© 2004 International Council for the Exploration of the Sea
The effects of marine sand and gravel extraction on the macrobenthos at a commercial dredging site (results 6 years post-dredging)
The Centre for Environment, Fisheries & Aquaculture Science, Burnham Laboratory Remembrance Avenue, Burnham-on-Crouch, Essex CM0 8HA, England, UK
*Correspondence to S. E. Boyd: tel: +44 1621 787245; fax: +44 1621 784989. e-mail: S.E.Boyd{at}CEFAS.co.uk.
Benthic recolonization was investigated at a site historically used for the extraction of marine sand and gravel. The main objective was to assess the effects of different levels of dredging intensity on the recolonization of benthic fauna and sediments. Preliminary observations from this study indicated that the fauna within an area of seabed exposed to high dredging intensities remained in a perturbed state some 4 years after the cessation of dredging. Thereafter, annual monitoring surveys of the benthos and sediments at the "treatment" and "reference" sites have followed the recolonization process. Results from univariate and multivariate data analyses show that distinct differences in the nature of assemblages at sites exposed to high and lower levels of dredging intensity persist at least 6 years after the cessation of dredging. This paper presents the physical and biological findings 6 years after dredging, together with a generic framework for evaluating post-cessation recolonization studies.
Keywords: aggregate extraction, dredging, impacts, North Sea, recolonization
Received 30 November 2003; accepted 20 November 2004.
 |
Introduction |
|---|
 Top Introduction MethodsResultsDiscussionConclusionsReferences |
|---|
Much of the seabed surface around the England and Wales coastline consists of coarse material, i.e. various proportions of sand and gravel (CIRIA, 1996). Where these resources are present in sufficient quantity, are of the right composition, and are accessible to commercial dredgers, they may be exploited as a source of aggregate for the construction industry, to supplement land-based sources, or as a source of material for beach nourishment (Singleton, 2001).
As the extraction of marine sand and gravel has its primary impact at the seabed, assessment of the effects of this activity has conventionally targeted bottom substrata and the associated benthic fauna (Millner et al., 1977; Desprez, 2000; Van Dalfsen et al., 2000). Historically, the scientific study of coarser substrata has presented a significant challenge, largely on account of the difficulties of obtaining reliable quantitative samples (Eleftheriou and Holme, 1984). As a consequence, information on the nature and distribution of benthic assemblages, and on their wider role in the marine ecosystem, is considerably more limited than in areas of soft sediments.
Of those studies which have considered the effects of marine aggregate extraction, most have concentrated on establishing the rates and processes of macrobenthic recolonization upon cessation of dredging (Cressard, 1975; Kenny et al., 1998; Desprez, 2000; Sardá et al., 2000; Van Dalfsen et al., 2000; Van Dalfsen and Essink, 2001). These studies indicate, typically, that dredging causes an initial reduction in the abundance, species diversity, and biomass of the benthic community (for review see Newell et al., 1998) and that substantial progress towards full restoration of the fauna and sediments can be expected within a period of approximately 2–4 years following cessation (Kenny et al., 1998; Sardá et al., 2000; Van Dalfsen et al., 2000; Van Dalfsen and Essink, 2001). For example, Van Dalfsen et al. (2000) suggested that recolonization of a dredged area by polychaetes occurred within 5–10 months after the cessation of dredging in a site located within the North Sea, with restoration of biomass to pre-dredge levels anticipated within 2–4 years. Such studies have been mainly concerned with the effects of dredging operations conducted over a relatively short time scale, e.g. up to periods of 1 year (Kenny et al., 1998; Sardá et al., 2000; Van Dalfsen et al., 2000; Van Dalfsen and Essink, 2001). Under such circumstances, any more subtle effects, e.g. on seasonal recruitment success to the locality, arising from prolonged dredging over several years would clearly be expected to be minimal. Few studies have addressed the consequences of long-term dredging operations (Desprez, 2000). Thus, there is limited information which is directly applicable to the impacts of commercial dredging operations in UK waters where the life-time of a typical production licence is at least 15 years. The aim of this study was to assess the status of the seabed substrata and associated benthic assemblages at a former extraction site which was intensively dredged over a 25-year period. Preliminary observations on the status of this extraction site, 4 years after the cessation of dredging, were reported in Boyd et al. (2003). In this paper we examine the findings from surveys carried out between 2000, and 2002 i.e. 4, 5, and 6 years on, and investigate whether different historical levels of dredging intensity affect the subsequent rate and nature of benthic recolonization at an aggregate extraction site following cessation of dredging.
|
Methods |
|---|
|
TopIntroductionMethodsResultsDiscussionConclusionsReferences |
|---|
Study site
A full description of the study site (designated "Area 222") together with an account of the dredging history is reported in Boyd et al. (2003). It is located approximately 20 miles east of Felixstowe off the southeast coast of England (Figure 1) in water depths of between 27-m and 35-m Lowest Astronomical Tide (LAT). The tidal ellipse in the region is rectilinear and is aligned in a NNE-SSW direction which is thought to be modified by an adjacent deeper channel that encroaches into the northern edge of the extraction site. Maximum spring tidal current velocities reach 1.5 m s–1 and there is evidence for a NNE nearbed residual tidal direction (Boyd et al., 2003). This site, with dimensions of approximately 900 m by 300 m, was first licensed for sand and gravel extraction in 1971, with a peak in extraction activity recorded as 872 000 t in 1974. Extraction continued at levels >100 000 t per annum until 1995, before the site was relinquished by the industry in 1996. At this site, the sand:gravel ratios of dredged cargoes were adjusted by screening, with excess sand being discharged overboard at the site of dredging.
