The use of offshore meiobenthic communities in laboratory microcosm experiments: response to heavy metal contamination

Marine Biology and Ecology, 211 (1997) 247-261

Journal of Experimental

ELSEVIER

JOURNAL OF EXPERIMENTAL MARINE BIOLOGY AND ECOLOGY

The use of offshore meiobenthic communities in laboratory microcosm experiments: response to heavy metal contamination Melanie C. Austen*, Andrea J. McEvoy Plymouth Marine Laboratory, Prospect Place, West Hoe, Plymouth PLl 3DH, UK

Received 3 April 1996; revised 27 August 1996; accepted 26 September

1996

Abstract A microcosm, originally developed for intertidal, estuarine meiobenthic communities, has been used to determine the effects of the heavy metals copper, zinc, cadmium and lead on offshore meiobenthic nematode communities. Significant differences were observed in community structure between controls and all metals except cadmium. The dose response of the offshore meiofauna to experimental contamination was rather confusing as copper and zinc low doses appeared to have much more drastic effects than the high doses. We speculate that at the highest copper and zinc dose levels the metals acted as preservatives such that animals died but did not decompose. This indicates that metals will affect the microbial component of the sediment as well as the meiobenthos in this type of experimental design. The response to the contaminants of offshore sediment biota differed from that previously observed in intertidal estuarine biota. This may be because fauna in the estuarine environment are subjected to greater levels of natural physicochemical stress and are therefore more generally tolerant. This suggests that environmental impact assessments should bioassay communities which naturally inhabit the environment to be assessed. The methods used have potential in the development of a community level bioassay particularly since it appears that the dominant nematode component of meiobenthic communities, from a wide range of habitats, can be easily maintained in simple microcosms. 0 1997 Elsevier Science B.V. Keywords:

Bioassay;

Community;

Meiofauna;

Metals; Microcosm;

Nematode

1. Introduction Meiobenthic communities are composed of animals of small size which have short generation times (mostly days to weeks), continuous reproduction all year round and in *Corresponding author 0022.0981/97/$17.00 0 1997 Elsevier Science B.V. All rights reserved PfI SOO22-0981(96)02734-7

248

M.C. Austen,

A.J. McEvoy

I J. Exp. Mar.

Bid.

Ed.

211 (1997)

247-261

situ direct benthic development i.e. with no planktonic larval stage. In addition the communities generally have high densities and high diversity. Because of these features it is quite easy to maintain and manipulate quite natural meiobenthic communities, particularly the normally dominant nematode component, in simple laboratory microcosms (Austen, 1989; ‘Olafsson and Elmgren, 1991; Sundelin and Elmgren, 1991). This has opened up possibilities for manipulating such communities to look at the ecological effects of natural environmental variables (Austen, 1989; ‘Olafsson and Elmgren, 1991) and anthropogenic factors on whole communities (e.g. Warwick et al., 1988; Sundelin and Elmgren, 1991; Austen et al., 1994; Millward and Grant, 1995; Carman et al., 1995). A community level bioassay is being developed using natural, intertidal meiobenthic communities incubated in simple closed microcosm systems with laboratory contaminated or potentially polluted field sediments (Austen et al., 1994; Austen and Somerfield, 1997). These experiments have indicated that this technique can be used to detect differences in the ecological effects of different types of metals at a range of doses and to discriminate field sediments with different contaminant loadings. Austen et al. (1994) also demonstrated that communities from different locations may have different sensitivities to contaminants. If this is generally the case bioassays must be developed which can be firstly ecologically relevant to the point of discharge and hence the impact of the contaminant, and secondly also be used to differentiate potentially sensitive areas to contamination. Meiobenthic communities are highly structured by the granulometry of the sediment substrate in which they live but the previous studies have shown that for estuarine communities the microcosms can be used to maintain fauna from both muddy and sandy sediments (Austen et al., 1994). Due to the extremely variable environmental conditions in estuaries (salinity and temperature, desiccation etc.) it might be expected that intertidal estuarine meiobenthic communities are sufficiently robust to withstand laboratory manipulation. In contrast communities from offshore sub-tidal areas are subjected to much less environmental variation and therefore may be more susceptible to disturbance effects during laboratory manipulation making them much less reliable in microcosm experiments. However, providing such communities can be maintained and used in laboratory microcosm experiments they may also be more sensitive to the effects of contamination. Additionally it would be advantageous in the development of a meiobenthic community level bioassay if the technique could be applicable in as many different environments as possible. In the present study the microcosm technique has been used to incubate meiobenthic communities in offshore sediments with defaunated sediment (also collected offshore) contaminated with four metals at a range of concentrations. The methodology is very similar to that employed by Austen et al. (1994) so that the results should be directly comparable.

