Toxicity of sediment-associated tributyltin to infaunal invertebrates: Species comparison and the role of organic carbon

PII:

Marine Environmental Research, Vol. 43, No. 3, pp. 219-241, 1991 Published by Elsevier Science Ltd Printed in Great Britain 0141-1136/97 %15.00+0.00 SO141-1136(96)00090-6

ELSEVIER

Toxicity of Sediment-Associated Tributyltin to Infaunal Invertebrates: Species Comparison and the Role of Organic Carbon James P. Meador, Cheryl A. Krone, D. Wayne Dyer & Usha Varanasi Environmental Conservation Division, Northwest Fisheries Science Center, National Marine Fisheries Service, National Oceanic and Atmospheric Administration, 2725 Montlake Blvd East, Seattle, Washington, USA (Received 25 November 1995; revised version received 2 April 1996; accepted 11 April 1996)

ABSTRACT Experiments with three species of infaunal invertebrates (a polychaete, Armandia brevis and amphipodr, Rhepoxynius abronius and Eohaustorius washingtouianus) with dtferent modes of feeding gave a wide range in toxic response to sediment associated tributyltin (TBT), while exhibiting consistent lethal tissue residues. These studies showed that bioaccumulation and toxicity of sediment-associated TBT were strongly controlled by the organic carbon content of the sediment, which we concluded was primarily due to its influence on interstitial water (IW) concentrations of TBT. Major dtferences in the response to sediment-associated TBT werefound between species, which was attributed to dtxerences in their rates of uptake and elimination of this compound. Predictions for bioaccumulation and toxicity for each species based on these toxicokinetic rates were matched closely by observed values. Based on comparisons of water-only and I W exposures (when water and sediment concentrations of TBT were in equilibrium) and predictions made with toxicokinetic rates, the major route of uptake for each of the species tested appeared to be from dissolved TBT. We determined the mean (sd) organic-carbon normalized sediment-water partition coeficient (K,,) to be approximately 25 100 (5500) for TOC values ranging from 0.3 to 1.O%, which was fivetimes higher than the reported K,,. Additionally, we determined the dissolved organic carbon-water partition coeficient (K& to be 1652, which was three fold lower than the K,,. The results also showed that the K,, could be influenced by infaunal organisms, presumably by reducing IW concentrations below predicted values, which raises questions about the environmental relevance of sediment bioassays using these organisms and the expected IW concentrations present in$eld sediments. Published by Elsevier Science Ltd

INTRODUCTION Before its use was restricted, tributyltin (TBT) had been widely used as an antifoulant on ships, nets, piers,’ buoys, and other nautical devices to retard the growth of fouling 219

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organisms. In 1988, the United States Congress passed the Organotin Antifouling Paint Control Act (OAPCA; US Congress, 1988) to limit the use of TBT, making it the only pesticide to be specifically regulated by the US Congress. TBT concentrations in water dropped dramatically in the years subsequent to OAPCA (Valkirs et al., 1991); however, sediment concentrations have shown only modest declines (Valkirs et al., 1991). Similarly, Dowson et al. (1993) found that TBT concentrations in water have significantly declined since 1991 in estuaries in the UK, whereas sediment concentrations have decreased to a lesser extent. Recent surveys show TBT concentrations in sediment over time are relatively constant (Krone et al., 1996; Fent, 1996); hence sediments may have become a source, whereas they were once a sink for TBT. Several benthic areas in Puget Sound have been examined and environmental levels of TBT were found to be in the ppb (ng gg’) to low ppm (ug gg’) range (Krone et al., 1989a,h). Similarly, Langston & Burt (1991) and Dowson et al. (1993) demonstrated that TBT in sediment can easily approach hundreds of ng gg’ in British estuaries. Another study of 34 US coastal sites, which was part of the National Oceanic and Atmospheric Administration’s Mussel Watch Project, found TBT present at 25 of these sites at concentrations up to approximately 200 ng g-i (Wade et al., 1990). In a larger study that examined 99 sediment samples from 41 embayments on the US coastline, 60% of butyltin-containing sediments had concentrations greater than 200 ng gg’ and in a few cases sediments reached 1600 ng ggi (dry weight; Krone et al., 1996). Other studies around the world have found sediment concentrations of TBT in the 3 to 10 ug gg’ (dry weight) range, especially when close to marinas and shipyards (Krone et al., 19896; Espourteille et al., 1993; Fent, 1996). Acute toxicological data suggest that TBT is extremely toxic to marine invertebrates (Maguire, 1987; Cardwell & Meador, 1989; Heard et al., 1989; Fent, 1996) and because of its persistence in sediment, is still of concern. It is known that marine species exhibit a range of responses when exposed to TBT in water (Cardwell & Meador, 1989; Meador, 1996; Fent, 1996); however, their response to TBT-associated sediments, has rarely been studied and detailed knowledge on partitioning behavior in relation to the mechanisms of bioaccumulation is lacking. It is important to determine the portion of a toxicant that is available to an organism to more accurately predict bioaccumulation and the levels that cause adverse effects. In this study we chose to examine the effects of organic carbon on TBT partitioning and how it controls the bioaccumulation of TBT in sediment dwelling organisms. In marine systems, one can assume (within reasonable spatial and temporal limits) that other variables (e.g. pH, alkalinity, and hardness) that may control the bioavailability of TBT are relatively constant. Redox state may also be an important variable in the control of TBT partitioning and worthy of study because of the potential differences in desorption from sediment to overlying water. Because TBT is ionic in aqueous solutions, Mn and Fe oxyhydroxides (and other surfaces) may be important controlling factors due to their complexation capacities. Grain size may control TBT partitioning; however, it is known that organic carbon content is generally inversely correlated to the mean grain-size distribution for a sediment because it is often found as a coating on particles. Even though some or all of the aforementioned parameters may affect TBT bioavailability, we expected that organic carbon would be the dominant controlling factor in most natural situations because of its moderately high K,,. This study was designed to assess and compare the responses of selected benthic infauna to sediment containing tributyltin and to examine the effects of sediment organic

Toxicity of tributyltinto infaunal invertebrates

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carbon on the control of toxicant exposure. We were also interested in determining the major route of TBT uptake and therefore tested species with different modes of feeding for comparison. We tested the amphipods Rhepoxynius abronius (meiofaunal predator) and Eohaustorius washingtonianus (detritivore), and a polychaete, Armandia brevis (nonselective deposit feeder). For two of the species, we varied organic carbon in sediment, while keeping grain size relatively constant, to explore how this variable may influence the toxic response. Our null hypothesis was that total organic carbon in sediment would have no effect on the toxicity response for R. abronius or A. brevis. To evaluate our results and help develop predictions in other situations, we measured the sediment-interstitial water (K,,) and dissolved organic carbon-interstitial water (K& partition coefficients. While interested in the concentrations of TBT in sediment and interstitial water that cause toxicity, we were also interested in trying to predict toxic responses and to test interspecific variability.

