Meiofaunal bioturbation effects on the partitioning of sediment-associated cadmium

JOURNAL OF EXPERIMENTAL MARINE BIOLOGY AM3ECOLoGy

ELSEVIER

Mei~fau~a~ ~i~tur~ati~n ef%xts on the partitioning of sediment-associated cadmium Andrew Department

S.Green”, C. Thomas Chandler

ofEwironmental Health Sciences, School ofPublic Health, Uniwr.sity of South Carolina, C~~~ntbia. SC 29208. USA

Received I October 1993: revision received 25 January 1994; accepted 28 February 1994

Abstract Two experiments were conducted using laboratory cultured meiofauna to measure bioturbatio~ effects on cadmium partitioning. The benthic harpacticoid copepod, Amphiascus tenuiremis cf. Mielke, and the foraminiferan, Atnmorzia beccarii cf. Linne’ served as bioturbators. The first experiment utilized single and combined species treatments which were compared to controls having no fauna. All treatments and controls used sediment spiked with Cd at the concentration at which 5 9, mortatity occurs in the copepods (LC, >. Cadmium concentrations were measured in the overlying water, pore wafer, and sediment after 96 h to see if bioturbation altered the partitioning of cadmium. The second experiment employed the same design as the first except that the number of organisms used was tripled to simulate reported field densities. Results from the first experiment revealed significantly higher amounts of Cd in the pore waters of all treatments relative to the controls. There were also between treatment differences showing that copepod-only treatments resulted in the highest concentration of pore water Cd. With the second experiment, it was assumed that this bioturbat~on effect would be even greater due to the increased number of organisms used. However, this was not seen. Treatment pore water Cd concentrations of Experiment 2 were not significantly higher than Iike treatments from Experiment 1. Both experiments demonstrated that meiofaunal bioturbation has a significant effect on the partitioning of Cd in muddy sediments. Kewords:

Bioturbation;

Cadmium; Meiofau~a; Partitioning; Sediment

Macroinvertebrate burrowing, deposit feeding, defecation, and tube-building strongly influence the physicoch~mical dynamics of muddy sediments (Aller, 1978, 1980 and * Corresponding author. 00X-0981/93/$7.00 0 1994 Elsevier Science B.V. All rights reserved SSDl 0022-0981(94)00039-G

1983; Eckman et al., 1981; Aller & Yingst, 198.5;Woodin & Marinelli, 1991). Sediment diffusive permeabililies, depths and shapes of aerobic zones, porewater pH, etc. are all affected by bioturbation (Aller, 1980, 1983). For example, with regard to metals, deposit feeders can alter dissolved and solid phase Fe and Mn profiles through burrow construction, irrigation and enhanced molecular diffusion (Aller, t 978). Iron and manganese oxides are known to be sign~cant sinks for other divaient metals in aerobic sediments including toxic metals such as cadmium and copper (Davies-Coiley et al., 1984; Cowan et al., 1991). Macrofauual enhancement by bioturbation of light- and heavy-metal movement into sedimentary porewaters may increase (or reduce) the bioavailability of toxic metals depending on a whole suite of physicochemical factors, and ultimately influence sediment toxicity. The meiobenthos (i.e. sediment-d~~elling invertebrates > 0.063 and < t .Omm in size) are much more numerous and ubiquitous than macrofauna, yet me~ofaun~ effects through biot~rba~~on on ~thro~ogenic tuxicant p~t~tjoning in sediments and porewater are virtually unknown (see Coulf & Chandler, 1992, for review). Only one study to date, (Aller & Aller, 1992) has examined meiofaunal bioturbative effects on solute movement (Cl- and Br- ) in sediments, Meiofauna occur at such high densities ( > lO’/m’ typically) in the aerobic zone of muddy sediments that they dramatically alter sediment structure through creation of burrows, tubes, mucous nets, and sediment peltetization (Cullen, 1973; Severin et al., 1982; Jensen, 1983; ChandIer & Fleeger, 1984, 1987; Chandler, 1989; Nehring et a!., 1990; Reichelt, 1991). In addition to the above structural changes, meiofauna can modify solid-phase distributions in sediments through the burial and subsequent degradation of sedimented phytoplankton (Webb & Montagna, 1993). Meiofauna can also enhance molecular transport and reaction rates of naturally occurring compounds in the aerobic sediment layer (Aller & Aller, 1992). Enhancement presunlably occurs through ‘~rrleio-bioturbation”, but taxon specific differences in effects on solute mobility at the meiofaunal scale remain unknown. The objectives of our study were to (1) determine if nleiobenthic bioturbation affects partitioning of the toxic melal cadmium among the overlying-water, particulate and porewater phases of muddy sediments, (2) compare taxon-specific influences on Cd partitioning for two meiofaunal taxa having very dissimilar bioturbatjve effects, and (3) determine the relative importance of nleiofaunal density on Cd petitioning. Two experiments were conducted using the free-burrowing harpacticoid copepod, Ar@zia.\-cus teruGremiscf. Mielke, and t.he sediment-pelletizing foraminiferan Ammonia bewcrrii cf. Linne’. The copepod Amplziaws terzuire~ti~is a predominantly infaunal organism which generates extensive burrow networks in the upper 1 cm sediment layer. The foraminiferau Anztno~~in beccarii is cosmopolitan and alters sediment structure by pelletizing and mucous-binding fine sediments swept around its body by 100’s of reticulopodia (Chandler, 1989). Our first experiment utilized both taxa in single-species and dualspecies treatments referenced against control sediments (no fauna) to see if their characteristic bioturbation influenced the amount of Cd associated with particulates, porewater and overlying water. The general null hypothesis for this experiment was that for each taxon separately and in combination, there would be no statistical difference between Cd concentrations in the three sediment phases exposed to meiofauna1 bioturbation versus non-bioturbated controls. If rejected, it would suggest that meiofaunaf

