Long-term changes in polychaete assemblages of Botany Bay (NSW, Australia) following a dredging event

Marine Pollution Bulletin 52 (2006) 997–1010 www.elsevier.com/locate/marpolbul

Viewpoint

Long-term changes in polychaete assemblages of Botany Bay (NSW, Australia) following a dredging event Ceridwen Fraser a

a,b,*

, Pat Hutchings b, Jane Williamson

a

Marine Ecology Group, Department of Biological Sciences, Macquarie University, North Ryde, NSW 2109, Australia b Aquatic Zoology, Australian Museum, 6 College Street, Sydney, NSW 2010, Australia

Abstract The long-term effects of marine aggregate dredging on near-shore benthic assemblages are still largely unknown, despite a global increase in demand for, and extraction of, marine aggregates. This study assessed the state of recovery of polychaete assemblages in Botany Bay, temperate NSW, Australia, at sites dredged for aggregate material more than 10 years previously. Sedimentary and faunal samples were collected from impact sites in Botany Bay, and from reference sites in nearby Pittwater estuary. This study was based on, and included data from, a study conducted by the Australian Museum at the same sites in the 2 years following cessation of dredging. Abundance, species richness and evenness of polychaetes, as well as overall polychaete assemblage structure, were compared between localities over time.  2006 Elsevier Ltd. All rights reserved. Keywords: Dredging; Impacts; Botany Bay; Recovery; Taxonomic resolution

1. Introduction Dredging is one of the most common anthropogenic disturbances in marine soft-sediment environments, and dredging of marine aggregates, for use in construction, land reclamation and beach replenishment, has become an important and growing industry in recent decades (ICES, 2001). Impacts of dredging on benthos can be direct, by displacement or burial (Newell et al., 1998), or indirect via a permanent change in environmental factors such as depth, turbidity, sediment characteristics and water flow patterns (Jones and Candy, 1981; ICES, 2001). Sediments in the dredged area are often completely defaunated initially (Newell et al., 1998), and although recolonisation of impact sites is generally rapid (Newell et al., 1998 and

*

Corresponding author. Address: Aquatic Zoology, Australian Museum, 6 College Street, Sydney, NSW 2010, Australia. Tel.: +61 2 93206140. E-mail addresses: [email protected], [email protected] (C. Fraser). 0025-326X/$ - see front matter  2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2005.12.016

references therein), the long-term environmental impacts of marine aggregate extraction are usually site-specific and difficult to predict (Kenny and Rees, 1996; Newell et al., 1998; Desprez, 2000; ICES, 2001; Boyd et al., 2004). Recovery of the seabed and benthos following dredging depends on a variety of factors, such as the intensity of dredging, and the nature of the benthos. Sites that have been dredged repeatedly over several years may take longer to recover than sites dredged over short periods (Boyd et al., 2004, 2005). Availability of new recruits also influences the speed and success of recovery (Rosenberg, 1977), as does the horizontal migration of adults from refuge patches or surrounding unaffected areas (van Dalfsen et al., 2000). Small-scale dredging disturbances can recover more rapidly than spatially extensive disturbances (Hall, 1994). Local hydrodynamics are a major controlling factor in recovery processes of the seabed and its benthos (Van Der Veer et al., 1985; Kenny and Rees, 1996; Boyd et al., 2004). For similar faunal assemblages to occur at dredged sites, the original topography, sedimentary characteristics and hydrodynamics must be restored (Boyd et al., 2004). Dredging scars can remain for decades at low energy sites

998

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

(Boyd et al., 2004). Recovery processes and durations in soft-sediment communities cannot, therefore, be easily compared between sites with different physical characteristics. Benthic faunal assemblages of fine-grained sedimentary environments tend to recover more rapidly than those of coarse-grained sedimentary environments (see review in Newell et al., 1998). Few studies have followed dredged sites to complete recovery, and although recovery may be reported as well underway by the end of monitoring (Kenny and Rees, 1996; Newell et al., 2004), the actual duration of a complete return to a pre-dredging state is usually predicted rather than measured (Boyd et al., 2004). This study examined the long-term effects of a dredging event on polychaete assemblages of Botany Bay, Sydney, Australia. From late 1992 until late November 1993 areas of the seafloor of Botany Bay (NSW, Australia) were dredged to provide infill for a third parallel airport runway at Sydney (Kingsford Smith) Airport. The Australian Museum studied the recolonisation and recovery of the sites from April 1994 until October 1995. Reference sites with similar physical characteristics, such as depth and sediment type, were chosen in nearby Pittwater estuary. The dredging removed much of the coarse aggregate material from the sediment, deepened the sites and, initially, destroyed most of the benthos. By the end of the study, the benthic assemblages were still markedly different to those of the reference sites (Wilson, 1998). This study revisited the same sites as the 1994/1995 study to assess any longer term changes in benthic assemblages. A pilot study carried out in Botany Bay in July 2004, as well as data from the previous study, showed that polychaetes dominated the benthos at these sites. Polychaetes are an extremely diverse group, with a wide range of reproductive strategies and feeding guilds (Hutchings, 1998, 2003), and are likely to be suitable surrogates for soft-sediment faunal assemblages (Olsgard and Somerfield,

2000). Our study, using data from the 1994/1995 study as well as data from samples collected in September and December 2004, used polychaetes to indicate the state of recovery of dredged areas. We addressed whether changes had occurred in the following parameters at impact sites: (1) sediment grain size characteristics; (2) water depth; (3) species richness, abundance and evenness of polychaetes; (4) polychaete assemblages. We also examined whether taxonomic resolution affected the interpretation of results. 2. Materials and methods 2.1. Study areas Sites at two locations within the Sydney region, Botany Bay and Pittwater, were examined (Fig. 1). These sites were also used in the Australian Museum’s Federal Airports Commission (AMFAC) study of 1994–1995. 2.1.1. Botany Bay Botany Bay is located within the Sydney metropolitan region (3355 0 S, 15111 0 E). It is a large (49.1 km2), shallow bay with a wide (1.1 km) mouth. Hydrodynamics in the bay have been altered by anthropogenic constructions with hard edges, such as the large shipping container port and the parallel airport runway on the northern side of the bay. The foreshores of Botany Bay have been well developed, with land use ranging from heavy industrial (e.g., chemical manufacture) to residential. A few natural areas have been preserved. 2.1.2. Pittwater Pittwater is located 32 km (3336 0 S, 15119 0 E). Although bury estuary system, Pittwater having no major watercourses

to the north of Sydney part of the large Hawkesreceives little freshwater, emptying directly into it