MATERIALS

AND METHODS

Rhepoxynius abronius (Phoxacephalidae) and Eohaustorius washingtonianus (Haustoriidae) were collected in August, 1991 from West Beach, Whidbey Island, Washington (a minimally contaminated area). Rhepoxynius abronius were obtained by dredging in about 3 m of water (-4 m MLLW) with an amphipod dredge. Eohaustorius washingtonianus were collected during low tide (- 1 m MLLW) in August, 1992. Because these two species occurred within a few meters of each other, we assumed that they came from almost identical exposure histories and hence their differences in response to TBT would be due to innate physiological traits and not adaptation. All animals were held for less than 10 d before testing. We tested adults, which were generally 3-5 mm in length and randomized over treatments. The mean (sd) final dry weight was 1.3 (0.2) mg for R. abronius and 0.5 (0.2) mg for E. washingtonianus individuals. The polychaete Armandiu brevis was collected in June, 1991 on the intertidal mudflats of Mitchell Bay on San Juan Island, Washington, which had no detectable TBT. Most individuals were adults (some juveniles) with a mean (sd) individual dry weight of 1.4 (0.4) mg. Sediments were manipulated by mixing three components to vary TOC and hold grain size (by weight) relatively constant. The majority of the sediment mixture was composed of sand from West Beach, Washington (sieved at 500 pm, TOC = 0.13%), which was the native substrate for R. abronius and E. washingtonianus. To this we added a fine mud from Polnell Point (northern Puget Sound, TOC = 1. 1%), which is a reference site we use for other studies and is uncontaminated. The third component added was organic matter (TOC = 2.5% dry weight) collected from the headbox at our marine station in Mukilteo, Washington. This material is a fine, brown flocculent material that is presumably made up of zooplankton and phytoplankton casts and other natural organic material. These three components were added together in different proportions to achieve variable organic carbon content. Because sand was the largest component ( > 80% by weight) in all treatments, the grain-size distribution between treatments did not vary greatly. For the Ew test (see below), we used sediment from Mitchell Bay, San Juan Island, Washington for the substrate, which was not mixed with any other materials. Each of the above sediment mixtures were added to a brown glass jar, spiked with a concentrated solution of TBT in acetone (0.5 l.d g-l sed) and tumbled on a rolling mill for

222 48

J. P. Meador

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h at room temperature. Sediment in the jar was moist, not a slurry. Sediment (250300 g dry weight) were then dispensed to l-liter Pyrex beakers and filtered (5 urn), UV sterilized seawater (overlying water =OW) was gently added. After 3 d, the OW was siphoned off and replaced (day 0). The following tests contained these treatments: R. abronius (Ra), 3 TOC x 4 TBT; A. brevis (Ab), 3 TOC x 4 TBT; E. washingtonianus (Ew), 1 TOC x 7 TBT and were conducted for various lengths of time. Sediment TBT concentrations were as follows: Ra, l-10 ug gg’; Ew, 0.04-4.5 ug gg’; Ab, 0.4-6 ug gg’ and the factor difference between TBT treatments varied between 2.2 and 2.5, depending on the test. The Ab test contained chemistry reps (without worms) which included two reps for each of the three TOC treatments. Sediment samples for analysis were taken before animals were added to the beakers and at the end of the experiment, to determine if the concentration of TBT in sediment was consistent over time. Each test contained an acetone control for each TOC treatment and one native-substrate control (unamended). Because of its volatility, we expected very little acetone to remain after the addition of seawater. Individuals were added to beakers on day 0 and enumerated daily. The number of animals per beaker were as follows: Ra, 40; Ew, 20; and Ab, 15. Individuals were assumed to be feeding (except Ra) during the experiments and in a separate observation, A. brevis was observed to ingest sediment continuously. Beakers were aerated with a pipette bubbler and kept on a water table to maintain constant temperature. The fluorescent lab lights were on continuously to keep organisms buried in the sediment. OW was changed as needed in each experiment to ensure good water quality according to the following schedule: Ew (days 0 and 5); Ra (days 0 and 5); Ab (days 0, 2, 4, 7). In all tests, dead animals were removed daily, soaked in clean water for 1 h, and preserved by freezing. The gut was not purged of sediment; however individuals were on the sediment surface for a day or two before death and may have purged at this time. At the end of a test, live animals were placed in clean seawater for 6 h to purge their gut of sediment. The endpoint for LCsO determination (the water concentration that caused 50% mortality at a specified time interval) was the absence of any motion by an individual. Tissue residue of TBT was determined for animals at the time of death or on the last day and was used to calculate a bioconcentration factor (BCF) or LD5a (the lethal dose or tissue concentration that caused 50% mortality at a given time interval). Uptake clearance (kt) was calculated for the highest treatment concentrations, which corresponded to the shortest time of exposure (generally l-2 d). The elimination rate constants (k2) were taken from Meador (1996). With both the kl and k2 values, we predicted BCF, time to steady-state tissue concentration, and L&a. Dry:wet weight ratios were determined several times over the time period of these experiments by weighing several individuals of a species on blotter paper to 0.1 mg using a Mettler AT261 Delta Range balance. Individuals were dried at 70°C for 24 h, allowed to cool and reweighed. Initial total-lipid content was determined on a subsample of the field population before bioassay testing. Randomly selected animals ( = 0.2 to 0.3 g wet weight) were homogenized in a Tekmar Tissumizer and lipids were extracted according to a modified Bligh and Dyer technique (Herbes & Allen, 1983) designed for small sample sizes. Residues were weighed to 0.01 mg on a Mettler AT261 balance. Dissolved organic carbon (DOC) was sampled with a glass syringe and solutions were filtered with a pre-ashed 0.45 urn filter, stored frozen in a glass vial (cleaned of organics) with a Teflon lid and analysed by the ultra-violet promoted persulfate oxidation technique