2. Methods The methods are fully described in Austen et al. (1994). Briefly, offshore sediment and meiofauna were collected from a series of 0.1 m* box cores from Rame Head near

M.C.

Ausren, A.J. McEvoy

I J. Exp. Mar.

Biof. Ed.

211 (l997)

247-261

249

Plymouth, south-west England (SO0 18.40’N 04” 14.96’W), at a muddy sand site with a water depth of 45-50 m. Sediment to be amended with metals was collected in advance by scraping the surface 5 cm of sediment from the box cores. This was alternately frozen and thawed three times to defaunate it and then amended with metals as chlorides in seawater stock solutions. Metals were mixed into the sediment with a food mixer and amended sediment was left to equilibrate for 2 days at 5°C before microcosms were assembled. Targeted dose levels (Table 1) were chosen using data from Bryan and Langston (1992) on metals from 19 estuaries to correspond with low to high levels of contamination. There were three experimental concentrations of copper, zinc and lead corresponding to low, medium and high. A fourth metal treatment, cadmium, was used at a high concentration only since previous experiments had shown that even high concentrations of cadmium had no effect on estuarine meiobethic nematode communities (Austen et al., 1994). At the end of the experiment a small aliquot of sediment from one replicate from each treatment was analysed for the dosed metal. Sediment samples were dried at 80°C digested in concentrated HNO, and, after evaporation, the residues were dissolved in 1 M HCl. Metal concentrations were determined by flame atomic absorption using a Varian Spectra AA20 atomic absorption spectrometer with autosampler. Background correction was used for Cd and Pb. An air/acetylene flame was used for all metals. Live meiobenthic communities were collected in surface scrapes of the top 5 cm of sediment from box cores on 16 November 1993 and were transported back to the laboratory on the same day for immediate use. Microcosms consisted of 570-ml narrow-mouthed glass bottles, stoppered with a rubber bung with two holes and aerated via an air-stone diffuser (Fig. 1). A 72-g aliquot of fresh, meiofauna rich sediment was mixed in the microcosms with 72 g of sediment which had been amended with twice the required final concentration of contaminant. Control samples were the same as treatment samples except that 72 g of defaunated, uncontaminated sediment from Rame Head was used in place of the contaminant amended sediment. Each microcosm was then gently filled with filtered (1 pm) natural sea water at 35%0 salinity. There were four replicate microcosms for each metal contaminated treatment and thirteen control replicates which were randomly located spatially within crates. The microcosms were kept in the dark, initially at the field temperature on the day of collection but this was gradually raised at a rate of I-2°C per day until a steady experimental temperature of 20°C was reached. Experimental temperatures were higher than field temperatures to stimulate and optimise conditions for nematode reproduction and growth. The experiment was maintained for two months. Five replicate control microcosms, the initial control samples, were sampled at the beginning of the experiment and the remaining treatments and controls were sampled at the end. Control samples taken at the end of the experiment are hereafter referred to as the control samples. Individual microcosms were sampled by pouring the overlying water through a 63p.m sieve thus retaining any suspended meiofauna which was then backwashed into a sample pot. The remaining mud from the microcosm bottle was also poured and washed into the pot and the sample was then fixed in 5% formalin. Meiofauna was extracted using a modification of the flotation in Ludox technique involving three Ludox (specific gravity 1.15) extractions (Austen and Warwick, 1989). Since nematodes are the overwhelmingly

250

M.C. Austen, A.J. McEvoy I J. Exp. Mar. Biol. Ecol. 211 (1997) 247-261

air intake

Fig. 1. Schematic diagram of microcosm.

dominant meiobenthic taxa in Rame sediment only these were identified to genus, or species where practicable, and enumerated. Samples were mounted on slides in anhydrous glycerol, coded and analysed “blind” to remove bias i.e. the identity of the treatment was not known by the analyst during identification and enumeration. During microscope analysis the first 200 nematodes were identified on the slide and the remainder counted. Corrections were made for the proportion of the total sample identified. Data analysis followed methods described by Clarke and Warwick (1994); Clarke (1993) using the PRIMER (Plymouth Routines in Multivariate Ecological Research) software package. Multivariate data analysis was by non-metric multi-dimensional scaling ordination (MDS) with the Bray-Curtis similarity measure and using a range of transformations depending on the data. Untransformed data analysis is more sensitive to

