are not equally sensitive to all types of anthropogenic disturbance and are likely to respond differently to different types of perturbation. As such, they may provide a way to establish a multimetric bioassessment method that can be modified for different geographical regions. Several indices have been proposed for use in estuarine and coastal waters; some of these attempt to include the five-step environmental model of the WFD (Rumohr et al., 1996). Some authors have made attempts to use these tools as a proxy of the impact at sea (Hily, 1984; Rygg, 1985; Majeed, 1987; Dauer, 1993; Engle et al., 1994; Grall and Glémarec, 1997; Weisberg et al., 1997; Roberts et al., 1998; Borja et al., 2000, 2003; Eaton, 2001; Simboura and Zenetos, 2002). Moreover, these approaches are based upon different premises and do not address all of them directly in the establishment of the EQS, for the whole of Europe, sensu the WFD.
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In this case, the use of the diversity and abundance of macroinvertebrate taxa, together with the presence of disturbance-sensitive taxa and taxa indicative of pollution (biotic indices), are proposed to be measured as useful metrics in determining the ecological status. However, at present, the abovementioned tools do not fulfil all the requirements proposed in the WFD. A recent guidance document, written by the group implementing the WFD, declares that “methods combining composition, abundance and sensitivity may be the most promising” (Vincent et al., 2002). Borja et al. (2000) proposed a Marine Biotic Index (AMBI) to establish the ecological quality of soft-bottom benthos within European estuarine and coastal environments. Such an index, based upon the sensitivity/tolerance of benthic fauna to stress gradients, classifies the species into five ecological groups. The distribution of these ecological groups provides a biotic index of 5 levels of pollution classification, as in the WFD (Table 1). The index has been validated and applied to different impact sources and geographical areas (Borja et al., 2000, 2003), demonstrating its usefulness. Table 1. Summary of the AMBI values and their equivalences (after Borja et al., 2000). The last column shows the proposed equivalent Ecological Status (WFD), as used in this study.
Dominating Site Pollution Ecological Benthic Community Health Ecological Group Classification Status I Normal 0.0 < BC ≤ 0.2 Unpolluted High Status Impoverished 0.2 < BC ≤ 1.2 III Unbalanced Slightly Polluted Good Status 1.2 < BC ≤ 3.3 Moderate Status Transitional to pollution 3.3 < BC ≤ 4.3 Meanly Polluted IV-V Polluted 4.3 < BC ≤ 5.0 Poor Status Transitional to heavy pollution 5.0 < BC ≤ 5.5 Heavily Polluted V Heavily polluted 5.5 < BC ≤ 6.0 Bad Status Azoic (7.0) Azoic Azoic Extremely Polluted AMBI value
In this study, the possible incorporation of new metrics to the AMBI will be explored, following the requirements of the WFD. The data provided by the monitoring of macrobenthic communities in an area affected by a submarine outfall in San Sebastian (Basque Country, North of Spain) (Figure 1), will be used as a case study.
Methodology The biotic index An extense presentation of the index development and its application has been described in Borja et al. (2000). The updated species list, with more than 2,000 taxa from all European seas, including their assignment to the ecological groups, together with the AMBI 2.0 program (AZTI’ Marine Biotic Index) to calculate and represent the index, are available (free of charge) in AZTI’s web page (www.azti.es). The list and the program are being continuously updated.
The case study In the spring of 2001, as a transitory solution until complete cleaning of the water within the context of the sewerage scheme, the initial discharges (old outfall) from the San Sebastian and Pasaia area (north of Spain), were diverted to a submarine outfall. This outfall
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is located 1.2 km from the coast in an approx. 47 m water depth (Figure 1a).
(a)
(b)
Figure 1. (a) Location of the case study area, in the north of Spain, showing the position of the old outfall and the new submarine outfall, and (b) the sampling stations over the area.
The benthic communities were studied 5 months before the diversion, and 4 and 16 months after the diversion. Benthos was sampled with a box-corer grab, at 9 sampling stations (Figure 1b); 3 replicates were taken at each sampling site. All of the samples were sorted out, identified, and counted. Benthic communities at two separate stations (sampled before the diversion, in areas of 50 and 160 m water depth, some 5 miles apart, not shown in Figure 1) were used, as a proxy to reference conditions over the region (data from Martínez and Adarraga, 2001). Here, data have been used on species richness (number of species), abundance (as number of individuals, per square metre), Shannon’s diversity (based upon abundance), and AMBI (calculated as mentioned above).
Incorporating new metrics The metrics used, in order to accomplish the requirements of the WFD, require that the final value must range between 0 (bad ecological quality) and 1 (high ecological quality), with 5 levels of quality. In this study, it is proposed to combine three different metrics: diversity, richness and the biotic coefficient (AMBI), following the criteria listed in Table 2. These metrics provide a value which is a rate between the value obtained and that of the reference site (in this particular case, taking as reference those values equivalent to an Ecological Quality Ratio (EQR) of 1). For each of the metric values there is an Equivalent Assigned Value (EAV, in Table 2). The EQR is calculated as the mean of the EAVs for each station. Then, each EQR has an associated ecological status.
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CLASSIFICATION TOOLS FOR MARINE ECOLOGICAL QUALITY ASSESSMENT Table 2. Calculating the Ecological Quality Ratio (EQR), with diversity, richness and AMBI, together with the equivalence in terms of the ecological status (WFD). EAV: Equivalent Assigned Value.