Toxicity of tributyltin to infaunal invertebrates

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(Standard Methods No. 505a; APHA, 1985) using a Shimadzu Total Organic Carbon Analyzer (Model 5000). Total organic carbon (TOC) in sediments was measured according to the method of Hedges & Stern (1984) using a Carlo Erba CHN analyzer Model 1106. The certified standard was acetanilide from Sanda Inc. Sediment grain size composition was determined by sieving and the pipette method as described by Buchanan & Kain (1971). Total ammonia was determined with a Hach test kit. Interstitial water (IW) was measured once at the end of the experiment in all test containers and was separated from bulk sediment by two different methods of centrifugation. For the Ra and Ab test, a polycarbonate (PC) bottle (250 ml) was fashioned into a separation device by cutting off the bottom and placing dual polystyrene petri-dish bottoms (Falcon number 1006) with drill holes (2 mm), back to back, inside the PC bottle near the screwcap end. Two layers of Nitex netting (300 urn pore size) were placed on top of the petri bottoms. Sediment was placed inside the container and centrifuged at 4°C with the screwcap end down, at 900 g for 20 min. Four devices were used concurrently. Interstitial water was collected and centrifuged again in clean, unmodified polycarbonate bottles in a Sorvall H-T6000B at 2400 g for 20 min, which was sufficient to remove particles larger than 0.1 pm, and then analysed for TBT. For the E. washingtonianus test, an air-powered centrifuge (CRC 18901) was used. Sediment was added to polypropylene cups with 1 mm holes, which were lined with Nitex netting (300 urn). IW was collected in a Teflon lined catch basin and drained through a Teflon tube to a polycarbonate jar. This solution was centrifuged again at 2400 g for 20 min and frozen until analysis. In each experiment, sediment was centrifuged for IW before the animals were sieved out. This procedure did not seem to affect the animals and in most cases the number assumed alive at the end of the test were found alive after sieving. In only a few cases was there a discrepancy between the number exposed and the number found (minus those found dead on the surface during the experiment). Partitioning of TBT to DOC (Kdoc) A separate study was conducted to determine the partition coefficient between DOC in interstitial water and TBT. This was accomplished by mixing bulk seawater with interstitial water to prepare samples with a range of DOC concentrations, then equilibrating these mixtures with different levels of TBT. These solutions were then passed through C,s Sep-Pak cartridges as described by Landrum et al. (1984). Our treatments consisted of DOC concentrations ranging from 7 to 200 mg C liter-’ with TBT held at 50 ng ml-’ and TBT concentrations ranging from 1 to 100 ng ml-’ with DOC held constant at 54 mg C liter-‘. For each solution, a 50-ml capacity gas/liquid-tight syringe was used to pass 100 ml of the solution through a pre-treated Cis Sep-Pak cartridge. Pretreatment consisted of passing 5 ml of methanol through the cartridge followed by 5 ml of water that was free of organic carbon. The effluent was collected in a polycarbonate bottle and stored at 5°C for TBT analysis. The flow rate was kept to x 10 ml mini with a Sage model 234 syringe pump. The same apparatus was used to pass 100 ml methylene chloride (CH2C12) through selected cartridges to extract the free TBT retained by the cartridge, which was generally not effective. The CHzC12 extracts were collected in amber glass bottles and stored at 5°C for TBT analysis. The amount of TBT retained by the cartridge was determined by difference (total minus that which passed through the cartridge).

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To determine the amount of organic material retained by a cartridge after elution with CH2C12, one cartridge was weighed before treatment. 50 ml of a solution containing 200 mg liter-’ DOC was passed through this cartridge, followed by 50 ml CH&lz. The cartridge was then dried at 110°C to constant weight, resulting in a mass difference of 1.3 mg. Assuming that all retained material was dissolved organic matter, and this material was no more than 40% carbon, we can conclude that no more than 0.5 mg (5%) of the DOC was retained on the cartridge after elution with CH2Cl2. Our observation for DOC retention agrees favorably with that (11 f 3%) found by Ozretich et al. (1995) for variable DOC in estuarine interstitial waters. Because we considered this retention minor, our results did not include this in the calculations. Preparation of the samples for TBT analysis by GC/FPD was performed by the standard TBT protocol discussed below. No further extraction was performed on the CH&lz solutions before concentration to 10 ml, and no tropolone was added to these solutions. The water samples were processed according to the protocol, except each extraction into CH2C12 was performed using 20 ml of CH&X&ropolone for each sample. All storage vessels used for this study (unless otherwise noted) were made of polycarbonate and were rinsed with methanol before use. Chemical analyses

Tributyltin chloride is an ionizable, organometallic compound with a log K,, reported to be 3.74 (Laughlin et al., 1986) and was obtained from Aldrich Co. (96% pure). Butyltins were determined in the tissue and sediment according to Krone et al. (1989a, 1996). Water samples were placed in a separatory funnel and extracted twice with 50 ml of CH2C12 containing 0.1% tropolone. Approximately 10 g of sediment was extracted by tumbling with 50 g sodium sulfate and 100 ml CH2C12 (containing 0.1% tropolone) in amber glass bottles for 16 h. The CH2C12 was decanted, additional CH$&/tropolone was added, and the bottles were rolled for an additional 6 h. These extracts were then combined. Tissue, CH2C12 containing 0.1% tropolone, and sodium sulfate were macerated/extracted using a Tekmar Tissumizer. The solutions were decanted and the extracts combined. The methylene chloride extracts from the water, tissue, and sediment samples were concentrated and exchanged to hexane. The extracted alkyltins were converted to their n-hexyl derivatives through the Grignard reaction and the organic phase subjected to silica/alumina mixed-bed cleanup using pentane as eluting solvent. The alkyltins were determined by GC with flame photometric detection (GC/FPD). GC/FPD was carried out on a Varian model 3700 gas chromatograph equipped with a fused silica capillary column (bonded SE-54, ca. 30 m x 0.25 mm i.d.), Varian flame photometric detector (fitted with a 610 nm bandpass filter for tin-selective response), a Varian 8040 autosampler and a Varian Star integrator. Tripentylmonobutyltin at 1 ng ul-’ (as tin) was employed as GC internal standard. Calibration was done by the internal standard method using peak heights and a calibration curve with five concentrations. Quality control procedures included the use of recovery standards, method blanks, spiked blanks, calibration standards and reference materials (RM). We used three reference materials; muscle of seabass (Okamato, 1991) for tissue and two sediments prepared in our laboratory (Krahn et al., 1988). Our analysis of the seabass determined the mean (sd) concentration to be 1.3 (0.5) ug g-r (n=4; certified at 1.3 (0.1) ug gg’). Our two RM