251

M.C. Austen, A.J. McEvoy I J. Exp. Mar. Biol. Ecol. 211 (1997) 247-261

changes in dominant species but increasingly severe square root and double square root transformations are more sensitive to changes in species composition. Pairwise ANOSIM tests were carried out to determine if there were significant differences between meiobenthic assemblages in different treatments. It should be noted that in ANOSIM tests p values cannot be adjusted to allow for multiple comparisons as the tests are not independent. SIMPER was used to determine the contribution of individual species towards dissimilarity between treatments and controls. Univariate indices were computed: abundance, number of species, Shannon-Wiener diversity using log, (H’). species richness and evenness. ANOVA was used to test for overall differences between these indices and the Tukey HSD multiple comparisons test was used in pairwise comparisons of treatments and controls. In all the above statistical significance testing a significant difference was assumed when p -=c0.05.

3. Results 3.1. Sediment

chemistry

Final sediment concentrations from one microcosm replicate for each treatment at the end of the experiment are given in Table 1. The metal concentrations in the high dose treatment sediments were much lower than the targeted values. 3.2. Multivariate

data analysis

The results of significance testing using ANOSIM for differences between treatments and the final controls were the same regardless of data transformation used, therefore the results with the intermediate -\/ data transformation are presented in Table 2. All of the controls and the treatments were significantly different from the initial controls. The lead medium, all the copper treatments and all the zinc treatments were significantly different from the control samples. The cadmium treatment was not significantly different from the control samples. The MDS ordinations shown in Fig. 2 are again all based on d transformation of the data since there were hardly any differences between ordinations based on different transformations. In the MDS ordination with all variables included (Fig. 2a) there is

Table I Targeted concentration of metals (Fg g concentrations at the end of the experiment Control

Copper Cadmium Zinc Lead

’ dry wt.) in Rame Head microcosm Medium

Low

sediments

and

actual

High

Target

Actual

Target

Actual

Target

Actual

Target

Actual

0 0 0 0

55 0 156 56

600

194

1200

1270

705 700

151 247

1410 1400

1376 1343

1800 1.8 2115 2100

1480 2.524 1530 1580

(1) (m) (h) (h)

(h) (1) (m) (h)

Initial control

1.OO* 1.00*

1.oo*

l.OO* 0.73*

1.oo*

1.00* 1.00* l.OO*

1.oo*

0.993*

Final control

1.00* 0.65* 0.35* 0.11 1.00* 1.oo* 0.51* 0.22(0.09) 0.36* 0.27(0.09)

results (I statistic) of pairwise

Cu (1)

0.19 0.51* 0.85* 1.OO* 0.98* 1.00*

1.oo*

0.91* 0.98*

Cu (m)

between

Cu (h)

0.45* 1.00* 0.99* 0.23(0.09) 0.48* 0.76* 0.46*

differences

0.28 0.78* 0.81* 0.81* 0.34(0.09) 0.64* 0.64* 0.61(0.06)

tests for pairwise

* Denotes significant differences when p < 0.05. Numbers given in brackets denote p values when 0.05 Sp 5 0.1

Zn Pb Pb Pb

Zn (I) Zn (m)

Cu Cu Cu Cd

Final Control

Table 2 ANOSIM

Cd (h)

0.26* 0.05 0.60* 0.35(0.09)

1.oo*

1.00*

treatments

using d

Zn (I)

0.27(0.06) 0.78* 1.OO* 0.96* 1.00*

and controls

Zn (m)

0.58* 1.oo* 1.00* 1.00*

transformed

Zn (h)

0.15 0.38* 0.09

nematode

Pb (1)

0.37(0.09) 0.30

abundance

data

Pb (m)

0.33(0.06)

%

2

k ;”

2 % 2

$ ? ti

5 e

5

i h S

9

2

F S “3 L i

M.C.

Austen,

A.J. McEvoy

I J. Exp. Mar.

Bid.

Ed.

control

211 (1997)

1

247-261

253

initial

cu(1) @

Cu(m)

4

Cu(4

v v v0

.a (1) Zn (4 Zn 03 Cd (h)

a

Pb(Il

A

Pb(m)

A

Pb (hi

Fig. 2. MDS ordination of 4 transformed Rame Head nematode data. (a) all treatments (stress = 0.06) (b) controls, copper and zinc treatments (stress = 0.05) (c) controls and lead treatments (stress = 0.17).