Diversity
Richness
AMBI
EAV
EQR
Ecological Status
0-1.2
0-15
5.5-7
0
0-0.25
Bad
1.2-2.4
15-30
4.3-5.5
0.25
0.25-0.5
Poor
2.4-3.6
30-45
3.3-4.3
0.5
0.5-0.7
Moderate
3.6-4.8
45-60
1.2-3.3
0.75
0.7-0.9
Good
>4.8
>60
0-1.2
1
0.9-1
High
Results and Discussion Figure 2 shows the main results obtained at each of the locations and for each of the sampling periods. Before the discharge diversion, the highest richness (> 70 species) is observed at the reference stations (R-50 and R-160) (Figure 2a); the lowest value (11 species) is reached near the old outfall (COAST) and in the immediate surroundings; these are more affected by pollution. After diversion, there is a progressive improvement in the richness values near the old outfall (approaching 40 species after 16 months). However, the new submarine outfall, together with those to the south (stations OUTFALL, S1 and S2), experiences some deterioration in richness, after the diversion. No clear trends are observed in the stations northwards to the submarine outfall. Before diversion, the reference stations, with those situated far to the north from the old outfall, present the lowest abundances (Figure 2b). The most affected station (COAST) show high abundance values (> 5,000 ind.m-2), due to the presence of very abundant small opportunistic species (such as Capitella capitata or Malacoceros fuliginosus, which represent more than 95% of the total abundance). After the diversion, this station experienced an immediate increase in abundance, followed by a decrease; this was due to the new conditions i.e. the absence of organic inputs to the system. On the other hand, the abundance dramatically increased in the newly affected stations, with values > 4,000 ind.m-2. Before the diversion, the highest diversities (between 3.5 and 5.7 bit.ind-1) were found at the reference stations and the non-affected stations (Figure 2c). After the discharge diversion, the old outfall station (COAST) improves significantly its diversity, from 4 values. The area near the new outfall reaches values around 1 bit.ind-1. On another occasion, the most affected area, in terms of diversity, is that situated to the south of the new impact source point (OUTFALL); changes to the north are indistinguishable. The same pattern can be seen in relation to the Biotic Coefficient (Figure 2d). The reference and the non-affected stations present low biotic coefficients (unpolluted or slightly polluted (following the terminology of Borja et al., 2000, see also Table 1)) before the diversion. For contrast, the stations more affected by the discharges present high biotic coefficients (heavily polluted). Following the diversion, there is an improvement in terms of this index at the COAST station, but the area around the new discharge is of poor quality. In general, well-marked gradients, both spatial (Figure 3) and temporal (Figures 2d and 3), can be detected by means of the Biotic Coefficient.
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Richness (n sp.)
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90 80 70 60 50 40 30 20 10 0
(a)
Abundance (n.m-2)
10000
(b)
8000 6000 4000 2000
Biotic Coefficient
Shannon Diversity
0 6 5
(c)
4 3 2 1 0 6 5
(d)
4 3 2 1 0 COAST
S2
SW
SE
S1 OUTFALLNW
NE
N
R-50
R-160
Figure 2. Evolution of several structural parameters at the studied locations: (a) taxonomic richness; (b) abundance; (c) Shannon diversity; and (d) biotic coefficient (AMBI). Key: black columns show data 5 months before the discharge diversion from the old outfall (COAST), to the new submarine outfall (OUTFALL); grey columns show data 4 months after diversion; and white columns 16 months after diversion.
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1.8
1.5
1.1
2.2
1.8
2.3
2.1
1.5
5.8
1.5 2.3
1.3
2.5
1.4
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4
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AFTER-2
Figure 3. The Biotic Coefficient values (AMBI), both in spatial and temporal perspective.
In studying the different relationships between the biological indicator metrics used here (Figure 4), the Biotic Coefficient is related negatively with diversity and richness and is related positively with abundance. Further, abundance is related negatively with richness and diversity and, finally, richness and diversity are positively related. This pattern indicates, at least in this particular case study, that there exist close relationships between the metrics proposed by the WFD for the determination of the ecological status. Taking into account the small discriminatory effect of the abundance, in terms of detecting the impact of pollution on communities (see data in Figure 2b), those structural parameters helping in the determination of such an impact have been used in this study. These are, as described in the methodology, diversity, richness and biotic index. In this way, when calculating the EQS for the different sampling stations, before and after diversion and based upon the methodology shown in Table 2, the results are very comprehensible under the terms of the WFD (Figure 4). Hence, the area near the old outfall (COAST), improves in its quality, from a ‘bad’ ecological status to a ‘moderate’ status, after 14 months of the discharge diversion. In the case of the new submarine outfall (OUTFALL), together with the surrounding affected area (S1, S2), the opposite situation occurs: from a ‘moderate’ or ‘good’ status, the area changes to a ‘poor’ or ‘bad’ status. Sampling areas used as reference, or areas to the north of the new outfall (out of the main tidal current direction), remain with ‘high’ or ‘good’ ecological status. Finally, other areas are less affected by the discharges (SW, SE), but change their status over time, with a small worsening in their quality. An interesting point to note is that the deterioration of the benthic communities in the area affected by the new discharges is much faster (less than six months) than the recovery of the communities in the areas positively affected by the waste elimination (more than one year). The results obtained here suggest that the combined use of biotic indices (such as AMBI), in combination with other structural parameters of the community (such as richness and diversity), could be useful in determining the ecological status of the European transitional and coastal waters. Furthermore, this approach can accomplish all the requirements of the WFD (using diversity and richness, together with the presence of disturbance-sensitive taxa and taxa indicative of pollution or biotic indices).
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Figure 4. Relationships between the different biological indicator metrics. Statistically significant correlations are indicated by * (p