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bioturbation signi~~~tiy affects Cd partitioning in whole sediments. In our second experiment, the same design and procedures were employed as in the first except that copepod and foraminiferal densities were tripled to simulate natural field densities. Our second null hypothesis was that Cd concentrations in bioturbated sediments, pore waters and/or overlying waters of triple density treatments would not be statistically higher from co~~sponding Cd concentrations in the first low density experiment. If rejected, this would suggest there is a positive relationship between meiofaunal density and the bioturbation effect on Cd partitioning among the particulate- pore water- and/or overlying water phases. 2. Materials and methods

Oxidized surface sediments (O-3 cm) were collected from Bread and Butter Creek, North Inlet, SC (33” 20’ N, 79” 10’ W), sieved through a 63-pm sieve and cleansed] condensed, as described by Chandler (1986). For each expe~ment, 200 g of autoclaved stock sediments were blended with 250 ml aerated artificial seawater (30 ppt). The resuhing mixture was then mixed with 4 1 of aerated, artificial seawater and alIowed to re-sediment for 24 h at 4 “C. The supernatant was then removed via aspiration leaving z 600 ml ofwet sediments. The sediment organic content was 3.8j’d (by C:H:N analysis) and solids content (dry weight) was 13%. 2.2. Test ~rg~nisrn~ Copepods and foraminifera used in both experiments were obtained from ffowthrough sediment microcosm cultures of monospecific stocks of Amphiaseus tenuiremis and ~rnrn~~z~abeccarii (~h~dler, 1986). Cultures were fed 2 times per week with a mixed diet of the unicellular algae, I~oc~~~i~ gaibana cf. Haines, Dunalie~la tertjolect~ cf. Butcher, and Fh~e~da~~~lurntri~ornutl~~lcf. Bohlin. One day before initiation of each experiment, test organisms were harvested from 2 1culture dishes by aspirating 50- 100 ml of the upper sediment layer from each culture dish. Sediments and organisms were then gently sieved through a 0.125-mm sieve. Material retained on the screen was washed into a Petri dish, and adult copepods and foraminifera were randomly sorted under a dissection stereo microscope. Test individuals were counted and pipetted into 25 mm Petri dishes and held for 12-18 h in darkness in an incubator at 21 “C until the initiation of each experiment. 2.3. Spiking procedure A 1000 ppm A.CS. certified Cd standard (Fisher Scientific Co.) was used as the stock solution for both experiments. Based on an earlier toxicity study using Amphiascus tenujremjs (Green et al., 1993), the sediment was spiked with Cd to a concentrative (LC,) of 15 mg Cd/kg sediment for both experiments. The total solids content of the sediment was used to calcuiate the requisite volume of Cd stock solution needed to