Fig. 1. Site locations. The locality ‘Impact A’ comprised the sites Bot 1 and Bot 2, the locality ‘Impact B’ comprised Bot 3 and Bot 4, and the locality ‘Reference’ comprised Pit 1 and Pit 2.

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

(Hutchings and Recher, 1974). It is 5.5 km long and has an area of 17.3 km2. The eastern and southern foreshores are developed (primarily residential) and most of the northwestern foreshores are part of Ku-ring-gai National Park. 2.2. Sampling 2.2.1. Sites Four impact sites in Botany Bay were monitored. Two of the sites, Bot 1 and Bot 2, were located to the southeast and south of the new airport runway (Fig. 1). Sites Bot 3 and Bot 4 were located in a relatively sheltered position to the west of the new runway. On 1 May 2001 Botany Bay became a Recreational Fishing Haven. NSW Fisheries banned all commercial fishing, including prawn trawling, in the Bay from that date. Prior to May 2001 commercial prawn trawling was carried out in Botany Bay and may have occurred at Bot 1 and Bot 2 between 1993 and 2001. Two of the impact sites of this study, Bot 3 and Bot 4, were in a restricted area, and no prawn trawling is assumed to have occurred at these sites after the parallel airport runway commenced operation in 1994. At the time of dredging, there were no areas in Botany Bay that (a) offered similar depth and sedimentary characteristics to the post-dredging impact sites, and (b) were adequately protected from dredging effects to act as reference sites for the study (Wilson, 1998). Reference sites with similar sediments and depths were therefore chosen from the nearby Pittwater estuary (sites Pit 1 and Pit 2). A network of submarine cables traverses Pittwater near the study sites. The presence of the cables ensured that no benthic trawling could have occurred at the sites between the Australian Museum’s Federal Airports Commission (AMFAC, 1994–1995) and this current (2004) study. The four Botany Bay impact sites were considered as two ‘localities;’ Impact A (sites Bot 1 and Bot 2) and Impact B (sites Bot 3 and Bot 4). This was for two reasons. Firstly, the two localities were separated, while the sites within them were close together; Impact B was between the airport runways, while Impact A was to the southeast of the parallel runway. Secondly, dredging continued at the Impact B sites slightly later than at the Impact A sites, and it could therefore be supposed that some differences in recovery may have taken place. To balance the design, the two reference sites in Pittwater were considered as a single reference locality. 2.2.2. Sampling periods Sites were sampled in mid September 2004 and early December 2004. This placed sampling periods in separate seasons and allowed observation of as much small-scale temporal variation in the benthos as possible within the timeframe of this study, thereby minimising the confounding effects of natural variation. All sampling was carried out over two consecutive days.

999

Four replicate samples were taken from each site; i.e., eight were collected from each locality, in two groups. The locations of replicate samples within each site were randomly selected. One 0.05 m2 Van Veen grab sample was taken from within 50 m of each sample location. Processing of samples was carried out as in the AMFAC 1994–1995 study. Grab samples were immediately sieved through 1 mm mesh bags, placed in 7% formalin/seawater with Biebrich Scarlet vital dye, and, after 3 days, transferred to ethanol (Smith and Rule, 2000; Newell et al., 2004). Polychaetes that could not be identified to lower than family level (eight animals) were not included in the analyses. The species list used for the analysis of 1994– 1995 data was updated to incorporate name changes due to taxonomic revisions to allow direct comparisons to be made. Analyses were performed at generic and family levels as well, permitting this study (incorporating the original AMFAC study data) to provide further evidence in the debate over the use of low taxonomic resolution data. A reference collection was deposited at the Australian Museum. 2.2.3. Sediment samples Sediment samples were sorted into different size classes by washing though a series of sieves, as per Morrisey et al. (2000), of mesh sizes 2 mm, 415 lm, 250 lm and 75 lm. Each fraction was oven-dried at 80 C, and weighed. Samples with grains greater than 2 mm diameter were classified as gravel, and those with grains smaller than 75 lm diameter were classified as mud. All sediments between 75 lm and 2 mm were considered sand. 2.3. Analyses Data were analysed using GMAV5 for Windows (Underwood et al., 1997). Analyses of variances (ANOVAs) were carried out on sediment, depth, species richness, abundance, and evenness data. Data that did not meet the assumptions of normality and heteroscedasticity were transformed. Abundance data were transformed by the natural log (x + 1); and evenness data by arc sine (proportion). Sediment grain size data were analysed using two-factor ANOVAs, with localities as a fixed factor and time as a random factor. As the proportion of gravel in any sample was negligible throughout the study, the proportion of mud was essentially complementary to the proportion of sand. Either the ‘percentage mud’ or the ‘percentage sand’ values should therefore adequately represent the sedimentary characteristics of any sample in this study. Sediment data could not be compared between sites, as only one sample was taken from each site. Replication within localities, using the samples from each of the two sites within each locality, nonetheless allowed analysis of the data and comparison of sedimentary characteristics between localities.