|
Sampling design
Since 1993, every vessel dredging on a Crown Estate licence in the UK has been fitted with an Electronic Monitoring System (EMS) which automatically records the date, time, and position of all dredging activity, every 30 s, to disk. EMS information was interrogated in order to locate areas of the seabed within the Area 222 extraction licence which had been subjected to different levels of dredging intensity. Replicate samples of the macrofauna and sediments were collected from areas representing two different levels as follows (i) >10 h of dredging within a 100-m by 100-m block during 1995 and (ii) <1 h of dredging within a 100-m by 100-m block during 1995. EMS data from 1995 were chosen since this year was the last year that the licensed extraction site was dredged heavily. The location and intensity of dredging was comparable between 1993 and 1995. In addition, a reference site (Reference site 1) was sampled in 2000–2002 and this was augmented by sampling at a second reference site (Reference site 2) in 2001 and 2002. Boyd et al. (2003) present an account of the design in terms of the likely dredging impact. Stations were randomly distributed within each area ("stratified random sampling") and allocated in proportion to the size of the sampling box (Green, 1979). Reference sites were selected as being representative of the wider environment surrounding the extraction site and outside the influence of any potential effects on the benthos from dredging (see Figure 1). Selection of appropriate reference sites was aided by the use of sidescan sonar and video images of the seabed (see Boyd et al., 2003 for methodology) and following criteria given in CSTT guidelines (1997) and Boyd (2002). There was also no evidence of the effects of other forms of seabed disturbance at the reference locations, i.e. effects of trawling activity. Data arising from this design provide a comparative evaluation of "treatment" and "reference" groups (e.g. Skalski and McKenzie, 1983). Note that the "reference" areas are not necessarily representative of baseline conditions, as there is insufficient historical information on which to determine what actually constitutes the likely pre-dredging status of Area 222.
Area 222 was not dredged in the 4 years prior to sampling. Sampling was conducted in July 2000, 2001, and 2002, that is 4, 5, and 6 years after the cessation of dredging. Sampling details for locations sampled as part of this study are presented in Table 1.
|
Sample collection
Samples for analysis of the macrobenthic fauna and sediment particle size were collected with a 0.1-m2 Hamon grab from RV "Cirolana". This device was employed because it has been shown to be particularly effective on coarse substrata (Kenny and Rees, 1994, 1996; Seiderer and Newell, 1999).
Following estimation of sample volume, a 500-ml subsample was removed for laboratory sediment particle size analysis. The whole sample was then washed over 5-mm and 1-mm square mesh sieves to remove the fine sediment. The two resultant fractions (1–5-mm and >5-mm) were back-washed into separate containers and fixed in 4–6% buffered formaldehyde solution (diluted in seawater) with the addition of "Rose Bengal", a vital stain.
Macrofauna samples were processed according to the guidelines given in Boyd (2002). The >5-mm sample fraction was first washed with freshwater over a 1-mm mesh sieve in a fume cupboard, to remove excess formaldehyde solution, then back-washed onto a plastic sorting tray. Specimens were removed and identified, where possible, to species level. The 1-5-mm fraction was first washed over a 1-mm sieve then back-washed into a 10-litre bucket. The bucket was filled with freshwater and the sample was then gently stirred in order to separate the animals from the sediment. Once the animals were in suspension, the sample was decanted over a 1-mm mesh sieve. This process was repeated until no more material was recovered. Specimens from this fraction were placed into labelled petri dishes for identification and enumeration. The sediment was then placed on plastic trays and examined under an illuminated magnifier for any remaining animals such as bivalves not recovered in the decanting process, which were then added to the petri dishes.
Sediment particle size analysis
The sediment subsamples from each grab were analysed for their particle size distributions. Samples were first wet-sieved on a 500-µm stainless steel test sieve using a sieve shaker. The <500-µm sediment fraction passing through the sieve, was allowed to settle from suspension in a container for 48 h. The supernatant was then removed using a vacuum pump and the remaining <500-µm sediment fraction was washed into a petri dish, frozen for 12 h, and freeze-dried. The total weight of the freeze-dried fraction was recorded. A subsample of the <500-µm fraction was then analysed using a laser sizer and a percentage weight for each size class was calculated. The >500-µm fraction was washed from the test sieve into a foil tray and oven dried at 
90°C for 24 h. It was then dry sieved on a range of stainless steel test sieves, placed at 0.5-phi intervals, down to 1 phi (500 µm). The sediment on each sieve was weighed to 0.01 g and the values recorded. The results from these analyses were combined to give a full particle size distribution for each sample.
Acoustic surveys
Sidescan sonar surveys were undertaken using the DatasonicsTM SIS 1500 digital chirps system using the Triton IsisTM data acquisition software. The DelphmapTM software package was used to post-process the data, and provided georeferenced mosaic images of the sonar data. Such surveys were undertaken in order to provide an indication of the spatial distribution of sediments in the wider area encompassing the dredged sites and to provide information on the distribution and stability of bedforms. Such information contributes to an evaluation of inter alia the physical "recovery" of sites, e.g. the persistence of dredged tracks or pits.
Data analysis
Sediment variables
Particle size distribution data have been presented using cumulative frequency distribution curves. Changes in the shape of the curve for any given sample when compared to another, reflect the variations in the particle size distribution of those samples.
A correlation-based principal components analysis (PCA) was applied to ordinate results from the sediment analyses (Clarke and Warwick, 1994).
Analysis of similarities (ANOSIM, Clarke, 1993) was performed on sediment particle size data to test the significance of differences in particle size composition between treatments.