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achieve a nominal concentration of 15 mg Cd/kg sediment. The spiked sediment slurry was then homogenized on a magnetic stirrer for 1 h at which time the sediment was ready for use. The sediment preparation and spiking procedures were identical for both experiments. Our rationale for the use of the LC, (concentration at which .5p/b mortality is estimated to occur) was that an LCr,, or lOY, mortality in control organisms is normally the acceptable limit in control organisms (APHA, 1985; ASTM, 1993). Therefore, it was decided that the LC, would be a concentration that would both be safe for the overwhelming majority of the organisms and easily measured by atomic absorption spectrophotomctry. Another possible choice would have been the “No Observable Effects Concentration” or NOEC which represents the highest concentration that does not result in significant mortality relative to the controls. Although the NOEC (28.56 mg Cd/kg sediment) for Anzphi~scus ferzuirenzisdid not result in significant mortality relative to controls, it did result in 18”, mortality which for the purposes of this experiment was deemed too high (Green et al., 1993). We assumed that the foraminifera would exhibit similar, non-toxic responses since Cd toxicity for A1~w2or2iu beccari was unknown. This assumption was supported by the occurrence of extensive bioturbation in foraminiferan Cd treatments. 2.4. Experimental desigtl The first experiment implemented four treatments with three replicates per treatment. The four treatments were as follows: (A) no fauna (control); (B) 50 foraminifera; (C) 25 foraminifera and 25 copepods; (D) 50 copepods. Test chambers (sixteen 100-ml beakers) were filled with 60 ml of artificial seawater (30 ppt) using a Repipet precision dispenser. Twenty ml of the prepared test sediments were then slowly extruded over the bottom of each chamber with a Brinkmann pipettor fitted with a 50 ml glass serological pipette. This created a 1.2 cm layer of sediment. The chambers were then placed in humidified trays and incubated static at 21 “C in the dark for 48 h. This was to allow a sufficient pre-trial time period for Cd to diffuse:associate with the sediment, pore water, and overlying water compartments. After incubation, three chambers were randomly chosen for each treatment and each received the proper number and type of organisms as required for a particular treatment. Organisms were recounted as they were transferred from holding dishes to test chambers. One extra test chamber was used as a control in which physical water quality parameters were measured. Dissolved oxygen, temperature, salinity, and pH were measured at 0 and 96 h. After 96 h, test chambers were removed from the incubator and samples were taken for Cd concentration analyses. The second experiment was identical to experiment 1 except that the number of test organisms was tripled, i.e.: (A) no fauna (control); (B) 150 forams; (C) 75 forams and 75 copepods; (D) 150 copepods. 2.5. Artalr~iealchemistry* Samples were collected from all test chambers for both experiments for Cd assay of overlying water, pore water, and sediment particulates. Overlying water was obtained

by taking 5-ml samples midway in each chamber’s water column with a S-ml Finnpipette@. The pH of each sample was then reduced below 2 with concentrated nitric acid (HNO,, AR select). The remaining overlying water was aspirated off along with the top 2 mm sediment layer. This left only a I cm deep layer of sediment. With a Finnpipette@, two 1.3 ml aliquots of whole sediments were taken from between the depths of OS1.0 cm. To extract pore water from sediments, the 1.3 ml sediment aliquots were centrifuged in 1.5 ml micro-centrifuge vials at 14000 x g for 10 min (Carignan et at,, 1985; Schufts et al., 1992). The sediment-Fred sup~rnatant pore water was decanted off and its pH reduced below 2 using HNO,. In 25 ml Erlenmeyer flasks, 0.25 g of centrifuged sediment was mixed with 10 ml of HNO,. The remaining centrifuged sediment pellets were then transferred and digested in a microwave digestor (CEM MDS-81) for 35 min. The resulting concentrated solution was diluted up to 100 ml in deionized water before analysis. Total Cd ~~~cen~ation of all samples was measured by flame atomic absorption spe~tro~hotomet~ (~erkin-Elmer 5100 PC). Sample blanks that matched the salinity and pH of definitive samples were run with all analyses. The detection limit for samples was calculated from blanks to be 0.005 mg Cd/l. Any concentrations below the detection limit were defined as half the detection limit or 0.0025 mg Cd/l for the purpose of statistical analysis. Only measured concentrations were used in statistical analyses and in discussion af results.