1000

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

Depth, species richness, abundance and evenness data were analysed using three-factor, partially nested ANOVAs with the factors time, site and locality, with site nested within locality. The overall difference between localities was not of major interest in this study; of greater importance was the difference between temporal trajectories of localities. This was shown by the interaction term between time and locality, with a significant interaction indicating that temporal trajectories were not the same at all localities. Student–Newman–Keuls (SNK) tests were then carried out to determine when significant differences existed. SNK tests were also used to examine smaller-scale spatial variation, shown by differences between sites within localities at each time. Multivariate analyses were used to compare the polychaete fauna of the localities, using PRIMER, version 5 (Clarke and Gorly, 2001). Bray–Curtis similarity matrices (Bray and Curtis, 1957) were created on fourth-root transformed data. Rare species were excluded from the data prior to analysis to prevent their masking the major differences between samples. Three primary data sets were created for analysing: (1) differences between samples associated with the immediate effects of dredging at the impact sites (April 1994 and May 1994); (2) the state of recovery (similarity of samples from impact and reference localities) at the end of the AMFAC study (September 1995 and October 1995), and (3) the state of recovery some 10 years after completion of dredging (September 2004 and December 2004). Non-metric multidimensional scaling (nMDS) ordinations were performed on the similarity matrices to show differences in assemblage structure between samples. Significance was determined by two-way crossed analyses of similarities (ANOSIM), as described in Clarke and War-

wick (2001), at species, generic and family levels. Similarity percentage (SIMPER) analyses were performed to determine which species were responsible for differences between localities. The abundances of the five most dominant species were compared at each locality over time to show successional patterns. 3. Results Ninety-nine polychaete species from 32 polychaete families were identified. Approximately 70% of the species found in September and December 2004 were also found in the AMFAC study of 1994–1995. All ANOVAs indicated that significant (time · locality interaction p < 0.05) differences existed between localities at some stage during the two studies (Figs. 2–6). The general trends of all faunal response variables shown at species level were also apparent at generic and family levels, although fewer significant differences between localities at different times were observed (data not presented). Sites were occasionally significantly different within localities. This small-scale variation did not obscure the larger patterns. 3.1. Changes in depth and sedimentary characteristics The temporal trajectories of sedimentary characteristics were not significantly different between localities (time · locality interaction F = 1.26, p = 0.250). Impact B was, however, significantly muddier overall than either Impact A or the reference locality (F = 28.78, p = 0.000) (Fig. 2). While the localities apparently followed the same pattern of change over time (as shown by the time · locality interaction), Impact B appears to have become much sandier between August 1995 and September 2004.

Fig. 2. Mean percentage mud in sediment samples from each locality at all sampling times. Gravel was almost totally absent from each sample, so the approximate percentage of sand at each locality can also be gauged from this plot.

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

1001

Fig. 3. Mean water depth at each locality.

Fig. 4. Mean species richness of polychaetes per sample at each locality. Letters along the top of the chart show where localities were significantly different at each sampling period, according to SNK tests, with (a) Impact A different to the reference locality, (b) Impact B different to the reference locality, and (c) Impact A different to Impact B.

As with the sedimentary characteristics, the temporal trajectories of water depth were not significantly different between localities (time · locality interaction F = 1.22, p = 0.236). Although the inclusion of sites as a factor in the model prevented any test for overall differences between localities, Fig. 2 shows that Impact A was several metres deeper than either Impact B or the Pittwater reference locality throughout most of the AMFAC study of 1994–1995 (Fig. 3). 3.2. Species richness The temporal trajectories of species richness were significantly different between localities (time · locality

interaction F = 2.14, p = 0.015) (Fig. 4). Although the two impact localities initially showed similar patterns, with species richness increasing in the first months following dredging, they diverged as early as August 1994, and from June 1995 Impact B (between runways) consistently had significantly lower species richness than Impact A (Fig. 4). Species richness at the reference site was consistently higher throughout the majority of sampling periods (Fig. 4). Impact B did not show the same patterns of species richness, in relation to the reference locality, as did Impact A. Impact B consistently had significantly lower species richness than the reference locality, throughout the study period, with the exception of September 2004.

1002

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

Fig. 5. Mean abundance of polychaetes per sample at each locality, scaled logarithmically. Letters along the top of the chart show where localities were significantly different at each sampling period, according to SNK tests, with (a) Impact A different to the reference locality, (b) Impact B different to the reference locality, and (c) Impact A different to Impact B.

Fig. 6. Mean Pielou’s evenness of polychaete species at each locality. Evenness values did not differ significantly between localities at any sampling period between January 1995 and December 2004.

3.3. Abundance Temporal trajectories of abundance were significantly different between localities (time · locality interaction F = 2.45, p = 0.006). Abundances of polychaetes were similar at impact localities over time (Fig. 5), with the exception of August 1994. Abundances were low at each site in the first months following dredging, but increased until they peaked in December 1994, after which they appear to have stabilised (Fig. 5). Impact A reached similar abundances to the reference locality by August 1994. In December 1994, and in all subsequent sampling periods, both Impact A and Impact B did not have significantly different abundances to the reference locality. While SNK tests showed that Impact A was significantly different to the reference locality in October 1994,

the two sites (Bot 1 and Bot 2) within the locality had significantly different abundances at this time, with samples from Bot 2 having less than 15 individuals each, while samples from Bot 1 had up to 880. Although the mean abundance value for the locality was therefore similar to that of the reference locality, the variation within the locality was great, and a significant difference was found between localities at this time. 3.4. Evenness The spread of abundances over species within samples (evenness) did not differ significantly between any localities from January 1995 until the end of the study (time · locality interaction F = 0.86, p = 0.5988) (Fig. 6). Evenness values could not be calculated for all localities prior to

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

January 1995, due to low abundance values at impact localities in early sampling periods. 3.5. Multivariate analyses Non-metric multidimensional scaling (nMDS) ordinations showed that impact localities were relatively similar to each other, whereas the reference locality had a substantially different assemblage structure to either of the impact sites, shortly after dredging took place (Fig. 7a). This pattern was still evident, although less severe, at

1003

the end of the AMFAC study (Fig. 7b) and in September and December 2004 (Fig. 7c). ANOSIM results showed that all localities differed significantly in polychaete assemblage structure at each time (Table 1), although impact localities were closer to each other than to the reference locality in the nMDS plots. Analysing at different taxonomic levels (species, genus and family) did not change any of the ANOSIM comparisons. In the data set for September and December 2004 the two sampling periods were also significantly different (R 0.124, p = 0.005).