Macrofaunal assemblage structure
The total numbers of individual organisms and numbers of species were calculated for each sample group. This allows a visual interpretation of any trends (e.g. increasing or decreasing abundance at different sampling locations and over time) and their statistical significance, whereas this judgement is more difficult for results obtained by multivariate data analyses. The significance of differences between treatments was tested using one-way ANOVA.
A non-parametric multi-dimensional scaling (MDS) ordination using the Bray–Curtis similarity measure (Bray and Curtis, 1957) was applied to species abundance data.
Warwick and Clarke (1993) noted that in a variety of environmental impact studies, the variability among samples collected from impacted areas was much greater than that collected from reference sites. They suggested that this variability was in itself an identifiable symptom of perturbed situations. To test whether this pattern was evident with the data from dredged sites examined in this study, the comparative Index of Multivariate Dispersion (IMD) was calculated. IMD has a maximum value of +1 when all similarities among impacted samples are lower than any similarities among reference samples. The converse case gives a minimum for IMD of –1, and values near zero imply no difference between groups.
The comparative Index of Multivariate Dispersion is restricted to the comparison of only two groups, e.g. reference vs. high dredging intensity samples and therefore is usually complemented by calculation of the relative Index of Multivariate Dispersion (r.IMD; Somerfield and Clarke, 1997). This index has a value of 1 if the relative dispersion of samples corresponds to the "average dispersion". Values greater than 1 are obtained if replicate samples are more variable than average. In contrast, a value lower than 1 is achieved if replicate samples are less variable than average.
Analysis of similarities (ANOSIM, Clarke, 1993) was performed to test the significance of differences in macrofauna assemblage composition between samples. The nature of the groupings identified in the MDS ordinations was explored further by applying the similarity percentages program (SIMPER) to determine the contribution of individual species to the average dissimilarity between samples.
All multivariate analyses were performed using the software package PRIMER v. 5, developed at the Plymouth Marine Laboratory (Clarke and Gorley, 2001).
|
Results |
|---|
|
TopIntroductionMethodsResultsDiscussionConclusionsReferences |
|---|
Sediment characteristics
Sediment particle size characteristics are presented in Table 2 and the cumulative particle size distribution curves for each of the survey years are presented in Figure 2. Grain size descriptions relate to the Udden–Wentworth scale (Wentworth, 1922). Particle size data revealed that there was a high degree of variability between replicate samples, particularly in the gravel and sand components of the distributions.
|
|
In both 2000 and 2001, the replicate sediments sampled at the site of high dredging intensity show a large degree of variability in the gravel and coarse sand fractions. In 2002, an apparent reduction of the gravel component at this location produced less variability between replicate samples. Reference site 2 was sampled in 2001 and 2002 only, and sediments from this location show more variability between replicates than those found at either Reference site 1 or the site of low dredging intensity. An ordination by PCA of sediment particle size data is illustrated in Figure 3.
|
In terms of particle size distribution, sediments collected from the area of low dredging intensity and the reference locations were more similar to each other than to sediments from the area of high dredging intensity. This was due to the higher percentage of coarse sand from samples collected from the area of high dredging compared with the samples from the area of low intensity and reference locations. This is reflected in the PCA ordination by the separation of the high intensity samples from the low intensity and reference samples (Figure 4). The particle size distributions of samples from within the area of low dredging intensity were also more consistent over time, as depicted by the tighter clustering of samples in the PCA ordination. In contrast, there was a much higher degree of particle size variability between replicate samples collected from the area of high dredging intensity and the reference locations, as represented by the much wider spread of samples from these locations in the PCA ordination. The separation of sediments collected from the area of low dredging intensity and the reference locations is largely on account of the higher silt/clay content of some of the reference samples (Figure 4).
|
In all, 70% of the total variation is explained by the first two principal components, indicating that the two-dimensional ordination gives an appropriate representation of the similarity between the collected sediments. Table 3 shows the analysis of similarities results (ANOSIM, Clarke, 1993) for particle size data between samples collected from the different treatments over the 3-year period of study. Sediments at all locations were significantly different (p < 0.05) from each other in terms of particle size characteristics, apart from the two reference sites in both 2001 and 2002, the site of higher dredging intensity and Reference site 2 in 2001, and the site of high dredging intensity in 2000 and Reference site 2 in 2001. Sediment characteristics differed over time at each location, although taken together these differences were not, generally, found to be statistically significant at p < 0.05.
|
Macrofaunal assemblage structure
In all, 289 taxa were identified from the 75 Hamon grabs collected from the different treatments. Excepting values of abundance in 2002, population densities and numbers of species of macrofaunal invertebrates were significantly lower (p < 0.05) in the site exposed to the highest level of dredging intensity compared with the site of lower dredging intensity and reference conditions (Figure 5). In 2002, higher densities of Pomatoceros lamarcki (Quatrefages, 1866) recorded in one of the samples from Reference site 2 increased the variability around the mean. In this year, therefore, there was no recorded difference between the dredged sites and reference conditions in terms of population densities, although there was still a difference between sites of higher and lower dredging intensity.
|
In general, differences between the site of higher dredging intensity and other sample locations were due to the absence or reduced abundance (p < 0.05) of a range of macrofaunal species characterizing nearby sediments, including the tube worm P. lamarki, the pea crab Pisidia longicornis (Linnaeus, 1767), the polychaete Lumbrineris gracilis (Ehlers, 1868), and the brittle star Amphipholis squamata (Chiaje, 1829). Densities of these species were variable between different locations and between different years (Figure 6). Densities of Amphipholis squamata increased between 2000 and 2002, while densities of Lanice conchilega (Pallas, 1766), a sand-dwelling polychaete were significantly higher (p < 0.05) at all sampled sites in 2001.