almost a straight line horizontal gradient of treatment effect. At the extreme end there is a cluster of initial control samples. Quite close to this cluster the majority of samples are tightly clustered including the controls and cadmium treatment and all the lead treatments. The copper high and zinc high doses are at the peripheral edges of this cluster, then the copper medium dose followed by the zinc medium dose treatments. At the extreme end of the gradient there is a rather loose arrangement of the remaining zinc low and the copper low treatments. The MDS ordinations have been repeated using subsets of the data to look for dose related trends and to look for similarities and differences in the effects of the different contaminants. An MDS ordination of controls, all zinc and all copper dose treatments (Fig. 2b) has the surprising feature that, particularly for zinc, there is a gradient from high to low dose

254

MC. Austen, A./. McEvoy I J. Exp. Mar. Biol. Ecol. 211 (1997) 247-261

treatments but the high dose treatments are most similar to the control and the low dose treatments least similar. This gradient is repeated for the copper treatments but the distinction between high and medium doses is not as marked. The copper low treatments appear to be more dissimilar from the control than the zinc low treatments. In the MDS ordination of lead treatments and controls (Fig. Zc) there do not appear to be any dose related effects. The species complement of the initial controls, controls and treatments remained quite similar but the proportional abundance of each species did vary from the start to the end of the experiment and also where there was a treatment effect. The control samples were dominated by Actinonema pachydermatum. The second ranking dominant species was Axonolaimus sp.. In rank order of abundance there was then a group of common species with similar abundances to each other: Sabatieria sp., Molgolaimus sp., Desmodora pontica and D. tenuispiculum; and then another group of common but even less abundant species: Halichoanalaimus (long tailed species), Comesa sp. and Neotonchus sp., Parasphaerolaimus sp., Dorylaimopsis punctata, Matylinnia complexa and Viscosia sp. Initial control samples were similar in composition but with much higher abundances of most of the common species. Pomponema sp. was quite common in the initial control samples but was very rare in the final control samples. In the copper low and medium treatments there was a reduction in most species compared to the controls with the reduction being much greater in the low dose treatment. In the copper high dose most of the more common species remained abundant, Desmodora pontica had a slightly higher abundance than in the controls but Sabatieria sp. had a 50% reduction in mean abundance. The zinc low and medium treatments also had a reduction in most species. Differences between the two doses were due to slightly higher abundances in the medium dose. The zinc high treatment had a much lower abundance of Axonolaimus sp. than the control where it was quite abundant. Other common but lower abundance species had small shifts in abundance with some increasing whilst others decreased. The copper high dose treatment had higher abundances of the dominant species but lower abundances of both the common (but not dominant) and the rare species than the zinc high treatments. The lead treatment samples (medium and high dose) had much lower abundances of Actinonema so that unlike the controls it was not a dominant species. Otherwise in the lead medium treatment average abundance of most species either decreased slightly e.g. Desmodora pontica, D. tenuispiculum and Sabatieria sp. or stayed roughly constant e.g. Axonolaimus sp. In the lead high dose treatments some species increased slightly e.g. Desmodora pontica, Sabatieria sp. and Axonolaimus sp. whilst others decreased slightly e.g. Dolicholaimus sp., Halalaimus longicaudatus and Viscosia sp. 3.3. Univariate

data analysis

Initial control sample nematode abundance (mean = 3910 standard error = 240) was significantly higher than nematode abundance in all of the control and treatment samples at the end of the experiment. For all the other univariate measures there were no significant differences between the initial controls, the controls and most of the treatments at the end of the experiments (the exceptions are detailed below).

M.C.

Austen,

A.J. McEvoy

/ J. Exp. Mar.

Biol.

Ecol.

211 (1997)

247-261

25s

Univariate measures and pooled confidence intervals at the end of the experiment are plotted graphically in Fig. 3. In comparison with the controls nematode abundance in the zinc low and medium and the copper low treatments was significantly reduced. These

1000 I

500

E

[[,I

I f

0

I

E

1

i

50

-

40

-

30

-

20

-

10

-

11

IE E

E

I

I

1

1

E

I

1

1

1

1

1

OF L 3-

I

2-

I

1

1

1

IOr 765432I -

I

07 1 .oc

0.9 0.8

-

0.7

-

0.6

-

I1

I

11

EIEEEI

0.5r

~$ImhImhImh E 0 8 3

r5

a”

Fig. 3. Graphical summary of means and 95% pooled confidence intervals of univariate indices for nematode communities. The number of replicates was either 8 (control), 3 (lead high) or 4 (all other treatments). H’ is Shannon Wiener diversity; richness is Margalefs d; evenness is Pielous’ .I.