Fig. 7. Non-metric multidimensional scaling (nMDS) ordination of samples from each locality in (a) April and May 1994, (b) August and October 1995, and (c) September and December 2004.

Table 1 Results of the two-way crossed ANOSIM testing for differences in polychaete assemblage structure in each of the temporal datasets analysed in this study Dataset

Comparisons

R

p

April 1994, May 1994

Global test Pairwise Impact A vs Impact B Impact A vs Reference Impact B vs Reference

0.678

0.001

0.151 0.98 0.993

0.01 0.001 0.001

Global test Pairwise Impact A vs Impact B Impact A vs Reference Impact B vs Reference

0.893

0.001

0.633 0.983 0.994

0.001 0.001 0.001

Global test Pairwise Impact A vs Impact B Impact A vs Reference Impact B vs Reference

0.69

0.001

0.478 0.811 0.828

0.001 0.001 0.001

August 1995, October 1995

September 2004, December 2004

1004

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

3.5.1. April 1994–May 1994 In the first two sampling periods following dredging, several species were primarily responsible for the differences between localities (Table 2a), according to the SIMPER analysis. These differences are not particularly important, however, given the extremely low abundances of all polychaetes at the impact sites in these months. Of greater interest is the presence or absence of particular species. Euchone limnicola (Family Sabellidae) was found at Impact B but not at Impact A, and Polycirrus sp. 1 (Family Terebellidae) was found at Impact A but not at Impact B. The majority of the differences between impact and reference localities in earlier sampling dates were due to low abundances or absences at the impact localities of most of the species at the reference locality. The main exception was Sthenelais pettiboneae (Family Sigalionidae), which was found in lower abundances at the Reference locality than at either of the impact localities. 3.5.2. August 1995–October 1995 The differences between the impact localities in August and October 1995 were mainly due to greater abundances of species at Impact A than at Impact B (Table 2b). Two species that were in high abundances at Impact A were absent at Impact B (Dipolydora socialis (Family Spionidae) and Maldane sarsi (Family Maldanidae)). Polycirrus sp. 1 was a dominant species (averaging > 20% of the total abundance) at both localities, but was found in far greater abundances at Impact A than Impact B (Table 2b). Euchone limnicola, on the other hand, was a dominant species at Impact B but was present in relatively low abundances at Impact A in these months (Table 2b). The differences between both impact localities and the reference locality were mainly due to greater abundances of Levinsenia gracilis (Family Paraonidae), Microclymene sp. 2 (Family Maldanidae), Aricidea (Allia) sp. (Family Paraonidae) and Scalibregma inflatum (Family Scalibregmatidae) at the reference locality (Table 2b). Microclymene sp. 2 was a dominant species at the reference locality, but was completely absent from Impact A, and was present only in low abundances at Impact B. 3.5.3. September 2004–December 2004 Mediomastus australiensis (Family Capitellidae) was a dominant species at both impact localities in September and December 2004 (Table 2c), but was more abundant at Impact B, accounting for 9.1% of the differences between these localities. E. limnicola, which was more abundant at Impact B than Impact A throughout the study (Table 2c), was responsible for 7.6% of the difference. A significant proportion of the differences between Impact A and the reference locality was due to two species, Microclymene sp. 2 and Mediomastus australiensis. Microclymene sp. 2 was found at high abundances at the reference locality but at relatively low abundances at Impact A, while the reverse was true for M. australiensis. A very similar pattern was shown for Impact B vs Reference.

Two other major contributors to the differences between Impact B and the reference locality were Levinsenia gracilis, which was absent from Impact B and present at considerable abundances at the reference locality, and E. limnicola, which was present in high abundances at Impact B and at low abundances at the reference locality. 3.6. Succession of dominant species The successional patterns at both impact localities showed three polychaete species replacing each other as dominant (Fig. 8a and b). The sabellid E. limnicola was the first to reach high abundances, appearing in high numbers at Impact A in August 1994, and peaking at both localities in December 1994, before declining sharply by January 1995. The terebellid Polycirrus sp. replaced E. limnicola as the dominant in the following months, peaking in June 1995 at Impact B, and in August 1995 at Impact A, before declining. Mediomastus australiensis, a capitellid, was a dominant species at both impact localities in September and December 2004, although E. limnicola was co-dominant at Impact B. The reference locality did not show the replacement patterns observed at the impact localities (Fig. 8). The dominant species at the reference locality were Microclymene sp. 2, L. gracilis, and Polycirrus sp. (Fig. 8c). Although the abundances of these three species fluctuated somewhat over time, they remained the characteristic dominants of the locality throughout the study, although Polycirrus sp. was at negligible abundances in September and December 2004 at all three localities. 4. Discussion The long-term effects of marine aggregate dredging on near-shore benthic assemblages are still largely unknown (Boyd et al., 2004), despite a global increase in demand for, and extraction of, marine aggregates (ICES, 2001). This study assessed the state of short (18 months) and long-term (10 years) recovery of polychaete assemblages in Botany Bay, Australia. In this study, recovery was defined as when the levels of response variables at impact sites reached those of nonimpacted reference sites. This definition is useful when the overall health and functioning of faunal assemblages at the impact sites is of higher priority than a return to the physical and biological conditions that occurred there prior to dredging. 4.1. Depth Dredging has a substantial long-term impact on the physical characteristics of soft-sediment environments. The depths of the impact localities in this study were still markedly deeper in 2004 than pre-dredging depths of 6–8 m (Australian Museum, 1993). This is surprising, considering the type of sediment. Depth can affect benthic