|
The MDS ordination for macrofaunal assemblages collected at sites of high and lower dredging intensity and at the two reference sites is presented in Figure 7. While the reference samples and low dredging intensity samples show tight clustering of replicates, indicating a high stability of the spatial pattern, the high intensity replicate samples are much more diffusely distributed. This separation of the individual replicates from the area of high dredging intensity indicates that they are biologically dissimilar.
|
The comparative Index of Multivariate Dispersion (IMD) has been calculated in order to contrast the multivariate variability among samples taken from the dredged sites with samples from the reference locations (Table 4). Comparisons between the site of high dredging intensity and the reference sites and the sites of high and lower dredging intensity give the most extreme values of IMD, i.e. close to +1. In comparison, there is little difference between the low dredging intensity and reference samples in terms of variability in multivariate structure. Thus, a pattern of high variability in multivariate structure with increased disturbance is clearly evident, in all years, at Area 222. Calculation of the relative Index of Multivariate Dispersion (r.IMD) confirms the conclusions from above, e.g. that there is an increased variability in community composition at the site of high dredging intensity in comparison with the other sampled locations (Table 5).
|
|
Macrofaunal assemblages generally discriminated well between different sampling locations in each year. ANOSIM results in Table 6 also confirm the patterns observed in the MDS ordinations. Macrofauna assemblages at all locations were significantly different (p < 0.05) from each other in terms of species composition, apart from the two reference sites in both 2001 and 2002, and the site of lower dredging intensity and Reference site 1 in 2001.
|
Further exploration of the community groupings subject to differing levels of dredging impact was undertaken using the similarity percentages program (SIMPER). Results revealed that the average similarity between replicate samples collected for each of the groups was low, particularly for samples collected from the area of high dredging intensity (see Table 7). This reflects the relatively few shared species found between replicate samples obtained from the area of high dredging intensity.
|
The output from SIMPER also indicates which taxa contribute the most towards similarity between replicate samples from within each of the groups. Characterizing species from each of the groups were similar over time. From the area of high dredging intensity, characterizing species tended to be infaunal species typically associated with sandy sediments. Juvenile animals also typified high intensity samples. This suggests an active process of recolonization by juvenile animals invading the dredged deposits. In contrast, those species characterizing the areas of low dredging intensity and reference areas were typically larger and included both infaunal and epifaunal species and these species represented a range of different phyla.
Information from SIMPER and ANOSIM also reveal that the differences between the area of high dredging intensity and the reference areas are more pronounced than those between the area of low dredging intensity and the reference areas. These differences in the sample groups are maintained over the 3-year period of study.
Sidescan sonar surveys
Sidescan sonar surveys were conducted at Area 222 in 2000, 2001, and 2002. Figure 8 shows the output of the sidescan sonar survey conducted in 2002. Operational factors such as weather conditions and the acoustic resolution applied which may affect the quality of the acoustic record have been taken into account when comparing the output from the sidescan sonar surveys over time.
|
The substrata and seabed features within and in the vicinity of Area 222 identified from the 2001 and 2002 sidescan sonar surveys are consistent with those observed in 2000 (Boyd et al., 2003). Disturbed sandy sediments interspersed with patches of sandy gravel and occasional small outcrops of consolidated clay predominate in the northern part of the extraction site. EMS records indicate that this area was subjected to the most intensive dredging activity in the years immediately prior to relinquishment, and some evidence of the effects of the trailer suction hopper dredging consistent with observations made in other studies (e.g. Diesing et al., in press), remains within this part of the site (Figure 9). The sidescan sonar survey conducted in 2001 and 2002 extends to the north of the extraction site and encompasses an area of disturbed seabed previously only surveyed in part in the 2000 survey. This area of seabed to the northeast of the extraction site is uneven, consisting of a series of interconnected pits which is consistent with the effects of static suction hopper dredging (Figure 8). Thus, as noted in previous investigations, it appears that the seabed in this area has been dredged (without a licence) some time prior to the introduction of the EMS in 1993 (Boyd et al., 2003). The area of disturbed seabed extends up to 1000 m away from the northern limit of the extraction site and is characterized by stable slightly muddy sandy gravels, interspersed by patches of clean rippled sand which form the base of the pit structures. Sediment transport features associated with the zone of out of area dredging also appear to extend up to 2500 m away from the northern boundaries of the former extraction site.
|
A number of large sand waves, whose crests run at right angles to the tidal axis, are present 150 m to the north of Area 222. The presence of these features may be the result of deposition and subsequent entrainment of screened sands produced during the dredging activity within and adjacent to Area 222. Furthermore, the sidescan sonar data indicate that those substrata surrounding Area 222 which have not been directly or indirectly affected by historic dredging activity are similar, being composed of a mixture of sand, gravel, and to a lesser extent silt with the occasional outcrop of clay. It should be noted that while sidescan sonar is effective in describing the nature of the sediment surface, it provides no information on buried substrata.
Changes in the physical status of the seabed have been assessed over the surveyed areas between 2000 and 2002. Particular attention has been given to two areas of seabed that show a persisting physical impact from historic dredging activity (Figures 9 and 10). Sidescan imaging shows little change in the nature and the distribution of the substrata over the wider survey area between 2000 and 2002. In each year, the seabed substrata surrounding the areas of dredging impact (including those found at the reference sites) are stable, mixed sediments.
|
Within the licensed extraction site, there appears to be some variability in the spatial distribution of sediments over time. Figure 9 shows a sidescan sonar image of the same area of seabed within and immediately surrounding Area 222, in each of the 3 survey years. These images show that while there has been some redistribution of sandy material within the licensed extraction site there has been little change in the overall amount of sand present. Trailer suction hopper dredge tracks are present within and immediately to the north of the extraction site in all years and there appears to be little modification of their appearance over the duration of these surveys (Figures 9 and 10).