Toxicity of tributyltin to infaunal invertebrates

225

sediments, SQl (spiked) and DUW (naturally contaminated), have historical mean (sd) TBT concentrations of 111 (19) ng gg’ (n = 6) and 29 17 (863) ng gg ’ (n = 9; dry weight), respectively. In this study our mean (sd) analysis for SQl was 127 (9) ng g-’ and for DUW was 3038 (299) ng g-‘. TBT concentrations were adjusted for the recovery of tripentyltin and are reported as ng TBT ml-’ or g-’ (1 ug g-l = 3.46 nmol g-l and 1 nmol gg’ = 290 ng g-l). Recoveries for water, sediment, and tissue measurements were generally in the range 70 to 110%. All sediment and tissue concentrations expressed as dry weight, were noted. All reported TBT concentrations were measured, none were nominal. The limit of detection for tissue was 0.75 pg g-’ (dry weight), 0.01 ng gg’ (dry weight) for sediment, and for water was 0.3 ng ml-‘. Statistical analyses The LCsOs and LDsos were determined by the GLMLCp method (Kerr & Meador, 1996). In a few cases, the data did not fit the binomial model for the GLMLCp method and we used the moving average or probit technique with software supplied by C. Stephan (Stephan, 1977) to calculate these statistics. Some toxicity values (IW and sediment L&s) were computed with data from each treatment in an experiment and some were determined with all data from an experiment (IW and sed,,-LCs,,, and LDSo values). Control mortality in three of the experiments exceeded the 10% level (ASTM, 1990). For the experiments where control mortality exceeded 15% (Ew day 41, one of two replicates and Ab, three of six replicates) we added the number of mortalities that occurred in the acetone control to the treatment survivors to offset the mortality due to factors unrelated to TBT toxicity. A comparison of the LCsos before and after application of this correction showed only minor changes and they were not significantly different. Standard deviations (sd) are reported to show the range in the data and the standard error of the mean (sem) is reported when comparisons of means are intended. Analysis of covariance (ANCOVA) was used to test for differences between slopes for the relationship between log,sK,, and TOC when organisms were present or absent for most of the test. The model was: log,0 K,, = PO+ BIX +

1321 + /%1x

(where X = TOC (covariate); I = indicator variable (0 or tively); IX = interaction term between TOC and presence fitted by the regression. If the coefficient of the interaction concluded that the slopes between categories (present or ent. All statistical assumptions about this test were met. The following equations were used in this study: 1. 2. 3. 4.

1 for absent or present, respecof organisms; B were coefficients term (/?3) was significant, it was absent) were significantly differ-

BAF (bioaccumulation factor) = [tissue] / [sed] BAF,,, (lipid and sediment organic-carbon normalized BAF) = ([tissue]/fiip)/[sed.,l BCF (bioconcentration factor) = [tissue]/ [water] or can be predicted with (k, / k2) f,, (fraction of steady state) = 1 - eMk2*t

5. k, (uptake clearance constant) = ,J:iei 6. kz (elimination rate constant)

t (units are ml g-l t-l)

-(tissue], = -((In Itlssuel _ _. )/t) (units are time-‘) I-

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7. Kdoc (DOC to IW partition coefficient) = [IW,,,,,,]/([IW,,,] * f& 8. K,, (sedoc to IW partition coefficient) = ([sediment]/fJ[IW,,,,11 9. K, (sediment to IW partition coefficient) = [sediment]/[IW] k2/k, * LD50 10. LC50 predicted for time t = , _ e_ k2It 11. sed,, (organic-carbon normalized sediment concentration) = [sediment]/f,, 12. TSS (time for tissue concentration to reach 50% (TS&) and 95% (TSS& of steady state) = 0.693/k* and 2.99/k*, respectively. 13. [water&, (equilibrium concentration of unbound TBT in water) = [water,,,,,]/ (1

+

fdoc

* Kdoch

where, flip is the dry weight fraction of tissue that is lipid, f,, is the dry weight fraction of sediment that is organic carbon, and f&,c is the fraction of water that is dissolved organic carbon (g ml-‘). RESULTS Table 1 lists the TOC and DOC values observed in these experiments and the corresponding partition coefficients. Manipulation of the organic carbon content with little variation in grain size was successful with a range of over seven fold in organic carbon content. The ratio of added TOC to naturally occurring TOC in sediment is listed for each treatment in Table 1. TOC was measured at the beginning and end of a test and found to be generally consistent over time (Table 1). It appears that DOC in IW was positively correlated with TOC in sediment; however some treatments showed large variation. As seen in Table 1, both the mean partition coefficient (KJ and the organic carbon normalized partition coefficient (K,,) determined for sediment and interstitial water varied by a factor of 34, with individual values ranging up to an order of magnitude. There was no correlation between either sediment (r* = 0.0004) or IW (r* = 0.003) concentrations of TBT and K,,. There was a strong negative correlation between TOC and loglOK,, (Fig. l), especially when organisms were present. The ANCOVA determined that the slopes between groups (present and absent) were significantly different (p = 0.014), indicating that the relationship between loglOK,, and TOC varied with the presence of organisms. In those beakers without organisms for most of the test, the trend between TOC and loglOK,, was substantially reduced and when one value from the Ra test (K,, = 4.72, TOC = 0.12%) was eliminated, the coefficient of determination (r*) was reduced from 0.44 to 0.16 and the slope became closer to zero (-0.37 to -0.18). The mean (sd) log,,K,, for beakers with organisms present for the majority of the experiment was 4.73 (0.27) and 4.40 (0.09) for those beakers where organisms were absent during most of the experiment. The mean (sd) back-transformed values are 53 300 (45 500) and 25 100 (5500), respectively. This K,, value for treatments where organisms were absent does not include the one point from the Ra test mentioned above. Elimination of this point reduced the standard deviation for the K,, by 50% and caused the slopes between the present and absent group to become more dissimilar (p = 0.004). Because the sediment was generally undisturbed for most of the test, we feel that the mean K,, measured in those beakers would more accurately represent the true K,, for TOC values between 0.3 and 1.O%; hence this became our standard value for some of the calculations. The partition coefficient determined by the regression between free TBT and that associated with DOC in the interstitial water was 1652 (3.22 in loglo units) with a standard

143 (85) 153 (107)

0.55 (0.12) 0.82 (0.11)

0.87 (0.09)

3

(!f)

0.57 (0.13)

0.31 (0.06)

0.30 (0.05)

1

*

137 (93)

2

0.47 (0.13)

0.62 (0.2)

0.69 (0.06)

0.59 (0.05)

5.0

(Z::)

(E) 203 (20)

203 (51)

119 (10)

158 (33)

318 (35)

209 (51)

0.29 (0.05)

0.28 (0.04)

(C)

119 (53)

KP

0.08 (0.01)

ow (mg liter-‘)

DOC

0.12 (0.03)

IW (mg liter-t)

Day end (%)

tests

TOC Day 0 (%)

1

1

TOC Treat

TABLE 1 parameters for tributyltin

4.37 (0.07)

4.53 (0.1)

84

86

84

4.59 (0.07)

90

4.67 (0.06)

92

94

4.86 (0.09)

4.44 (0.07)

97

Sand (%)

5.02 (0.18)

togto&

3.5

0.9

0.05

2.9

0.9

0.08

Ratio TOC

4

4

4

7

4

4

4

n

Mean (sd) total organic carbon (TOC; percent dry weight) in sediment and dissolved organic carbon (DOC; mg liter-‘) in interstitial water and overlying water (OW). DOC determined on last day of test. K, is the mean (sem) sediment to IW partition coefficient and K, is the (sem) log,, sedoc to IW partition coefficient. Sand is the percentage (dry weight) of particles > 63 urn diameter. Ratio TOC is the dry weight of added TOC to that naturally present in the preamended sediment. Ew sed test was not amended with TOC. n is the number of TBT centrations in that treatment excluding the controls. * is not measured. See methods for definitions and equations.