256

M.C. Austen, A.J. McEvoy

I J. Exp. Mar. Biol. Ecol. 211 (1997) 247-261

treatments also had significantly lower numbers of species than the controls. Shannon Wiener diversity was also lower in the copper and zinc low treatments than the controls but only the copper low treatment had a significantly different species richness and evenness from the controls. None of the other treatments differed significantly from the controls in any of the univariate measures.

4. Discussion The response of the Rame Head meiofauna to experimental contamination is rather confusing. The copper and zinc low doses appeared to have much more drastic effects than the high doses. Our experimental practice involving randomisation of microcosm location during the experiment and then “blind” analysis of coded samples to remove analyst bias (the identity of the treatment was not known by the analyst during identification and enumeration of the nematodes) has convinced us that this must be an actual experimental treatment effect. The low dose Rame samples appeared to be poorly preserved with animals in degraded condition and abundances were extremely low. In contrast in the high dose level treatments and the copper medium dose treatments the samples were in better, but not pristine, condition with higher abundances. They were distinguishable from the controls only in the multivariate analysis. The high and medium dose levels were comparable to those used in a previous experiment (Austen et al., 1994) and were deliberately chosen at the extreme end of the range recorded in British waters by Bryan and Langston (1992) to ensure that a stress response was elicited in the meiofauna. Why were the high dose zinc and high and medium dose copper treatments more similar to the controls than lower dose treatments? We speculatively hypothesise that the contaminants at these high doses were so toxic that all the bacteria and meiofauna in the samples were killed and preserved so that complete decomposition of nematodes did not occur. The control samples had low natural levels of metals so it seems unlikely that the bacteria and meiofauna in the sediment at Rame were adapted in any way to tolerate such extreme chemical stress. To partially test this hypothesis we carried out a small scale experiment to compare how effectively meiofauna samples collected in fresh sediment could be preserved in seawater in sealed bottles with either metal contamination, formalin preservation or no treatment. The methods were essentially those described above: Aliquots (90 g wet weight) of fresh mud samples collected from Rame Head were mixed in 570-ml glass bottles with 90 g wet weight defaunated mud which was either contaminated (metal treatments) to attain the same high dose levels of copper, zinc and cadmium as the original experiment or left untreated (controls and formalin treatments). The control and metal treatment bottles were filled with 1 p,m filtered seawater but formalin treatments were filled with 8% formalin in seawater. The bottles were sealed without aeration for a period of two months. All control and metal treatment samples were then fixed in formalin, the meiofauna were extracted and nematode abundances were determined (Table 3). In the copper and zinc treatments there was no reduction in nematode abundance from

M.C. Austen, A.J. McEvoy

I .I. Exp. Mar. Bid. Ed.

211 (1997) 247-261

257

Table 3 Nematode abundances in 180 g (wet weight) sediment (90 g fresh field collected sediment + 90 g defaunated treated sediment) which had been kept in seawater in sealed non-aerated 570 ml bottles for two months Treatment

Targeted

contaminant

or treatment

Nematode

abundance

1

concentration

Replicate

Co&o1

None

F&malin Copper Zinc Cadmium

8% by volume in seawater 1800 pg g ’ sediment (dry weight) 21 I5 pg g -’ sediment (dry weight) 1.8 pgg-’ sediment (dry weight)

220 1232 1060 1184 73

Replicate

2

98 1545 2072 1265 72

the day 0 formalin fixed samples, the nematodes were in rather poor condition but still identifiable to species. In the control samples and cadmium treatments the nematode abundance was an order of magnitude lower than the copper and zinc treatments but the remaining nematodes were in reasonable condition suggesting that they were alive, despite extremely anaerobic conditions, just before the samples were fixed in formalin. The abundance of nematodes in the copper and zinc treatments suggests that at these dose levels the metals acted as a preservative by preventing complete bacterial or fungal breakdown and decomposition of the dead nematodes. The poor condition of the nematodes implies that these contaminants did not kill them very quickly or efficiently and that some autolysis may have taken place. Nematodes are digested extremely rapidly by fish (Gee, 1989; Coull, 1990) and it is probable that their complete decomposition in sediment, observed in the control and cadmium samples, is also usually rather rapid. This result implies a further advantage of using a microcosm experimental approach with whole, natural benthic communities in sediment as a bioassay rather than in vitro methods (e.g. Millward and Grant, 1995). Toxicants may have more direct impact on associated microbial communities in the sediment than the meiobenthos. This may be reflected in meiobenthic community structure because of alteration of food supply or, as in this experiment, altered decomposition processes. Without microbial analysis in parallel with meiobenthos analysis it will not be possible to distinguish which group of organisms are the most sensitive to the pollutant. However, the response of the meiofauna, particularly if a range of concentrations is used for testing, should still act as a sentinel for effects on sediment ecology. Clearly the reduction in abundance of nematodes during the experiment indicates that there is a microcosm effect which, given the extensive manipulation of the sediment and meiofauna, is hardly surprising. However the microcosm effect is uniform across all treatments and most species and is not so great that it masks experimental differences between treatments and controls. Our results indicate that a reasonably complex, species rich and abundant offshore nematode community can be maintained in these experimental systems. The copper and zinc dose levels were similar to those used in the microcosm experiments with sandy Exe and muddy Lynher meiofauna by Austen et al. (1994) yet similar dose responses were not observed in those experiments. The control samples sediments at Rame had low natural levels of metals compared to the sediment in the Lynher estuary but in the Exe estuary metal levels were even lower (Austen et al., 1994).