|
Discussion |
|---|
|
TopIntroductionMethodsResultsDiscussionConclusionsReferences |
|---|
A number of studies have attempted to identify and explain distribution trends in benthic assemblages following the cessation of marine dredging (Cressard, 1975; Kenny and Rees, 1994, 1996; Kenny et al., 1998; Desprez, 2000; Sardá et al., 2000; Van Dalfsen et al., 2000; Van Dalfsen and Essink, 2001). Effects on the benthos and sediments identified in such studies show some parallels with the findings observed in this investigation. For example, Desprez (2000) showed that for an industrial extraction site off Dieppe, France, the structure of the benthic community changed from one of coarse sands characterized by the lancelet Branchiostoma lanceolatum (Pallas, 1744) to one of fine sands composed of the infaunal polychaetes Ophelia borealis (Quatrefages, 1866), Nephtys cirrosa (Ehlers, 1868), and Spiophanes bombyx (Claparède, 1870). Thus, the change in the assemblage structure reflected a change in sediment composition caused by dredging. Significant changes in particle size composition, resulting in a net fining of the sediment within extraction sites, have also been reported by Van Dalfsen et al. (2000) and Sardá et al. (2000) following sand extraction.
At Area 222, sediments collected from the northern sector of the extraction site tended to contain proportionally more sand and less gravel than most of the other sampled sediments. This part of the extraction site is coincident with the location of intensive dredging recorded by the EMS (see also Boyd et al., 2003). Sidescan sonar images of the extraction site also confirm the presence of dredge tracks within this area. Within the extraction site, screening of the dredged cargoes was undertaken to increase the gravel content with the return of finer material (usually sands) to the sea by means of a reject chute. Over time, this screening activity has the potential to significantly change the composition of sediments within a dredged area. Therefore, it is likely that, over a 25-year period of dredging, the intensively dredged areas of seabed at Area 222 have also undergone a similar transformation to those documented in the literature and thus have become sandier over time.
There is a large variability in the sediment characteristics sampled within the northern part of the extraction site. Presumably, this represents the uneven impact of the dredger draghead on the seafloor. Further evidence of the patchy nature of substrata is provided by sidescan sonar images of the dredged locations. This variability among replicate samples was also evident in biological samples collected from the area of high dredging intensity. Indeed, a high variability in the composition of sediments and benthic assemblages at dredged locations has been reported by Kenny and Rees (1994) and Sardá et al. (2000). Such observations lend further support to the hypothesis of Warwick and Clarke (1993), namely that variability in assemblage structure may be an identifiable symptom of perturbed conditions. In this case, however, a higher variability in assemblage structure at the extraction areas appears to be the result of increased habitat heterogeneity within the intensively dredged site which affects both the identity of species and the variation in abundances of each species. This propensity for extraction sites to exhibit variability in terms of sediment characteristics and species composition also has to be referenced against a high degree of natural variability and small-scale sediment patchiness that can be encountered in benthic ecosystems, even at locations like Reference site 2, which appear superficially to be relatively homogeneous.
In contrast to other studies that have demonstrated the rapid degradation of dredge tracks after cessation of dredging (Millner et al., 1977; Kenny et al., 1998), it appears that substantially longer periods, i.e. at least 9 years, are required for the complete erosion of dredge tracks in the disturbed area to the northeast of Area 222. Furthermore, the maintenance of a biological assemblage composed of juvenile animals at the site of high dredging intensity up to 6 years after cessation suggests that these species are unable to reach maturity owing to the unstable nature of sediments in the area. Thus, it appears that at the site of high dredging intensity the effects of dredging are still discernible on the composition of sediments and fauna even 6 years after cessation. This is in direct contrast to a body of case studies which together suggest that substantial progress towards restoration of the fauna could be expected within 2–4 years following cessation of marine sand and gravel extraction (Millner et al., 1977; Kenny et al., 1998; Desprez, 2000; Sardá et al., 2000; Van Dalfsen et al., 2000; ICES, 2001). This discrepancy between the Area 222 data and other studies may reflect differences in the magnitude of dredging disturbance, since many of the studies reported in the literature have been concerned with the effects of relatively short-lived dredging campaigns (Kenny et al., 1998; Sardá et al., 2000; Van Dalfsen et al., 2000), whereas Area 222 was dredged repeatedly over a 25-year period. Indeed, evidence from the current study suggests that the nature and speed of recolonization is affected by the intensity of dredging. This may also suggest that undisturbed deposits between dredged furrows provide an important source of colonizing species (Newell et al., 1998), allowing recolonization to proceed more rapidly in less heavily dredged sediments than in areas of intensive dredging. The discrepancy between our findings and those reported in the literature is also likely to reflect differences in the sediment composition at the dredging sites. Sites with sediments containing a higher gravel content typically support a richer assemblage than sandy substrata, and therefore it is to be anticipated that such sites will require a longer time scale for successful regeneration of benthic assemblages (see later).
Local environmental factors are also likely to affect the rate of re-establishment of benthic fauna and in particular the hydrodynamics of a site determine the sedimentary characteristics of an area and will ultimately be responsible for determining broad-scale community patterns (Warwick and Uncles, 1980; Hall, 1994; Rees et al., 1999). It is therefore apparent that any changes in the status of benthic assemblages in areas which have been subjected to extraction will need to be referenced against both variations in particle size and the hydrodynamic regime. Based on the water depth, tidal current, and wave data (Admiralty data), Area 222 is considered to be a site of "moderate" energy. This conclusion is important since it implies that many years, possibly decades, will be required for re-establishment of benthic assemblages in sites classed as "low energy".