Armandia brevis (Ab)

Eohaustorius washingtonianus (Ew)

Rhepoxynius abronius (Ra)

Species

Sediment

(IW) mean ratio con-

228

J. P. Meador et al.

error of 198 (n = 22). Based on equation (13) and this Kdoc, only a small amount of the TBT found in IW was predicted to be sorbed to DOC. The mean (sd) percentage of total TBT in IW that was not sorbed (= free) to DOC was as follows: Ra 90% (7%); Ab 82% (10%); and Ew 85% (estimated; DOC not measured). The measured concentration of TBT in sediment for a given treatment was generally close to the nominal concentration (mean (sd) was 89% (49%)). The mean change for TBT concentrations in sediment over time in a given experiment indicates that the variation between replicates for a treatment (over days) was generally small (Table 2). Both increases and decreases were observed, which was probably due to sampling and analytical variation. Because sediment sampled before exposure and after exposure contained essentially the same TBT concentration, we assumed that no substantial depletion by the infauna or bacterial degradation had occurred. Also, within a TBT treatment, over levels of TOC, the day 0 and day-end mean concentrations of TBT in sediment were all very similar. The mean (sd) coefficient of variation (CV) for all TBT treatments was 24 (9)% for A. hrevis and 9 (7)% for R. abronius, (n=4 TBT concentrations for each test), indicating that even though TOC varied, sediment TBT concentration was relatively constant for a given TBT treatment. The water quality parameters for each experiment (Table 2) were within the range allowed by ASTM (1990). Total ammonia was generally very low (225 ug ml-‘) in each test; however it was elevated (40-50 ug ml-‘) in a few IWs. At this pH and temperature, only 0.8 to 1.8% of the total ammonia would occur in the unionized form (Emerson et al., 1975). Mean (sd) dry to wet weight ratios for the amphipods were as follows: E. washingtonianus 0.21 (0.03) n=8; R. abronius 0.27 (0.03) n= 19; A. brevis 0.23 (0.02) n = 3. The wet weight LDs,,s and BCFs can be computed by multiplying the dry weight With organisms

- 0.87’TOC

Log Koc = 5.12

w

Ab+

l

Ew+

A

Ra+

0

Ab-

0

Ew-

Without organisms Log Koc = 4.6

0

0.1

0.2

- 0.37’TOC

0.3

0.5 0.6 0.4 TOC (percent)

0.7

0.6

0.9

1

Fig. 1. K,, as a function of organic carbon in sediment. Plot of the loglO organic carbon normalized sediment-interstitial water partition coefficient (loglO K,,) as it varied over percent organic carbon (dry weight) in sediment. Plotted in two groups, organisms present and absent for most of experiment. Treatments with organisms are closed symbols and treatments without organisms are open symbols. Each sediment test shown; Ra is R. abronius, Ab is A. hrevis, Ew is E. washingtonianus.

25

32

15

5

C Mort %

5.3 (0.5)

6.6 (1.3)

6.6 (1.3)

7.7 (-)

Lipid {% dry)

(Z)

1.8 (-)

(!)

(G3)

ow

PH

(Z) (2,

+

*

*

*

(Z)

19.9 (17.9)

*

*

7.2 (4.3)

Ammonia o w/” ml-‘) IW

(&)

IW

30.8 (0.4)

33.0 (-)

32.5 (0.7)

32.3 (1.2)

Saln %0

12.4 (0.6)

14.0 (0)

14.2 (0.4)

13.2 (0.4)

Temp aC

TABLE 2 and bioassay parameters for TBT sediment tests

5.9 (3.9)

9.4 (4.6)

+

(S)

TBT IW:O w

(273

:zi)

*

&

TBT Sed % A

10

41

9

10

Duration (days)

C mort is the acetone control mortality on the day indicated. Lipid is the mean (sd) percent dry weight lipid content at the beginning of the test. pH, total ammonia (NH, + NH4), saln (salinity), and temp (temperature) are means (sd) for all treatments over all days. Ratio of IW to OW is the mean (sd) ratio of the concentrations of TBT in interstitial water to that in overlying water measured on last day. Sed % A is the mean (sd) percent change in TBT concentration measured in sediment on the first and last days for all treatments. Duration of test in days. * is not determined and _ is insufficient data.

Ab

Ew

Ew

Ra

Test

Water-quality

g S t?. 2 Y B P 2 F

2 & 2 L s B _. F Q z S’ S

230

J. P. Meador et al.

values by these ratios. Mean (sd) lipid content is shown in Table 2. Based on the lipid elimination rates provided by Meador (1993) for the amphipod species, we did not expect a substantial decline in whole body lipid values over the time course of these experiments. Lipid content in the polychaete over time was not determined. There were large intraspecific and interspecific differences in response to TBT-associated sediment on day 10 and even greater differences occurred when the long term results were considered (Table 3). In Table 3, IW-LCsos, sed,-L&s, and LDses were calculated with all data from an experiment and in Figs 2 and 3, we present LCsa values for each TOC treatment. When expressed in terms of bulk sediment concentration, variable TOC in sediment had a dramatic effect on the toxicity response (Fig. 2); however when toxicity was expressed in terms of IW only small differences occurred over sediment TOC content (Fig. 3). The highest TOC treatment produced mortality that was less than 50% in both the Ab and Ra experiments, which precluded determination of the LCso; however, after day 8, the number of individuals dying was doubling daily and probably would have surpassed 50% mortality within a day or two. To further highlight the effect of TOC on the organism’s response, we plotted the tissue concentration in R. abronius as a function of TBT concentration in sediment and TOC level (Fig. 4). Most to all of the individuals died in the high TBT-low TOC treatments and exhibited tissue concentrations that were consistent with the LDs9. Because the LCso and RCF values between species were different, we compared observed values (Table 3) with p.redictions based an a species’ uptake and elimination rates (Table 4). Table 4 shows the k, values calculated in these tests, the kz values from a related study (Meador, 1996), and predictions for IW-L&e, sediment-L&,, time to steady-state tissue concentration, BCF, and BAFt,, values. The LDso was very similar in all three species and significant mortality in all species was observed when tissue levels were above approximately 35-80 ug g-’ (dry weight;

20% mortality

”

F

o

Ew

P 10 P

o

Ab

7

3

33% mortality

P

a, 8

/

0

~‘~~I”“I”‘~I”“I”“I”“I”“I~“‘t”“1””l

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1

TOC (% dry wt.) Day 10 LCso values for sediment-associated tributyttin (TBT). LC&J and 95% confidence intervals for each treatment baqed on whole sediment concentrations of TBT (dry weight) as a function of TOC (percent dry weight), in sedi,ment. Three sediment tests as listed in Table 11Highest TOC treatment produced mortalities less than 50%.