Table 2 Summary of the results of SIMPER species breakdowns in (a) April and May 1994 (the first two sampling periods following dredging), (b) August and October 1995 (the last two sampling periods of the AMFAC study), and (c) September and December 2004 (the sampling periods of this study)

(a) April 1994 - May 1994

(b) August 1995 - October 1995

Impact A

Reference

Euchone limnicola Aphelochaeta sp. 1 Sthenelais pettiboneae Impact B

Polycirrus sp. 1

Levinsenia gracilis Polycirrus sp. 1 Maldane sarsi Pseudopolydora paucibranchiata Augeneria verdis Micronephthys sp. 2

Impact B

Chaetozone setosa Leitoscoloplos sp. 1 Pseudopolydora paucibranchiata Maldane sarsi Sthenelais pettiboneae Euchone limnicola

Impact A

Reference

Levinsenia gracilis Microclymene sp. 2 Maldane sarsi Aricidea (Alllia) sp. Euchone limnicola Mediomastus australiensis Scalibregma inflatum Terebellides woolawa Levinsenia gracilis Microclymene sp. 2 Dipolydora socialis Mediomastus australiensis Aricidea (Alllia) sp. Chaetozone setosa Scalibregma inflatum

(c) September 2004 - December 2004 Impact A

Reference

Mediomastus australiensis Microclymene sp. 2 Galathowenia sp. Mediomastus australiensis Euchone limnicola Euchone limnicola Impact B Prionospio tridentata Levinsenia gracilis Spiochaetopterus sp. 1 Galathowenia sp. Leitoscoloplos sp. 1 Prionospio tridentata Chaetozone setosa Augeneria verdis Mediomastus australiensis Microclymene sp. 2 Levinsenia gracilis Impact A Aricidea (Allia) sp. Prionospio tridentata Goniada maculata

1005

Species names are aligned such that they are closest to the locality in which they were most abundant; for example, in September–December 2004, Euchone limnicola was more abundant at Impact B than at Impact A, while Galathowenia sp. was more abundant at Impact A than at Impact B. Bold font indicates species that made up greater than 20% of the abundance of the samples in one locality, while bold and underlined font.

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

Levinsenia gracilis Polycirrus sp. 1 Maldane sarsi Impact A Pseudopolydora paucibranchiata Augeneria verdis Sthenelais pettiboneae Micronephthys sp. 2

Impact A Dipolydora socialis Polycirrus sp. 1

1006

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

Fig. 8. Abundances of dominant species at (a) Impact A, (b) Impact B, and (c) Reference, over time. (Note differences in scales of abundance between figures.)

polychaete assemblages (Hutchings, 1998), and is an important consideration in assessing post-impact recovery. Fine-grained sediments, such as the sandy mud found in Botany Bay, should be quite easily resuspended in the water column, and dredging scars in areas with a high proportion of fine-grained sediments should fill in more rapidly than those with coarser, heavier grains (Desprez, 2000; ICES, 2001). Dredging scars can, however, endure for decades in low-energy areas (Boyd et al., 2004). Hydrodynamic conditions between the runways (Impact B) were probably very low-energy, and sediment movement at this locality may not have been great. The sites at the end of the parallel runway (Impact A), on the other hand, were in an area that is exposed to the major water movements of the Bay, and sediment movement would be expected to be considerable at these sites. That the dredging pits were still deep at Impact A in 2004 suggests that sediment movement was actually minimal in this area, and that the locality was too deep to be affected by wave action. This has implications for management of dredged sites. Deeply dredged pits can become deoxygenated (ICES, 2001 and references therein), so, if dredging is to take place in deep locations, beyond the reach of wave-induced water movement, dredging wider areas of seafloor in shallow bands may be better than dredging one or two points deeply. Although greater areas of the surface of the seafloor would be damaged by

dredging through this method, the refuge areas left between dredged strips could also facilitate horizontal migration and recolonisation of these sites (van Dalfsen et al., 2000). 4.2. Sedimentary characteristics Sediments at the impact localities had a high proportion of mud shortly after dredging. This is consistent with the findings of most dredging impact studies, which have found that dredged pits often initially develop a layer of fine sediments (see review in Newell et al., 1998). Fine sediments are more easily suspended in the water column, and it is therefore hardly surprising that they are the first to settle in dredging scars. The difference in sediment characteristics between the three sites throughout the 1994–1995 sampling periods suggests that hydrodynamic conditions were different at these localities. During 2004, the between runways sites (Impact B) were observed to be in an area that had relatively calm water even when much of the Bay was wind affected. Following dredging, fine sediments suspended in the water column may have settled more easily in the calm water overlying Impact B than at Impact A. By 2004, the proportion of mud at Impact B had dropped considerably, and the locality appeared to have similar sedimentary characteristics to Impact A and the reference locality. As this