258

M.C. Amen,

A.J. McEvoy I J. Exp. Mar. Bid. Ecol. 211 (1997) 247-261

The Exe copper and zinc treatment samples at all doses were degraded with low abundances and there was not a dose response gradient in community structure. Differences in sediment properties affect the bioavailability of metals (see discussion below). However, it seems unlikely that in the Exe sediment, which has a lower organic carbon content, metals are less bioavailable than in the Rame sediment. We can at present only speculatively suggest that because the heterotrophic bacteria of the estuarine environment are probably better adapted to tolerate and survive physicochemical stress naturally occurring in an estuary they might be also be less affected by the physiological stress of high metal levels. There is evidence that sediment dwelling bacteria develop metal resistant strains with prolonged exposure (e.g. Pickaver and Lyes, 1981; Hornor and Hilt, 1985) but we have not found similar references to increased general tolerance to stress of estuarine strains of microorganisms. The Rame meiofauna did not appear to be particularly sensitive to cadmium with no effects observed even at the high concentrations used. This was also observed for meiofauna in estuarine sediments (Austen et al., 1994) and is discussed in that paper. Fabian0 et al. (1994) found negative correlations between bacterial biomass and cadmium at similar concentrations to the dose levels we used and cite other work showing negative effects of cadmium on sediment bacteria. The reduced numbers of nematodes in the cadmium treatments of the small experiment described above indicate that in Rame sediment bacteria are much more cadmium resistant. Rame meiofauna was only significantly affected by the medium lead dose and not the low dose. There is evidence that lead is not very toxic to macrofaunal invertebrates (Bryan and Langston, 1992) even at concentrations as high as 2000 pg g ’ dry weight sediment. Our results imply that offshore meiofaunal nematodes may be more sensitive than macrofauna to lead. We conclude from our experiments and those of Austen et al. (1994) that different meiobenthic communities respond to the same contaminants but the extent of the response, the mode of the response and the dose eliciting a response are not constant between communities in different habitats. Of the tested contaminants copper and zinc appeared to elicit the greatest response from the meiofauna but the nature of the response was far from uniform. For example in the Lynher mud, zinc was more toxic than copper and the effect of each metal on community structure was different but in the Exe sand, copper was more toxic than zinc and the pattern of change in community structure was similar for both metals. In the Rame sediment copper and zinc were equally toxic at the low level doses with a similar pattern of change in community structure for both metals. As discussed in Austen et al. (1994) these inconsistencies are probably connected with the nature of the sediment (organic content, redox levels etc.) and seawater characteristics (e.g. salinity) from which the communities are derived which may affect bioavailability (Langston and Spence, 1994; Depledge et al., 1994) rather than being a specific response. Bioavailability and hence toxicity of contaminants depends on their partitioning between the sediment, pore water and overlying water and this can also be dependant on sediment organic carbon content (Di Toro et al., 1991). At lower metal doses some of the metals will be adsorbed onto POC or the sediment and then may not be bioavailable. At higher doses the uptake sites for metal binding to POC