Analysis of data included one-factor ANOVA to test significance of treatment effects (c( = 0.05). Tukey’s Studentized Range test was used to reveal significant differences in all pairwise comparisons within experiments. A two-tailed r-test was used to compare controls between experiments. One-tailed t-tests were used to compare like treatments between experiments due to the expected directionality of the comparisons. Aft statistical analyses were conducted using p~~~grarnsissued by the SAS Institute (1985).

3. Results In Experiment 1, there was a dramatic difference in the disturbance/bioturbation of sediments between controls and treatments. As expected, sediments in controls appeared no different at completion of the experiment than at initiation. Treatments with foraminifera alone had an extensively bioturbated layer of *O-3 mm deep with some trails extending to the 5-mm depth. Treatments with only copepods had an extensively bioturbated layer x O-5 mm in depth and visible burrows extending to the bottom of the test chambers (1.2 cm). The following values represent the average of physicochemical measurements ( & 1 SD) taken in the water quality control chambers at 0 and 96 h: temperature = 19.6 _c l.O”C; salinity = 30.3 + 0.2 ppt; dissolved oxygen = 6.8 +_ 0.9 mg/l; and pH = 7.3 + O.I. Analysis of variance revealed no significant differences in sediment-associated Cd concentrations among treatments and controls (F = 0.05, p = 0.98, n = I I). However, Cd concentrations were significantly higher (p
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ofmeiofaunally bioturbated treatments than in controis (Fig. 1). Pore water and overlying water Cd concentrations in the presence of copepods-onIy versus foraminiferan-only were si~~~&ant~y higher under copepod influences (Fig. I). Mean Cd concentrations in the overlying waters of foraminiferan-only treatments were aiso significantly lower than in the treatment where both foraminifera and copepods were present. Pore water concentrations of the copepod:foraminiferal treatment were intermediate and not significantly different from copepods alone nor forams alone. In Experiment 2 where the densities of meiofauna were tripled, there was correspondingly higher meiofaunal bioturbat~on in the faunai chambers than in the meiofauna-free controi chambers. Relative to Experiment 1, the bioturbati~n generated by the larger number of organisms was visibly intensified in the O-5 mm sediment zone. The following values represent the average of physicochemical measurements ( ? 1 SD) taken in water quality control chambers at 0 and 96 h: temperature = 20.6 f 0.42 “C; salinity = 29.0 + 0.0 ppt; dissolved oxygen = 5.75 k 0.4 mg/l; and pH = 7.6 2 0.1. As seen in Experiment 1, sediment-associated Cd concentrations in treatment and control sediments of the triple density Experiment 2 showed no significant differences (F = 1.46, p = 0.30, it = 11). Cadmium concentrations in Experiment 2 were aiso significantly higher in pore waters and overlying waters of meiofaunal treatments versus no-meiofauna controls (P tO.001, Fig. 2). For Cd concentrations in overlying water, only the copepod-alone treatment was significantly higher than controls. There were aiso no significant taxon-specific differences in overlying water Cd concentrations. Pore water Cd concentrations exhibited much stronger between treatment differences. Similar to Experiment 1, pore water Cd concentrations were significantly higher in the copepod-only treatment than in the forami~iferan treatment. The combined copepod:foraminiferan treatment again was intermediate and not significantly different from either singfe-taxon treatment. To test whether a higher density of copepods and/or foraminifera mobilized more Cd into pore waters, mean pore water concentrations from Experiment 2 were compared with those in Experiment 1 (i.e. low density treatments). waters

0

NF 0

FOR 50

FOR % COP

25

25

COP

0

50

Fig. I. Cadmium concentrations in sediment (SER). pore water (PW), and overfying water (OW) phases versus treatments (NF = no fauna, FOR = fo~aminiferans, FOR gL COP = foraminifcrans and copepods, COP = copepods). Each treatment represents three replicates with 50 organisms per replicate. Error bars show one standard deviation of the mean. Letters represent Tukey’s Studentized Range Test where means sharing any letter are not significantly different.