Fig. 2.

(2.3 ‘-“,.2)

7.4 (6.3 - 8.8)

LC50

7.8 x IO4 (6.3 - 10.6 x 104)

1.7 x 10s (1.5 -2.1 x 105)

3.5 x 106 (3.2 - 3.8 x 106)

se&

zw

34.3 (27.2 - 44.3)

49.2 (40.6 - 61.2) 16411 (2670)

-

1836

(273)

69.1

BCF

(64.4 - 74.6)

L&o

results from tributyltin

TABLE 3

4.6 (0.2)

-

-

97.5 (19.5)

(0.04)

0.41

BAFim

(K)

BAF

sediment tests

41

9

10

Day

7

7

12

n

Ab+

27.6 (24.9 - 30.8)

9.3 x 105 (8.3 - 10.6 x 105)

89.4 4213 30.7 10 12 (& (81.2 - 99.8) (467) (7.0) _ IW LCsO (ng ml-‘), Sed, LC5e (ng TBT g-’ organic carbon), and LD5c (ug g-‘) values with 95% confidence intervals shown. BCF, BAF, and BAFt, values are means (sem) from the treatments where most of the individuals were alive at the end of the test. All sediment and tissue concentrations as dry weight. IW is total dissolved concentration of TBT in interstitial water. Day indicates that day for determination of the mortality and accumulation estimates. n is the total number of treatments. + indicates that results of mortality response (LCsc and LDse) in treatments corrected (increased) by adding control mortality to offset non-toxicant mortality. - is insufficient data.

Ew+

Ew

38.7

Ra

(35.7 - 42.2)

zw

LC50

Test

Toxicity and bioaccumulation

J. P. Meador et al.

232

= 114-307 nmol g-t). The LDsc calculated in terms of lipid content, produced essentially the same differences between species as those observed for the corresponding LDSos based on dry weight because total, whole-body lipid levels (based on dry weight) for the three species were generally the same (Table 3).

o

Ab

0

Ra

20% mortality

T

w/

mortality

o:,,,,,,,,,,.,,,,,,,,,,,,,,,,.,,,,,~,,,,~,,,,,,,,,,, 0.1

0

0.2

0.3

0.4 0.5 0.6 TOC (% dry wt.)

0.7

0.6

0.9

1

Fig. 3. Day 10 interstitial water (IW) tributyltin (TBT) LCso values. LCsc and 95% confidence intervals for total dissolved TBT in interstitial water from two sediment tests as total organic carbon (percent dry weight) varied. See Fig. 1 for abbreviations. Highest TOC treatment produced mortalities less than 50%.

-

I

lb

1 Sediment TBT (pg g-’ dry wt.)

Fig. 4. Bioaccumulation of tributyltin (TBT) from sediment exposure in R. abronius. TBT concentrations (dry weight) in tissue for various combinations of sediment TBT and total organic carbon (TOC; percent

dry weight).

Values are means and standard day 10, whichever came first.

deviations

at time of death or on

256 (19)

854 (189)

455 (46)

Ra

EW

Ab

0.08 (0.02)

0.04 (0.002)

0.18 (0.0009)

15.7

2.3

48.5

LCs* IW (ng ml-l)

2.6

0.26

14.5

5688

21 350

1439

BCF IW

Steady-State

4.2

7.9

0.49

11x4F/w

Prediction

8.7

36.9

74.1

16.6

Time to 95% BB

and toxicity

28.6

7.0

58.5

LCjO IW (ng ml-‘)

3128

7045

1194

BCF IW

Day 10 Prediction

0.55

0.33

0.83

fss

Meador (1996). kt is mean (sem) for two or more determined with equations in methods section. Day estimated with observed value on day 10/f,,. Sedi(day 10 LCse sed,, * day 10 fss * f,,). All tissue and

17.3

3.9

Time to 50% BB

TABLE 4 and predictions of bioaccumulation

LCJo Sed (pgg-‘)

toxicokinetics

Clearance uptake (kt) determined for sediment test and elimination rate constant (k2) from treatments. LCse, BCF, BAFt,,, time to 50% or 95% steady-state body burden (BB), f,,, and kt 10 BCF determined from steady-state predicted value multiplied by f,, and steady-state BAF,,, ment LCss at steady-state determined for 0.5% TOC (f,,c =O.OOS) sediment using the equation: sediment concentrations expressed in dry weight.

kl (ml g-’ d-l)

Test

Tributyltin

?z zt $ ~ w

5 s ;

; ? 6 5 5 3 2

oy ti f. 2

234

J. P. Meador

et al.

DISCUSSION Experiments with three species of infaunal invertebrates, exhibiting different modes of feeding, showed that total organic carbon in sediment strongly influenced the bioaccumulation and toxicity of TBT, primarily through the control of IW concentrations. Data from these experiments showed large differences between species in their response to TBT that appeared to be predictable with toxicokinetic rate constants and the degree to which steady-state conditions had occurred between tissue and environmental concentrations. In each of these toxicity tests, the predictions for bioaccumulation and toxicity, based on the rates of aqueous uptake and elimination, were close to the observed results. Several important aspects of TBT partitioning were revealed by the present study. The salient finding for TBT partitioning chemistry is that the K,, for this compound was approximately 25 100, when organisms were absent for most of a test. Additionally, the K,, for TBT, within the range of TOC levels tested, was much higher than the reported K,,, whereas the Kdoc was approximately three fold lower than the K,,. Partition chemistry