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

change occurred at some stage during the 10 years between studies, it is not possible to determine whether it was rapid or gradual. Strong storm events in the intervening years may have led to suspension and resettlement of coarser sediments even in the protected area between the runways. Alternatively, the change may have been progressive and slow, with sand gradually moving down the walls of the dredged pit from surrounding areas. Sediment stability depends on many factors, including grain size characteristics and levels of bioturbation (Morrisey et al., 2000). Bioturbation may have played a role in the redistribution of sand at Impact B. 4.3. Faunal recovery Impact areas were almost completely defaunated during dredging (Wilson, 1998), and all organisms occurring at these sites must, therefore, have arrived post-dredging, either through vertical migration or recruitment (Hall and Frid, 1998). Considering the size of the areas dredged, and the limited mobility of many benthic polychaetes (Hutchings, 2000), the availability of larvae would have a substantial impact on recolonisation. The rate of recolonisation at dredged sites reflects the high dispersive capabilities of many polychaetes (Wilson, 1991). It took less than a year for polychaete abundances at the impact localities to reach, and remain at, levels similar to those of the reference locality. This is consistent with the findings of other postdredging studies. Newell et al. (2004) found that infauna at dredged sites showed rapid recolonisation and achieved high abundances, but comparatively low species richness and biomass, within months of dredging. Bonvicini Pagliai et al. (1985) also found that abundances reached reference levels within 6 months of dredging. Although polychaete abundances at Impact B (between runways) reached and remained at reference levels by December 1994, just over 12 months from cessation of dredging activity, species richness levels at this locality remained lower than reference levels throughout the two studies. This was in contrast to Impact A, at which species richness levels reached, and generally remained at, reference levels from August 1994. Thus, fewer polychaete species were present at Impact B, but those present occurred at high abundances. This may have been due to a restricted supply of larvae to this locality, limiting the number of species that could establish there; to larvae selecting not to settle there based on environmental conditions; or to differential survival of species under different environmental conditions. Any of these scenarios are plausible, given the broad range of life-history capabilities of this diverse group (Hutchings, 2003), and the different responses of species to disturbances in general (Connell, 1978; Pearson and Rosenberg, 1978; Goodsell and Connell, 2005). The replacement patterns observed at the impact localities in this study suggest a facilitation model of succession, in keeping with that observed by Lu and Wu (2000) in their study of recolonisation and succession of benthos at a sub-

1007

tropical location. In the model, early colonisers alter the habitat in such a way that other species are able to flourish in their stead (Connell and Slatyer, 1977). Euchone limnicola is a small, tube-dwelling, exotic sabellid that has previously been found at high abundances near ports (Parry et al., 1997; Currie and Parry, 1999). Population densities of this species can fluctuate greatly over time, and have been observed to peak in summer in southeast Australia (Parry et al., 1997), concurrent with the findings of this study. The peak in E. limnicola following dredging may, therefore, have been due to its reproductive cycle rather than any particular opportunism on its part, although the dredged areas may have presented an ideal habitat for settlement of larvae. Unfortunately, the lack of any suitable reference sites within Botany Bay make the distinction between natural fluctuations and opportunism difficult to determine in this instance. Polycirrus and Mediomastus are both genera that have been described as opportunistic following disturbances (Salen-Picard et al., 2003 and references therein), and their roles in the replacement sequences of this study could well have been as facilitators for later colonisers. Each is capable of modifying the microhabitat, for example by construction of tubes (in the case of E. limnicola) or by bioturbation. M. australiensis was only dominant at the impact sites in the 2004 sampling periods, suggesting that it may be a later-stage successor. Another member of this genus, M. ambiseta, has been observed to be a fairly late coloniser of disturbed sediment (Grassle and Grassle, 1974). The relative stability of the dominant members of polychaete assemblages at the reference locality during this study contrasts strongly with the replacement patterns of the impact localities. Whether the impact localities achieved stable assemblages by the end of the study is not certain, as the two sampling periods in 2004 were inadequate to show fluctuations in dominant species over time. The early successional stages are, however, apparent. Functional redundancy, whereby an ecosystem function can be performed by several species (Snelgrove et al., 1997; Hutchings, 1998), may have accounted for some differences in polychaete assemblage structure between localities. M. australiensis, a deposit-feeding capitellid, was found at high abundances at both impact localities, but only in low abundances at the reference locality, in the later sampling periods. A deposit feeding maldanid, Microclymene sp. 2, was, however, found at comparable abundances at the reference localities. M. australiensis and Microclymene sp. 2 may perform similar ecosystem functions, and so the differences between the reference and impact localities may be less than observed. This study would, ideally, have made use of several reference localities, preferably with one or more in Botany Bay itself. The use of a single reference locality was necessitated by a lack of suitable sites in both Botany Bay and other bays (Wilson, 1998). Although the original AMFAC study included a reference site in Port Hacking as well as those in Pittwater, the Port Hacking site was so small,

1008

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

and had such different sedimentary characteristics (Wilson, 1998), that it was not considered a suitable reference site for this study. The trends and impacts that occurred at each locality in the 10 years between October 1995 and September 2004 are not known. As Botany Bay is a heavily used area for both shipping and fishing, we can assume that there have been considerable impacts at varying scales during this period. Impact B, lying between two high-use airport runways, almost certainly receives runoff that includes contaminants such as heavy metals and petrochemicals. The Australian Museum’s final report to the Federal Airports Commission (Wilson, 1998) suggested that the area between the runways was at risk of becoming anoxic and faunistically depauperate. Despite a low diversity of polychaetes at Impact B, this extreme situation has not occurred. 4.4. Taxonomic resolution There is much controversy over the appropriate taxonomic resolution for benthic studies, with issues including processing time and level of expertise needed cited as reasons for identifying animals only to family level or higher (Dauvin et al., 2003). Identifying fauna to species level in environmental assessment studies has many benefits. Species are distinct biological units and substantially more can therefore be known about assemblages when species are identified, and data from studies where fauna was identified only to higher taxa cannot effectively be compared to any other studies (Hutchings, 1999). Comparison between the results of this study and the previous AMFAC study was greatly facilitated by comprehensive species-level data being available from the earlier study, as well as actual reference material that could be compared with the animals collected in 2004. One reason given in support of studies at lower taxonomic resolutions is that identification to species level by non-specialists is excessively difficult (Olsgard and Somerfield, 2000). Most polychaetes from the samples collected during this study were identified by a non-specialist, assisted largely by new tools such as an interactive electronic identification key (Wilson et al., 2003). Relatively few (<20%) of the polychaete species found had not yet been described in the literature, and all could be identified to genus. Unless emphasis is put on the importance of specieslevel identification, the need to describe species will no longer exist, and the problem will be exacerbated (Maurer, 2000). Although several studies, including this one, show that trends in macrobenthic assemblages can often be seen at family level (e.g., Vanderklift et al., 1996; Olsgard and Somerfield, 2000; Macfarlane and Booth, 2001; Roach et al., 2001; de Biasi et al., 2003; Olsgard et al., 2003; Thompson et al., 2003), most have also observed some loss of information compared to analyses at species level. In this study, the conclusions would not have been the same if polychaetes had only been identified to family level. Overall family richness and abundance trends would have