Pare water Cd concentrations were not significantly higher in like treatments of the triple density experiment versus like treatments in the single density experiment. Thus, elevating densities of either taxon separately or combined had no additive effect on the amount of Cd that pa~~t~o~ed into prrre waters. Sediment Cd concentrations were not sigmficantfy different between experiments for 41 treatments and controls. S~gn~~~ant~y higher Cd concentrations were seen in overlying waters of all Experiment 2 treatments relative to Experiment 1. However, with the large number of comparisons being made (i.e. 12) this statisticai result may be due to Type I error. ~onfe~on~s correction showed that only one of the three overlying water treatments in Experiment 2 was si~~fi~antiy higher than its counterpart in Experiment I (p = 0.02).

~eiobenthj~ bioturb~ti~n signifiea~~y alters the pa~it~oning of Cd from the particulate sediment phase to the pore water phase of Cd-spiked sediments. The strongest taxon-specific influence on Cd movement was an increase in pore water concentrations under capepod bioturbation. Foraminiferan bioturbation also significantly enhanced Cd pore water concentrations but to a lesser degree (35 + II”/; lower, IZ = 6) than copepod bioturba~o~. As might be expected For a 50:50”1, species mixture, combined co~epod~~or~~~~~eran tr~atrne~~tswere simply additive Zn their effects on pore water Cd concentrations and yielded almost perfect intermediate effects in both experiments, Surprisingly, elevated (3 x ) densities of each taxon separately and in combination had no significant effects on corresponding pore water Cd concentrations despite correspondingly greater copepod burrowing and foram pelletizing activities. No foraminiferal inhibition of copepod effects were seen despite pubfished evidence of strong foramin~fera~:~opepod trophic amensahsm between A~~~~~~~~~ beecar and another copepod species, _~~~?~~~~c~~~~ ~~~~~~~~~ which is very similar io ~~~~~~us~u~~~~~~~~~~~~ both phy~oge~etj~ally and ecologically (Chandler, 1989).