The recolonization of dredged sites
From studies of dredged sites (Cressard, 1975; Kenny and Rees, 1994, 1996; Kenny et al., 1998; Desprez, 2000; Sardá et al., 2000; Van Dalfsen et al., 2000; Van Dalfsen and Essink, 2001; Boyd et al., 2003) and from observations following defaunation as a consequence of storm disturbance (Rees et al., 1977), a general pattern of recolonization is emerging (see also ICES, 2001). The first stage involves the settlement of a few opportunistic species, which are able to take advantage of the dredged and sometimes unstable sediments (Hily, 1983; Kenny and Rees, 1994, 1996; Kenny et al., 1998; Desprez, 2000; Van Dalfsen et al., 2000; Van Dalfsen and Essink, 2001). Recolonization can either be by adults or larvae from the surrounding area if the sediments of the disturbed area are similar to the original substrata (Cressard, 1975) or by larvae from more distant sources if the sediment is markedly different (Santos and Simon, 1980; Hily, 1983). These species can substantially increase the overall abundance and the numbers of species during the early stages of post-dredging recolonization (Hily, 1983; López-Jamar and Mejuto, 1988; Kenny and Rees, 1994, 1996; Kenny et al., 1998; Desprez, 2000; Van Dalfsen et al., 2000). A second phase is characterized by a reduced community biomass which can persist for a number of years (Kenny and Rees, 1994, 1996; Kenny et al., 1998; Desprez, 2000). There is a natural expectation that biomass will remain reduced, while new colonizers "grow on" to maturity comparable with the pre-dredging age/size profile (Rees, 1987; Van Dalfsen and Essink, 2001). Furthermore, a reduced biomass may also be caused by the abrasive effects of increased sediment (mainly sand) inhibiting the growth and survivorship of epifauna such as hydroids and bryozoans. Paradoxically, it is this sandy sediment that is also responsible for the infilling of dredge tracks (Kenny et al., 1998; ICES, 2001; Limpenny et al., 2002), which in the longer term may promote physical stability. Over time, it may be expected that, at some sites, the bedload transport will approach the pre-dredged equilibrium, allowing the restoration of community biomass (Kenny et al., 1998). A similar model of response has been represented schematically by Hily (1983) and includes a further stage in which opportunists are replaced by a greater number of species. It was suggested that this replacement was the result of increasing levels of interspecific competition. However, this model was based on observations following the dredging of a sandy mud (Hily, 1983), and further evidence is required to establish whether such oscillations occur in more stable gravel habitats during the latter stages of succession. Evidence from Area 222 suggests that, where there are significant changes to the topography and composition of the sediments as a result of dredging activity, the maintenance of a biological community, over prolonged periods, at an early developmental stage can be expected.
Framework for future studies
As yet, coordinated studies on a wide geographical scale investigating the physical and biological status of commercial aggregate extraction sites in the UK and elsewhere are limited (Van Dalfsen et al., 2000). One consequence of the limited available information on the effects on the benthos of marine aggregate extraction is the difficulty it creates for the establishment of reliable empirical models for predictive purposes. A further difficulty in generalizing about the effects of extraction is the variability in both the dredging history and the particular dredging practices to which different extraction sites are exposed, i.e. a typology of dredging disturbance does not exist. Consequently, when seeking to develop and then apply predictive models, generalizations about the effects of marine aggregate extraction must be qualified by local information regarding the nature of dredging activity and the conditions under which extraction activity occurs. Based on existing evidence, however, the two most commonly encountered scenarios following marine aggregate extraction in the UK are:
- sites where the substratum has changed from a sandy gravel to a gravelly sand;
- sites where the substratum has remained unchanged.
- sites where the substratum has remained unchanged.
This is not to exclude the possibility of other consequences such as the exposure of clay depending on local circumstances (Boyd et al., 2003). In the first of the scenarios, there are a number of ways in which alterations to the sediment as a consequence of dredging could result. These include, but may not be limited to, the exposure of an underlayer of finer sediments, discharge of finer sediments from spillways (Hitchcock and Drucker, 1996; Van Dalfsen et al., 2000) or screening and the trapping of bedload in dredged furrows (Desprez, 2000; Sardá et al., 2000; ICES, 2001). The degree of change appears to depend both on the local circumstances (Desprez, 2000; Van Dalfsen et al., 2000; Boyd and Rees, 2003), and on the magnitude of perturbation i.e. differences in the intensity, type of dredging, or length of extraction period (Van Dalfsen et al., 2000; Boyd et al., 2003; Boyd and Rees, 2003). The colonizing fauna also appear to reflect this change to the substrata, through a shift in the proportions of sandy vs. gravelly fauna (Desprez, 2000). Accompanying this, it is postulated that there would be a net decline in biomass. This model of response is portrayed schematically in Figure 11. A similar model of response could account for changes at some sand extraction sites where the seabed substrata have changed from coarse to fine sands (Sardá et al., 2000; Van Dalfsen et al., 2000).
|
In the second scenario, sediments present at the seabed following the cessation of marine aggregate extraction are similar to those which existed prior to disturbance, i.e. sandy gravels. This scenario accords with current expectations regarding seabed status following licensed dredging and is consistent with the management aim of ensuring that the seabed environment is left in a comparable physical condition to that prevailing prior to the onset of dredging i.e. with a similar sediment type and evenness profile. From the limited available data concerning the effects of marine gravel extraction (Kenny and Rees, 1994, 1996; Kenny et al., 1998), it is reasonable to postulate that the fauna recolonizing such sites will follow classical successional dynamics (e.g. Grassle and Sanders, 1973). Indeed, this scenario was postulated in connection with the disposal of dredged coarse material arising from a port expansion or channel deepening (see Anon., 1996). Although such simplified models require further validation and/or refinement, they provide a useful framework for evaluating the outcome of post-cessation recolonization studies and recovery rates and eventually could provide a reliable predictive capability. A body of case studies on the consequences of marine aggregate extraction over sufficiently long time scales from sites exposed to commercial extraction practices is therefore required to underpin the derivation of reliable and scientifically credible models of response. Such a need applies equally to many other human activities which take place in the marine environment.