been similarly interpreted, but the importance of differences in assemblage structure, as determined by multivariate analyses, as well as patterns of succession, could not have been examined in detail. Olsgard et al. (2003) suggest that more studies comparing the results of analyses at different taxonomic levels need to be done across the globe before the effects of studies at low taxonomic resolutions can be properly understood. Analysis of data at various taxonomic levels, after identification of organisms to species, should be encouraged in impact-assessment studies in order to better gauge the viability of the use of lower taxonomic resolutions in studies that are severely restricted by time or money. 5. Conclusions and recommendations Aggregate dredging in marine sedimentary environments can affect benthic assemblages both in the shortand the long-term. This study suggests that long-term (years to decades) changes can vary substantially from short-term (months to years) changes in both the physical environment and the structure of polychaete assemblages. It would be interesting to assess these sites in another 10 years to further understand long-term effects. More studies are needed to adequately assess the longterm impacts of aggregate dredging on marine sedimentary biota in Australia. Studies should carry out monitoring of impact and reference sites until recovery is complete, rather than for an arbitrary length of time as is currently the case. The long-term effects of aggregate dredging have thus far been assessed primarily by studies that have, like this study, revisited dredged sites after a period (several years) has lapsed since initial post-dredging sampling (e.g., Desprez, 2000; Boyd et al., 2004). While such studies provide important information on the state of recovery of faunal assemblages at the time of revisitation, more studies are needed that systematically sample over long periods, to adequately show trends in long-term recovery. Acknowledgements This work was carried out in Aquatic Zoology at the Australian Museum. Thanks to all staff in Marine Invertebrates and Marine Ecology at the Australian Museum for their assistance with this work, particularly Alan Jones, who gave invaluable ecological advice, George (Buz) Wilson, Anna Murray, Kate Attwood and Maria Capa. Thanks also to Buz Wilson for permission to use data from the AMFAC study of 1994–1995. Thanks to Ian Graham for assistance with sedimentary analyses. References Australian Museum, 1993. Australian Museum Biological Survey of Botany Bay – Phase 1 Final Report, June 1993. Sydney: Australian Museum, 217 pp. Bonvicini Pagliai, A.M., Cognetti Varriale, A.M., Crema, R., Curini Galletti, M., Vandini Zunarelli, R., 1985. Environmental impact of

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010 extensive dredging in a coastal marine area. Marine Pollution Bulletin 16, 483–488. Boyd, S.E., Cooper, K.M., Limpenny, D.S., Kilbride, R., Rees, H.L., Dearnaley, M.P., Stevenson, J., Meadows, W.J., Morris, C.D., 2004. Assessment of the re-habilitation of the seabed following marine aggregate dredging. Science Series Technical Report, CEFAS Lowestoft, No. 121, 154 pp. Boyd, S.E., Limpenny, D.S., Rees, H.L., Cooper, K.M., 2005. The effects of marine sand and gravel extraction on the macrobenthos at a commercial dredging site results 6 years post-dredging. ICES Journal of Marine Science 62, 145–162. Bray, J.R., Curtis, J.T., 1957. An ordination of the upland forest communities of the Southern Wisconsin. Ecological Monographs 27, 325–349. Clarke, K.R., Gorly, R.N., 2001. PRIMER v. 5 User Manual Tutorial. PRIMER-E Ltd., Plymouth, 91 pp. Clarke, K.R., Warwick, R.M., 2001. Change in Marine Communities: An Approach to Statistical Analysis and Interpretation, second ed. PRIMER-E, Plymouth, 172 pp. Connell, J.H., 1978. Diversity in tropical rainforests and coral reefs. Science 199, 1302–1310. Connell, J.H., Slatyer, R.O., 1977. Mechanisms of succession in natural communities and their role in community stability and organization. American Naturalist 111, 1119–1144. Currie, D.R., Parry, G.D., 1999. Changes to benthic communities over 20 years in Port Phillip Bay, Victoria, Australia. Marine Pollution Bulletin 38, 36–43. Dauvin, J.C., Gomez Gesteira, J.L., Salvande Fraga, M., 2003. Taxonomic sufficiency: an overview of its use in the monitoring of sublittoral benthic communities after oil spills. Marine Pollution Bulletin 46, 552–555. de Biasi, A.M., Bianchi, C.N., Morri, C., 2003. Analysis of macrobenthic communities at different taxonomic levels: an example from an estuarine environment in the Ligurian Sea NW Mediterranean. Estuarine Coastal and Shelf Science 58, 99–106. 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. Goodsell, P.J., Connell, S.D., 2005. Historical configuration of habitat influences the effects of disturbance on mobile invertebrates. Marine Ecology Progress Series 299, 79–87. Grassle, J.F., Grassle, J.P., 1974. Opportunistic life histories and genetic systems in marine benthic polychaetes. Journal of Marine Research 32, 253–284. Hall, S.J., 1994. Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology: An Annual Review 32, 179–239. Hall, J.A., Frid, C.L.J., 1998. Colonisation patterns of adult macrobenthos in a polluted north sea estuary. Aquatic Ecology 33, 333–340. Hutchings, P., 1998. Biodiversity and functioning of polychaetes in benthic sediments. Biodiversity and Conservation 7, 1133–1145. Hutchings, P., 1999. Taxonomy of estuarine invertebrates in Australia. Australian Journal of Ecology 24, 381–394. Hutchings, P.A., 2000. Natural History. In: Polychaetes and Allies: the Southern Synthesis. In: Beesley, P.L., Ross, G.J.B., Glasby, C.J. (Eds.), Fauna of Australia, vol. 4A, Polychaeta, Myzostomida, Pogonophora, Echiura, Sipuncula. CSIRO Publishing, Melbourne, pp. 32–38. Hutchings, P.A., 2003. Polychaetes – their biological diversity. Records of the South Australian Museum Monograph Series 7, 39–49. Hutchings, P., Recher, H.F., 1974. The fauna of Careel Bay with comments on the ecology of mangrove and sea-grass communities. The Australian Zoologist 18, 99–128. ICES, 2001. Effects of Extraction of Marine Sediments on the Marine Ecosystem. ICES Cooperative Research Report No. 247, 84 pp.