M.C. Austen, A.J. McEvoy

I .I. Exp. Mar. Biol. Ecol. 211 (1997) 247-261

259

and to the sediment itself will be saturated. Hence more metal will be found as free ions in the pore water and, because the microcosms are closed systems, in the overlying water. This would explain the deviations between target values for metal concentrations in Table 1 and actual values at the end of the experiment since during sampling the overlying water in the microcosm was poured away through a 63-pm sieve and metal analysis was carried out on the remaining sediment only. The Lynher mud has the highest particulate organic carbon (POC) content and the Exe sand the lowest with Rame mud intermediate. Where there is a lower organic content in the sediment we would have expected to observe greater toxic effects since the free ions in water are the most bioavailable and toxic forms of metals (Bryan and Langston, 1992). This emphasises that our results with Rame meiofauna at these very high metal levels are not what we could have predicted from sediment parameters alone. However, Somerfield et al. (1994); Millward and Grant (1995); Austen and Somerfield ( 1997) have suggested that some meiofauna species can develop tolerance mechanisms. Differential sensitivity of nematodes from locations which have been subjected to different levels of contamination has been utilised by Millward and Grant (1995) in the development of their in vitro bioassay. Thus variability in response of different communities to contaminants may also be partially attributable to the different levels of environmental contamination to which they have been subjected on comparatively recent timescales. Cat-man et al. (1995) analysed meiofauna communities in microcosm experimental bioassays to study the effects of polynuclear aromatic hydrocarbons (PAH) on estuarine communities and processes. They found that their communities were not very sensitive to PAH and suggested this was probably as a result of chronic exposure. Berge (1990) also found that macrofauna communities from a eutrophic fjord had a much lower sensitivity to oil than those from a clean fjord. This again confirms that different benthic communities will have different sensitivities in different habitats emphasising the requirement for the development of comparatively realistic bioassays which utilise communities which are likely to be targeted by a pollutant. One drawback in the development of microcosm community level experiments as bioassays is that biological interpretation of the results is far from straightforward. Statistical analysis using ANOSIM attached significance to results and in some cases MDS ordinations also showed relationships between toxicant and/or dose and effects on the meiobenthic community structure. However, at the species level meiobenthic response to contaminants does not appear to be specific or uniform. The most extreme response observed was a general fall in abundance of all species. More often the response was an increase in some species, decrease in others and no change in the remainder. No pattern of response could be discerned that was contaminant or dose specific. Within the meiobenthos it is difficult to relate response of individual species to their population ecology since for most species there is little published information. More information is available for some species which have been used in laboratory cultures and in in vitro toxicity testing. These species however tend to be laboratory opportunists and are in fact rarely encountered in field situations (see Coull and Chandler, 1992 for review). Until further life history and toxicity studies are available for a wider range of species, a prospect that seems unlikely in the near future, biological

260

M.C. Ausren, A.J. McEvoy

I J. Exp. Mar.

Biol.

Ecol. 211 (1997)

247-261

interpretation at the more detailed species level in community level bioassays will remain problematic. The basic technique that we have used has discriminated effects between different contaminants and different doses even if the nature of the response is not understood. Our microcosm approach can be used to investigate further responses of meiobenthic communities to metals and other contaminants. This clearly also implies that the technique could be developed as a bioassay to detect significant changes in ecologically realistic whole communities in contrast to more usual methods of toxicity testing carried out with single species or limited artificial assemblages of species whose relevance to local field ecology is often questionable. Our initial results suggest that ecotoxicity testing with meiofauna communities for environmental impact assessment should be carried out directly on the communities which might be affected by a pollutant rather than on a proxy community not necessarily found where the impact is likely to occur. This does not present a problem since meiofauna communities, particularly the nematode component from a range of habitats including offshore subtidal environments, are so easily maintained in simple microcosms.

Acknowledgments This work was jointly funded by the UK Department of the Environment (Project No. PECD 7/7/397) and the National Rivers Authority (Project No. 0354). We are grateful to Steve Widdicombe for technical assistance in the field and Nick Pope, Gary Burt, Bill Langston and Sue Spence for assistance and advice with sediment chemistry, metal spiking and analysis of sediment. Thanks also to Steve Widdicombe, Frode Olsgard, Alistair Lindley, Bruce Coull and an anonymous referee for providing constructive comments on the manuscript.

References Austen, M.C., 1989. Factors affecting estuarine meiobenthic assemblage structure: a multifactorial microcosm experiment. J. Exp. Mar. Biol. Ecol., Vol. 130, pp. 167-187. Austen, M.C., McEvoy, A.J. and R.M. Warwick, 1994. The specificity of meiobenthic community response to different pollutants: Results from microcosm experiments, Mar. Poll. Bull., Vol. 28, pp. 557-563. Austen, M.C. and P.J. Somerfield, 1997. A community level sediment bioassay applied to an estuarine heavy metal gradient. Mar. Env. Res. Vol. 43, pp. 3155328. Austen, M.C. and R.M. Warwick, 1989. Comparison of univariate and multivariate aspects of estuarine meiobenthic community structure. Est. Coast. She(f Sci., Vol. 29, pp. 23-42. Berge, J.A., 1990. Macrofauna recolonisation of subtidal sediments. Experimental studies with crude oil in two Norwegian fjords with unequal eutrophication status. I. Community responses. Mar. Ecol. Prog. Ser., Vol. 66, pp. 103-I 15. Bryan, G.W. and W.J. Langston, 1992. Bioavailability, accumulation and effects of heavy metals in sediments with special reference to United Kingdom estuaries: a review. Environ. Pollur., Vol. 76, pp. 89-131. Carman, K.R., Fleeger, J.W., Means, J.C., Pomarico, SM. and D.J. McMillin, 1995. Experimental investigation of polynuclear aromatic hydrocarbons on an estuarine sediment food web. Mar. Em. Res., Vol. 40, pp. 298-318.