|
Conclusions |
|---|
|
TopIntroductionMethodsResultsDiscussionConclusionsReferences |
|---|
Many of the field studies reported in the literature are the results of investigations on the impacts of short-term dredging events, and these have proved useful in determining the rates and processes leading to benthic re-establishment following aggregate extraction. From such studies and those undertaken at sites exploited for commercial interests, a general pattern of response to marine aggregate extraction is emerging. This needs to be tested to establish its general validity in all environments, particularly in areas which have been exposed to industrial scale dredging operations over many years. From such work, it is clear that re-establishment of a community similar to that which existed prior to dredging can only be attained if the topography and original sediment composition are restored (for a contrary view, see Seiderer and Newell, 1999). Should physical stability of the sediments not be attained, however, it is hypothesized that communities will remain at an early developmental stage.
|
Acknowledgements |
|---|
This work was funded by the UK Office of the Deputy Prime Minister, The Department for the Environment, Food and Rural Affairs (Project code AE0915) and the Crown Estate. This work was also supported in 2004 with funding from the MEPF Aggregate Levy sustainability Fund (Project code C2228). The authors would also like to thank the following individuals for their input to this work: Clare Morris for sediment particle size analysis, Bill Meadows for sidescan sonar surveys, and Rebecca Kilbride for processing benthos samples at sea. We are also grateful to Unicomarine limited for contract analysis of macrobenthic samples.
|
References |
|---|
|
TopIntroductionMethodsResultsDiscussionConclusionsReferences |
|---|
-
Anon. (1996) Monitoring and assessment of the marine benthos at UK dredged material disposal sites. The Scottish Office Agriculture, Environment and Fisheries Department, Aberdeen, Scotland. Scottish Fisheries Information Pamphlet 21. 35 pp.
Boyd S. E. (2002) Guidelines for the conduct of benthic studies at aggregate dredging sites. (Compiler) UK Department for Transport, Local Government and the Regions, London and CEFAS, Lowestoft. 117 pp.
Boyd S.E., Limpenny D.S., Rees H.L., Cooper K.M., Campbell S. (2003) Preliminary observations of the effects of dredging intensity on the re-colonization of dredged sediments off the south-east coast of England (Area 222). Estuarine, Coastal and Shelf Science 57:209–223.[CrossRef]
Boyd S.E. and Rees H.L. (2003) An examination of the spatial scale of impact on the marine benthos arising from marine aggregate extraction in the Central English Channel. Estuarine, Coastal and Shelf Science 57:1–16.[CrossRef]
Bray J.R. and Curtis J.T. (1957) An ordination of the upland forest communities of the Southern Wisconsin. Ecological Monographs 27:325–349.[CrossRef]
CIRIA. (1996) Beach recharge materials – demand and resources. Construction Industry Research and Information Association. Report 154, London. 174 pp.
Clarke K.R. (1993) Non parametric multivariate analyses of changes in community structure. Australian Journal of Ecology 18:117–143.[CrossRef][Web of Science]
Clarke K.R. and Gorley R.N. (2001) PRIMER v. 5 User Manual/Tutorial(PRIMER-E Ltd, Plymouth) 91 pp.
Clarke K.R. and Warwick R.M. (1994) Change in Marine Communities: an Approach to Statistical Analysis and Interpretation(Plymouth Marine Laboratory, Natural Environment Research Council, UK) 144 pp.
Cressard A. (1975) The effects of offshore and gravel mining on the marine environment. Terra et Aqua 8/9:24–33.
CSTT. (1997) Comprehensive studies for the purposes of Article 6 & 8.5 of DIR 91/271 EEC, the Urban Waste Water Treatment Directive, second edition. Published for the Comprehensive Studies Task Team of Group Coordinating Sea Disposal Monitoring by the Department of the Environment for Northern Ireland, the Environment Agency, the Scottish Environmental Protection Agency and the Water Services Association. 60 pp.
Desprez M. (2000) Physical and biological impact of marine aggregate extraction along the French coast of the eastern English Channel: short- and long-term post-dredging restoration. ICES Journal of Marine Science 57:1428–1438.
Diesing M., Zeiler K., Klein H. Comparison of marine sediment extraction sites by means of shoreface zonation. Journal of Coastal Research Special Issue 39 (in press).
Eleftheriou A. and Holme N.A. (1984) Macrofauna Techniques. In Holme N.A. and McIntyre A.D. (Eds.). Methods for the Study of Marine Benthos 2nd edn (Blackwell Scientific Publications, Oxford, UK) pp. 140–216 387 pp. (ch. 6).
Grassle J.F. and Sanders H.L. (1973) Life histories and the role of disturbance. Deep Sea Research 20:643–659.[Web of Science]
Green R.H. (1979) Sampling Design and Statistical Methods for Environmental Biologists(John Wiley & Sons, New York) 257 pp.
Hall S.J. (1994) Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology: an Annual Review 32:179–239.
Hily C. (1983) Macrozoobenthic recolonisation after dredging in a sandy mud area of the Bay of Brest enriched by organic matter. Oceanologica Acta. Proceedings of the 17th European Marine Biology Symposium27 September to 1 October 1982Brest, France pp. 113–120M.