In sediment, the concentration of TBT was only 100 to 200 times as great as that in water; however, when expressed in terms of organic carbon, the partition coefficient was approximately 25 100 for TOC levels between 0.3 and 1.O%, When organisms were present, the K,, values displayed a strong inverse relationship with TOC content of the sediment producing a mean about twice as high and a CV about four times as large than the treatments without organisms present for most of the test. This two fold difference in K,, values found between treatments that did or did not contain organisms for the majority of the duration of the test was almost certainly due to the variation in IW concentrations. Even though the inverse relationship between K,, and TOC was evident without organisms present for several days, the K,, values were similar enough to allow determination of a mean K,, (25 100) with a CV of only 22%. As seen in Table 1, most of the elevated K,, values occurred in the Ra test, which contained the lowest TOC treatments and the highest number of individuals. Because only two of the seven total treatments contained substantial amounts of added TOC (Table l), we believe that the partitioning behavior observed here for TBT would be reflective of that expected under natural conditions. It is generally believed that the partition coefficient between water and organic normalized sediment concentrations (K,,) for nonionic organic compounds is essentially equivalent to a compound’s K,,, and that the Kdoc is approximately equal to its K,, (Di Toro et af., 1991). Our observed mean K,, (without organisms) was about five times higher than the reported K,, for TBT, which occurred over TOC levels considered average for US West Coast marine sediments (mean (sem) = 0.93 (O.OS)%; Meador et al., 1994). Because the TBT K,, does not match the predicted value, it is probable that other factors are important for controlling its partitioning behavior. Tributyltin is an ionizable, organometallic compound, hence its partitioning may be controlled by both organic carbon sorption and anionic surface complexation. Overall, the K,, values from our study are similar to those reported by Unger et al. (1988) for marine sediments from Chesapeake Bay, including the observation of a higher value for a low TOC sediment. Higher K, values could be construed to mean stronger associations of TBT with sediment as organic carbon content declined (i.e. less TBT desorbed into IW for a given

Toxicity of tributyltin to infaunal invertebrates

235

sed,), which may have been caused by alterations in the physicochemical nature of sediment or interstitial water. For example, the complexation or sorption capacity of sediment may have increased as TOC declined, possibly due to a higher redox state leading to increased oxyhydroxides of Fe and Mn. Another explanation for the anomalously high Kocs is that the activity of organisms under the sediment surface was responsible for lower IW concentrations of TBT by either diluting the IW with overlying water, which had lower concentrations of TBT, or accumulating this compound faster than it could desorb from sediment into IW. Because this trend was stronger in treatments with organisms present, either process (dilution of IW or uptake rates exceeding desorption rates) seems probable, but indistinguishable. Although not measured directly, we expected that TBT in sediment and porewater would rapidly come to equilibrium. Previous work by Unger et al. (1988) found rapid desorption rates for TBT, with 60% of equilibrium reached within 30 min. Similar studies with organic compounds in this K,, range support rapid desorption (Karickhoff, 1980; Karickhoff & Morris, 1985; Wu & Gschwend, 1986) which would make the hypothesis of uptake rates exceeding desorption rates unlikely. Dilution, due to a change in sediment porosity with varying percentage fines allowing more OW to be pumped in by organisms, may explain part of the anomaly; however almost all of the higher K,, values occurred in the Ra test and the differences between TBT in OW and IW in this test were not large (Table 2). Conversely, only minor differences in K, were observed in the Ab test, which displayed a large differential in TBT between OW and IW. Further studies are necessary to extend the findings with sediments containing more than 1% TOC and to help elucidate this important phenomenon of variable K,,, which may have a major impact on interpreting toxicity bioassays. The Kdoc for TBT (1652) was about 17 fold lower than the standard K,,, and approximately three fold lower than the K,, (Kdoc=0.3 * K,,). The Kdoc for many contaminants of interest is reported to be in the range of 0.1-1.0 * K,, (Landrum et al., 1985; Chiou et al., 1987; McCarthy et al., 1989) which we feel is comparable to our TBT value. It is possible that other factors in sediment, in addition to organic carbon, control the partitioning of TBT between sediment and water; however, once in the dissolved phase, TBT behaves more like an organic compound with respect to its partitioning between DOC and water. The DOC in IW was relatively high in some of the treatments (up to 150 mg C liter-‘) but because the Kdoc was relatively low, only a small fraction of the dissolved TBT would have been associated with DOC. Consequently, the limited partitioning of TBT to DOC in IW would have produced only a minor reduction in free interstitial water TBT and therefore very little change in the bioavailability of this compound to the infauna from this route of exposure. Accumulation of TBT We propose that dissolved TBT will account for almost all of the TBT found in the tissues of similar infaunal invertebrates when TBT concentrations in water and sediment are at, or close to equilibrium, implying that the mode of feeding may not be a factor in determining TBT accumulation under these conditions. While the route of exposure can not be discerned in an experiment where all phases (organism, sediment, and IW) are in steady state (Di Toro et al., 1991) the uptake kinetics can be affected, depending on the relative

J. P. Meador et al.

236

contribution

of food, sediment, and water (Bierman, 1990). Comparison of IW LCSos for and A. brevis, which were probably not close to steady-state conditions after 9 or 10 d (Table 4) exhibited values that were very close to predicted values for that time period. If ingestion or dermal contact were important, then for a given IW concentration, the rate of uptake would have been faster and the L&e lower for this period of exposure because of the increased tissue concentrations due to these other routes of accumulation that were not included in the calculation for aqueous uptake (k,). Consequently, because all species exhibited the same or lower uptake clearance of TBT in IW compared to that for water-only, it is likely that the majority of uptake of TBT from sediment exposure occurred through IW. Our results are consistent with studies of other compounds that concluded the majority of an organism’s tissue concentration will come from uptake of the dissolved compound when that compound exhibits a partition coefficient (log K,,) less than approximately 5 to 5.5 and sediment and water concentrations are at, or near, equilibrium (Landrum & Robbins, 1990; Meador et al., 1995). Of course, when aqueous exposure concentrations are far below equilibrium, which may be the case for epibenthic or pelagic organisms, the importance of sediment ingestion and dietary uptake may be enhanced. For example, Langston & Burt (1991) concluded that sediment was an important route of uptake for the deposit-feeding clam (Scrobiculuriu plunu) when exposed to TBT in water-only vs sediment exposures. Their assumed K, for sediment and water was 10000, resulting in water concentrations far below those expected under equilibrium conditions. When aqueous TBT concentrations are less than expected, food and sediment could become major contributors to TBT body burden; however we predict that these body burdens should be below those observed under worstcase exposure conditions when water and sediment are in equilibrium. Because TBT is an ionizable organometallic compound, there is no a priori reason to expect that it will bioaccumulate according to the predictions of EqP (Di Toro et al., 1991). A further confounding factor is that the K,, for TBT was five times less than the K OW? . hence, the predicted IW concentrations would be five times less than assumed by EqP. Consequently, the utility of the BAFi,, in this study is strictly for comparison. In this study we found large interspecific differences in BAFr,, values, which, in theory, were probably due to variation in rates of metabolism. Interestingly, while R. ubronius displayed an elimination rate (determined in clean sediment) that was only 4.5 times faster than E. washingtonianus (Meador, 1996) its near steady-state BAFi,, was predicted to be more than 16 fold lower. We are unsure which factors would explain the additional lowering of the BAFi,, over that expected from metabolism alone; however some of this may be due to variable TBT partitioning. Based on the disparity in their rates of elimination, the differences in BAF,,, for E. washingtonianus and A. brevis were as predicted. E. washingtonianus