M.C.

Austen,

A.J. McEvoy

I J. Exp. Mar.

Biol.

Ecol. 2/I

(1997)

247-261

261

Clarke, K.R., 1993. Non-parametric multivariate analyses of changes in community structure. Australiarl Aust. J. Ecol.,Vol. 18, pp. 117-143. Clarke, K.R. and R.M. Warwick, 1994. Changes in marine communities: an approach to statistical antr/ysi.\ and interpretation. Natural Environment Research Council, UK. 144 pp. Coull, B.C., 1990. Are members of the meiofauna food for higher trophic levels’? Trans. Am. Microsc. Sot,.. Vol. 109, pp. 233-246. Coull, B.C. and G.T. Chandler, 1992. Pollution and meiofauna: field, laboratory, and mesocosm studies. Ocranogr. Mar. Biol. Annu. Rev., Vol. 30, pp. 191-271. Depledge, M.H., Weeks, J.M. and P. Bjerregaard, 1994. Heavy metals. In, Hundbook of ecotoxicology rolume 2, edited by P. Calow, Blackwell Scientific Publications, Oxford, pp. 799105. Di Toro, M.D., Zarba, C.S., Hansen, D.J., Berry, W.J., Swartz, R.C., Cowan, C.E., Pavlou, S.P., Allen, H.E.. Thomas, N.A. and P.R. Paquin, 1991. Technical basis for establishing sediment quality criteria for non-ionic organic chemicals using equilibrium partitioning. Environ. Toxicol. Chem., Vol. 10, pp. 1541-1583. Fabiano. M., Danovaro, R., Magi, E. and A. Mazzucotelli, 1994. Effects of heavy metals on benthic bacteria in coastal marine sediments: a field result. Mar. Pollut. Bull., Vol. 28, pp. 19-23. Gee, J.M., 1989. An ecological and economical review of meiofauna as food for fish. Zool. J. Linn. SW.. Vol. 96, pp. 243-261. Homor, S.G. and B.A. Hilt, 1985. Distribution of zinc tolerant bacteria in stream sediments. Hydrohiol., Vol. 128, pp. 155-166. Langston, W.J. and S.K. Spence, 1994. Metal analysis. In, Handbook rfl ecotoxicology, Volume 2, edited by P. Calow, Blackwell Scientific Publications, Oxford, pp. 45-78. Millward, R.N. and A. Grant, 1995. Assessing the impact of copper on nematode communities from a chronically metal-enriched estuary using pollution-induced community tolerance. Mar. follut. Bul/..Vol. 30, pp. 70 I-706. ‘Olafsson, E. and R. Elmgren, 1991. Effects of biological disturbance by benthic amphipods Monoporeitr afinis on meiobenthic community structure: a laboratory approach. Mar. Ecol. Prog. Ser., Vol. 74, pp. 99-107. Pickaver, A.H. and M.C. Lyes, 1981. Aerobic microbial activity in surface sediments containing high or low concentrations of zinc taken from Dublin Bay, Ireland. Est. Coast. Shelf Sci., Vol. 12, pp. 13-22. Somertield, P.J., Gee, J.M. and R.M. Warwick, 1994. Soft sediment meiofaunal community structure in relation to a long-term heavy metal gradient in the Fal estuary system. Mar. Ecol. frog. Ser., Vol. 105, pp. 79-88. Sundelin, B. and R. Elmgren, 1991. Meiofauna of an experimental soft bottom ecosystem-effects of macrofauna and cadmium exposure. Mar. Ecol. Prog. Ser., Vol. 70, pp. 2455255. Warwick, R.M., Carr, M.R., Clarke, K.R., Gee, J.M. and R.H. Green, 1988. A mesocosm experiment on the effects of hydrocarbon and copper pollution on a sublittoral soft-sediment meiobenthic community. Mar. Ecol. frog. Srr.,Vol. 46, pp. 181-191.