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sediment mechanical breakdown and release of associated toxins is likely by the burrowing and feeding activities of benthic copepods and foraminifera. Amphiascus tenuiremis is a disruptive sediment burrower and sediment ingestor that can rapidly disaggregate muddy sediment conglomerates and effectively increase sediment particulate surface area (Chandler, unpubl.). Ammonia beccarii is a sediment sweeping foraminiferan that gathers sediment particles into hollow spheres around its body, removes almost all of the associated food particles (bacteria and diatoms), and then ejects itself from within the spheres. The net effect is a much coarser collection of 0.1-0.5 mm diameter sediment:mucous conglomerates (Chandler, 1989) having correspondingly lower surface area than copepod bioturbated sediments. Given these contrasting sedimentary effects, it is not surprising that copepods mobilized more Cd into sediment pore water than foraminifera. Our second experiment found no statistical difference between Cd concentrations in sediment pore waters where meiofaunal densities were tripled to observed field densities (78 organisms/treatment/lo cm’) versus those of Experiment 1 where densities were only 26 organisms/treatment/lo cm3. If one assumes that the biological activity of Amphiascus tenuiremis and Ammonia beccarii increased the rate of Cd transport into pore water as evidenced here, then one may also assume that there is a densitydependent upper bound for the rate of Cd mobilization that can be generated by these taxa. A density of only 2.6 copepods and/or foraminifera per cm3 may be at or near the maximal effect on Cd transport rate for these sediments. Thus the addition of triple densities generated no significant Cd increase. Perhaps if only 1.3 organisms per cm3 were used as a treatment level, a smaller pore water Cd concentration would have been seen. Such low densities for these and other major taxonomic groups are unrealistic for muddy sediments in the field however (Heip et al., 1982; Hicks & Coull, 1983). Our experimental design did not attempt to study the mechanics/dynamics behind Cd partitioning beyond a gross species-specific bioturbation level, but several possible explanations for our findings have been reported for macroinvertebrate:sediment systems. For example, macrofaunal bioturbation increases solute transport and diffusion, and alters in situ metal, oxygen and pH profiles (Aller, 1978, 1980, 1983; Eckman et al., 198 1; Aller & Yingst, 1985). Alterations of this type to the sediment matrix can directly or indirectly change the mobility of divalent metals such as Cd (Davies-Colley et al., 1984). Several studies have also documented the ability of meiofauna to extensively alter sediment structure/erodability through burrowing, pelletization, mucous deposition, etc. (Cullen, 1973; Chandler & Fleeger, 1984, 1987; Chandler, 1989; Reichelt, 1991). These kinds of bioturbation should change the diffusive properties of sediments in taxon-specific ways. Aller & Aller (1992) found that mixed assemblages of meiofauna (Z 30-100 individuals per 10 cm3) could increase Cl - and Br solute transport rates in the aerobic zone by a factor of 1.7-2.3 compared with meiofauna-free controls. In our study, Cd partitioning into pore water ranged between 2 and 3.6 x the no-meiofauna control concentrations of the first experiment and 1.7 and 2.3 x the no-fauna control concentrations of the second experiment. Given that a different solute was studied using only two taxa, it is remarkable that the trends for increasing Cd concentrations in our study were so similar to those reported for Cl- and Br- in Aller & Aller (1992).

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Meiofaunal bioturbation also increased porosity, created fluid motion and enhanced three-dimensional solute diffusion. Meiofa~~a~~~enhanced solute trar~sport no doubt increases metabolic rat~s~~rodu~tion of aerobic decomposers which may in turn mobilize sediment-associated toxins such as Cd. Several studies have shown that meiofauna! grazing also enhances bacteria1 growth rates (e.g. Fenchel & Harrison, 1975; Alongi, 1985). ~a~robenthjc bi~turbatjon commonly enhances solute transport 2-10 x in sedimeuts and has a much greater infhtence than the lne~ofanna on geochemical processes in anaerobic sediment horizons - at least in intertidaf to continental shelf sediments (Alier & Aller, 1992) However, the potential for broad-scale meiofaunal enhancement of toxin mobility in the aerobic zone of shallow to deep sea sediments should not be disregarded simply because of their smailer effect on solute mobility at local scales. ~~~obent~l~~soccur at densities of at feast 30- 100 ~nd~~~idua~s per IO cm3 in almost every fine sediment habitat spuddedto date (He@ et at., 1982; Hicks & Couth, 1983); and these densities occur un~~o~l~y across broad spatial scafes with scattered random patches of Z- 10 x these densities (Couil, 1988). This characteristicaliy broad, ubiquitous dispersion may make the meiobenthos as or more important in mobilization of sedimentassociated toxins than the larger macrobenthos which typically exhibit patchy, more locahzed distributions in sediments. This may be especially true in deep sea sediments where ~~a~robenthos rival meiobenthos in size but not number, and rna&ro~~~e~obenth~~ biomass ratios are -2 I:1 (Thief, 1975; Tietjen, ‘1992;Giere, 1993). ~~e~ofauna~densities (unlike the macrobellthos) remain at high levels in the deep-sea at l~~-~O~~ individuals per 10 cm’ excluding the foraminifera. Furthermore, w 80”; of all metabolic processes in deep-sea sediments may be attributable to meiabenthos (Shirayama & Worikoshi, 1989). As is true with most iaitiaI studies, more questions were raised than solved in this study of meiofaunal effects on Cd partitioning. Two iaboratory-cultured species having very dissi~~lilarbioturh~tive effects were used to demonstrate that meiofa~na1 bioturbation can influence Cd petitioning in marine sediments. In doing so, the study was also able to demonstrate taxon-specific influences on Cd partitioning for two meiofaunaf taxa. However, the use of cultured organisms under arti~~ja~~~controlted conditions raises the question ofhow these findings relate to natural field settings. The next logical step would be to compare field-coliected meiofauna and their bioturbation effects with the ultimate goal of a full meiofaunal community assessment of Cd (or other toxins) partitioning effects. In addition, this study used a spiking method to create Ed-contaminated sediments. Although spiking sediments is a common method in toxicity studies {Swartz et at., 1985, Mearns et at., 1986)_the question of ~u~~~b~urnwithin the sediment matrix arises. The present study used a 48 h incubation pre-trial period to allow time for Cd to equilibrate with the sediment, pore water, and overlying water compartments. This was considered a conservative time period as complete Cd-sediment-pore water equilibration should occur within 24 h in marine sediments (Oakley et al., 1981). It was therefore assumed that at the test initiative, Cd partitioning had reached equ~~ibr~~m.However. spiking approaches attempt to simulate field cont~n~inated sediments and as such+ are s~s~ept~h~eto errors inherent whenever ones tries to make simple what is normally a very complex biogeochemical system. Thus future