Hitchcock D.R. and Drucker B.R. (1996) Investigation of benthic and surface plumes associated with marine aggregates mining in the United Kingdom. Proceedings of the Oceanology International Conference 1996The Global Ocean – Towards Operational Oceanography (Spearhead Exhibitions Ltd, Surrey) vol. 2: pp. 221–234.
ICES. (2001) Report of the ICES Working Group on the effects of extraction of marine sediments on the marine ecosystem. ICES Co-operative Research Report 247, Copenhagen, Denmark. 80 pp.
Kenny A.J. and Rees H.L. (1994) The effects of marine gravel extraction on the macrobenthos: early post dredging recolonization. Marine Pollution Bulletin 28:442–447.[CrossRef][Web of Science]
Kenny A.J. and Rees H.L. (1996) The effects of marine gravel extraction on the macrobenthos: results 2 years post-dredging. Marine Pollution Bulletin 32:615–622.[CrossRef][Web of Science]
Kenny A. J., Rees H. L., Greening J., Campbell S. (1998) The effects of marine gravel extraction on the macrobenthos at an experimental dredge site off North Norfolk, UK (results 3 years post-dredging). ICES CM 1998/V: 14. 14 pp.
Limpenny D. S., Boyd S. E., Meadows W. J., Rees H. L., Hewer A. J. (2002) The utility of sidescan sonar techniques in the assessment of anthropogenic disturbance at aggregate extraction sites. ICES CM 2002/K: 04. 20 pp.
López-Jamar E. and Mejuto J. (1988) Infaunal benthic recolonisation after dredging operations in La Coruna bay, NW Spain. Les Cahiers de Biologie Marine 29:37–49.
Millner R. S., Dickson R. R., Rolfe M. S. (1977) Physical and biological studies of a dredging ground off the east coast of England. ICES CM 1977/E: 48. 11 pp.
Newell R.C., Seiderer L.J., Hitchcock D.R. (1998) The impact of dredging works in coastal waters: a review of the sensitivity to disturbance and subsequent recovery of biological resources on the sea bed. Oceanography and Marine Biology: an Annual Review 36:127–178.
Rees E.I.S., Nicholaidou A., Laskaridou P. (1977) The effects of storms on the dynamics of shallow water benthic associations. In Keegan B.F., Ceidigh P.O., Boaden P.J.S. (Eds.). Biology of Benthic Organisms(Pergamon Press, Oxford) pp. 465–474 630 pp.
Rees H. L. (1987) A survey of the benthic fauna inhabiting gravel deposits off Hastings, Southern England. ICES CM 1987/L: 19. 19 pp.
Rees H.L., Pendle M.A., Waldock R., Limpenny D.S., Boyd S.E. (1999) A comparison of benthic biodiversity in the North Sea, English Channel, and Celtic Seas. ICES Journal of Marine Science 56:228–246.
Santos S.L. and Simon J.L. (1980) Response of soft-bottom benthos to annual catastrophic disturbance in a south Florida estuary. Marine Ecology Progress Series 3:347–355.[Web of Science]
Sardá R., Pinedo S., Gremare A., Taboada S. (2000) Changes in the dynamics of shallow sandy-bottom assemblages due to sand extraction in the Catalan Western Mediterranean Sea. ICES Journal of Marine Science 57:1446–1453.
Seiderer L.J. and Newell R.C. (1999) Analysis of the relationship between sediment composition and benthic community structure in coastal deposits: implications for marine aggregate dredging. ICES Journal of Marine Science 56:757–765.
Singleton G.H. (2001) Marine aggregate dredging in the U.K: a review. Journal of the Society for Underwater Technology 25:3–14.
Skalski J.R. and McKenzie D.H. (1983) A design for aquatic monitoring programs. Journal of Environmental Management 14:237–251.[Web of Science]
Somerfield P.J. and Clarke K.R. (1997) A comparison of some methods commonly used for the collection of sublittoral sediments and their associated fauna. Marine Environmental Research 43:145–156.[CrossRef][Web of Science]
Van Dalfsen J.A. and Essink K. (2001) Benthic community response to sand dredging and shoreface nourishment in Dutch coastal waters. Senckenbergiana Maritima 31:329–332.
Van Dalfsen J.A., Essink K., Toxvig Madsen H., Birklund J., Romero J., Manzanera M. (2000) Differential response of macrozoobenthos to marine sand extraction in the North Sea and the western Mediterranean. ICES Journal of Marine Science 57:1439–1445.
Warwick R.M. and Clarke K.R. (1993) Increased variability as a symptom of stress in marine communities. Journal of Experimental Marine Biology and Ecology 172:215–226.[CrossRef][Web of Science]
Warwick R.M. and Uncles R.J. (1980) Distribution of benthic macrofauna associations in the Bristol Channel in relation to tidal stress. Marine Ecology Progress Series 3:97–103.[Web of Science]
Wentworth C.K. (1922) A scale of grade and class terms for clastic sediments. Journal of Geology 30:377–392.
CiteULike 
Connotea 
Del.icio.us What's this?
This article has been cited by other articles:
|
|
 |
|
  E. Bayer, R. A. Barnes, and H. L. Rees The regulatory framework for marine dredging indicators and their operational efficiency within the UK: a possible model for other nations? ICES J. Mar. Sci., November 1, 2008; 65(8): 1402 - 1406. [Abstract] [Full Text] [PDF]
|
|
|
|
|
|
P. D. Eastwood, C. M. Mills, J. N. Aldridge, C. A. Houghton, and S. I. Rogers Human activities in UK offshore waters: an assessment of direct, physical pressure on the seabed ICES J. Mar. Sci., April 1, 2007; 64(3): 453 - 463. [Abstract] [Full Text] [PDF]
|
|
| ||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||