1009

Jones, G., Candy, S., 1981. Effects of dredging on the macrobenthic infauna of Botany Bay. Australian Journal of Marine and Freshwater Research 32, 379–398. Kenny, A.J., Rees, H.L., 1996. The effects of marine gravel extraction on the macrobenthos: results 2 years post-dredging. Marine Pollution Bulletin 32, 615–622. Lu, L., Wu, R.S.S., 2000. An experimental study on recolonization and succession of marine macrobenthos in defaunated sediment. Marine Biology 136, 291–302. Macfarlane, G.R., Booth, D.J., 2001. Estuarine macrobenthic community structure in the Hawkesbury River, Australia: relationships with sediment physicochemical and anthropogenic parameters. Environmental Monitoring and Assessment 72, 51–78. Maurer, D., 2000. The dark side of taxonomic sufficiency (TS). Marine Pollution Bulletin 40, 98–101. Morrisey, D., Turner, S., MacDiarmid, A.B., 2000. Subtidal assemblages of soft substrata. In: Kingsford, M., Battershill, C. (Eds.), Studying Temperate Marine Environments: A Handbook for Ecologists. Canterbury University Press, Christchurch and CRC Press, Boca Raton, pp. 194–226, 335 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. Newell, R.C., Seiderer, L.J., Simpson, N.M., Robinson, J.E., 2004. Impacts of marine aggregate dredging on benthic macrofauna off the south coast of the United Kingdom. Journal of Coastal Research 20, 115–125. Olsgard, F., Somerfield, P.J., 2000. Surrogates in marine benthic investigations – which taxonomic unit to target? Journal of Aquatic Ecosystem Stress and Recovery 7, 25–42. Olsgard, F., Brattegard, T., Holthe, T., 2003. Polychaetes as surrogates for marine biodiversity: lower taxonomic resolution and indicator groups. Biodiversity and Conservation 12, 1033–1049. Parry, G.D., Currie, D.R., Crookes, D.P., 1997. Exotic marine pests in Portland Harbour and environs. Marine and Freshwater Resources Institute Report No. 1. Pearson, T.H., Rosenberg, R., 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: An Annual Review 16, 229– 311. Roach, A.C., Jones, A.R., Murray, A., 2001. Using benthic recruitment to assess the significance of contaminated sediments: the influence of taxonomic resolution. Environmental Pollution 112, 131–143. Rosenberg, R., 1977. Effects of dredging operations on estuarine benthic macrofauna. Marine Pollution Bulletin 8, 102–104. Salen-Picard, C., Arlhac, D., Alliot, E., 2003. Responses of a Mediterranean soft bottom community to short-term (1993–1996) hydrological changes in the Rhone river. Marine Environmental Research 55, 409– 427. Smith, S.D.A., Rule, M.J., 2000. The effects of dredge-spoil dumping on a shallow water soft-sediment community at Coffs Harbour NSW. Part 2. Results of additional monitoring. Report prepared for Coffs Harbour City Council, Coffs Harbour, NSW, June 2000. Snelgrove, P., Blackburn, T.H., Hutchings, P.A., Alongi, D.M., Grassle, J.F., Hummel, H., King, G., Koike, I., Lambshead, P.J.D., Ramsing, N.B., Solis-Weiss, V., 1997. The importance of marine sediment biodiversity in ecosystem processes. Ambio 26, 578–583. Thompson, B.W., Riddle, M.J., Stark, J.S., 2003. Cost-efficient methods for marine pollution monitoring at Casey Station, East Antarctica: the choice of sieve mesh-size and taxonomic resolution. Marine Pollution Bulletin 46, 232–243. Underwood, A.J., Chapman, M.G., Richards, S.A., Sage, M.B., 1997. GMAV5 for Windows [CD-ROM]. Institute of Marine Ecology, Sydney. van Dalfsen, J.A., Essink, K., Toxvig Madsen, H., Birklund, J., Romero, J., Manzanera, M., 2000. Differential response of macrozoobenthos to

1010

C. Fraser et al. / Marine Pollution Bulletin 52 (2006) 997–1010

marine sand extraction in the North Sea and the Western Mediterranean. ICES Journal of Marine Science 57, 1439–1445. Vanderklift, M.A., Ward, T.J., Jacoby, C.A., 1996. Effect of reducing taxonomic resolution on ordinations to detect pollution-induced gradients in macrobenthic infaunal assemblages. Marine Ecology Progress Series 136, 137–145. Van Der Veer, H.W., Bergman, M.J.N., Beukema, J.J., 1985. Dredging activities in the Dutch Wadden Sea: effects on macro-benthic fauna. Netherlands Journal of Sea Research 19, 183–190.

Wilson, W.H., 1991. Sexual reproductive modes in polychaetes: classification and diversity. Bulletin of Marine Science 48, 500–516. Wilson, G.D.F., 1998. A post-impact monitoring study of benthic fauna in areas dredged for the third parallel airport runway in Botany Bay. Report prepared by the Australian Museum’s Marine Invertebrate Section for the Federal Airports Commission. Available from: . Wilson, R.S., Hutchings, P.A., Glasby, C.J., 2003. Polychaetes: An Interactive Identification Guide. CSIRO Publishing, Melbourne.