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research in this area should also include studies of meiofaunal bioturbation on toxicant partitioning in field-collected, contaminated sediments. Clearly a better understanding of meiofaunal-induced changes in porosity, pH, O2 and other variables in the sediment microhabitat would improve our understanding of the small-scale mechanics influencing Cd mobility, bioavailability and especially toxicity. In much of contaminated sediment toxicology to date, interactions between the fauna1 and toxic components of sediments are almost always characterized as chemistry-impacting biology which leads to toxicity. However, the reverse view also needs to be considered. For example, if meiobenthos can elevate toxin concentrations in pore waters of bedded field sediments, then bioavailability and subsequent toxicity to less resistant fauna/taxa should also increase. Such biotic influences on toxin bioavailabilities may be particularly important to accurate development of sediment quality criteria since it is well known that the concentration:mortality curve for metals and lipophilic organics correlates better with pore water concentrations than bulk sediment concentration (Swartz et al., 1985; Kemp & Swartz, 1986; DiToro et al., 1991). In addition, it has been shown that Cd partitioning is controlled primarily by acid volatile sulfide (AVS) in anoxic sediments (DiToro et al., 1990, 1992; Ankley et al., 1991). AVS was not measured in this study as all sediments used were oxic based on DO levels, depth of sediment used, and the fact that Amphiasczrs tenuiremis is only found in the oxic zone in surface sediments. However, the effect of bioturbation on Cd partitioning in anoxic sediments where AVS is a primary factor can be speculated. Cadmium reacts with solid phase AVS when present to form cadmium sulfide (CdS) precipitate which is not bioavailable (DiToro et al., 1990, 1992; Ankley et al., 1991). Bioturbation can not directly alter the chemical state of CdS, but bioturbation can alter the depth and shape of aerobic zones, and the pH of porewaters (Aller, 1980, 1983). An increase in oxidation or a lower pH can result in Cd being released from its nonbioavailable state. Predictions of pore water chemical concentrations for the purpose of sediment quality criteria development are usually based on partitioning coefficients derived from sediment and solute chemical properties solely. One copepod bioturbation treatment in our study had 3.6 x the pore water Cd concentrations observed in the faunal-free control sediments due apparently to biological activity and not chemically-based sediment:pore water partitioning. Such biologically driven phenomena no doubt contribute substantially to the discrepancies often seen between predicted and measured chemical concentrations (and toxicities) in sediment pore waters. Meiofaunal bioturbative effects on sediment toxicant mobility is a factor that has been overlooked or ignored by ecologists, chemists, and toxicologists alike; but it is certainly deserving of further study with additional taxa and a broader range of sediment toxins. Acknowledgement We thank W.W. Piegorsch for help with the statistical analyses. This research was supported by the U.S. Environmental Protection Agency, Assistance No. R8170000-01 (B.C. Coull & G.T. Chandler, principal investigators) and the U.S. National Oceanic and Atmospheric Administration Coastal Ocean Program.

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