Toxicity

of TBT-associated

sediment

It is clear that organic carbon had a strong effect on the toxicity of sediment-associated TBT to these infaunal invertebrates. By varying organic carbon and holding grain-size relatively constant, we have demonstrated that this one factor can have a major effect on the bioavailability of TBT, as evidenced by the four or five fold difference in sediment L&s and tissue accumulation over the range of TOC levels tested. Although we can’t extrapolate beyond the results of this test, it is likely that as TOC levels increase, tissue

Toxicity of tributyltin to infaunal invertebrates

231

residues and toxicity would continue to decrease. Many other studies have shown similar results for organic toxicants (Adams et al., 1985; Nebeker et al., 1989; Swartz et al., 1990). The IW LCSOvalues for E. washingtonianus and A. brevis were two to four times higher than those observed for water-only exposure (Meador et al., 1993) which can be explained with the rates of uptake and elimination. While the kr based on IW concentrations for R. ubronius was remarkably similar to those values reported earlier for wateronly exposure (Meador et al., 1993; Meador, 1993), the IW kr values for both E. wushingtoniunus and A. brevis exposed to sediment were lower than those found in water-only exposures (Meador et al., 1993; Meador, 1996) which could lead to a lower than expected IW-BCF and a higher IW-LCSO (assuming k2 is the same for each exposure). In water-only experiments, the kr for E. wushingtoniunus was usually in the range of 2OOG4000 ml g-l d-i (Meador et al., 1993; Meador, 1996); however, in the sediment test, the two highest TBT treatments produced a much reduced uptake clearance in terms of IW exposure (k, = 854 ml g-i d-i). This kr from exposure to IW and the elimination rate determined in another study (Meador, 1996) were used in equations (3) and (10) to predict the BCF and LC5,, (Table 4), which were found to be very close to the observed values in these sediment tests (Table 3). Excellent agreement was also found between the observed and expected IW-LCSO and IW-BCF for A. brevis, even though there was a decrease of 1.6 fold in k, between water-only and IW exposures. The reduced ki and elevated L&s values are consistent with the inflated K,, values observed when animals were present. Because we measured IW concentrations at the end of a test, it is possible that when organisms were present, IW exposure concentrations of TBT were lower (due to dilution with OW). Once these test animals died, the IW concentrations could increase towards their equilibrium concentrations producing higher IW concentrations than actually experienced by the organisms. A k, based on the higher, endof-test, IW concentrations than those experienced by the organisms would produce lower kr values. Future research examining the IW concentrations over time in the presence of bioaccumulating organisms will help resolve the importance of this uncontrolled variable. Another explanation for the reduced ki values in E. wushingtoniunus and A. brevis for IW compared to those in water-only exposures (Meador et al., 1993; Meador, 1996) may involve behavioral modification. It is possible that these species may have been differentially stressed when placed in water-only exposure systems due to the lack of substrate. Tributyltin has been shown to cause an increase in swimming velocity and reverse the normal direction of phototaxis in a crustacean (Meador, 1986) which may be exacerbated in some species by unnatural surroundings (i.e. a lack of sediment). It is believed that when water is the main route of uptake in a species, the majority of the tissue burden will be due to accumulation through ventilatory/respiratory surfaces because of their large surface area and high water flow. Consequently, an increase in activity stimulated by this compound or a lack of substrate may cause an increase in the rate of uptake. The LDSOSfor the amphipods and polychaete in this study were very similar to those reported for other organisms exposed to TBT (Tas, 1993; Meador, 1996). One other study (Moore et al., 1991) with polychaetes (Neunthes arenuceodentutu) has found substantial lethality (79%) when tissue burdens (17 pg g-l) were in the same range as that found to be lethal for A. brevis in water-only exposures (Meador, 1996). The LDSOsreported for the water-only (Meador et al., 1993; Meador, 1996) and sediment tests for the amphipods were not substantially different if one considers the 95% CI; however the polychaete did exhibit a higher LDso when exposed to sediment. While the sediment exposure LDsO for

238

J. P. Meador et al.

A. brevis was higher than that for the two amphipods, its water-only LDso (41 pg g-l) was closer to that observed for the other species (Meador, 1996). This elevated LDsO may have been caused by sediment in the gut leading to overestimated tissue concentrations of TBT. Considering the range in values for the amphipods and the 95% CI, we conclude that these differences in LDsO are relatively small and have only a relatively minor effect on the conclusions. The difference between day-10 IW LCsO values for R. abronius and E. washingtonianus exposed to TBT in sediment was large, but less than that reported for water-only exposure (Meador et al., 1993). Based on the predicted IW L&OS at steady state, E. washingtonianus may be approximately 21 times more sensitive than R. ubronius and seven times more sensitive than A. brevis to the effects of sediment-associated TBT; however, differences in the toxic response between species is larger when based on sediment concentrations. It should be noted that if the k, values were underestimated because of an artifact of the bioassay, the steady-state predictions given in Table 4 would substantially underestimate the potential response by these species in the field.

CONCLUSIONS The findings from this study show that TBT associates with organic carbon, and possibly other sediment ligands, and that the amount of dissolved TBT is directly controlled by this association. When sediment and water concentrations of TBT are near equilibrium, it appears that aqueous uptake accounts for essentially all of the acquired body burden in these infaunal species with diverse feeding modes. The results of this study also highlight the important points that one species may not exhibit a response to a contaminant that is representative for other similar species and that a given sediment may produce a response that is not commensurate with the response produced by other sediments with varying physical/chemical features. Even though different species may exhibit very similar LDSos, the amount of toxicant available for uptake and the rates by which a species accumulates and eliminates the chemical will have a very large effect, both spatially and temporally, on its response.

ACKNOWLEDGEMENTS We would like to thank Casey Rice, Bob Snider, Le Tran, and Paul Plesha for technical support, Susan Picquelle for statistical advice, and William Reichert, Casey Rice, and John Stein for reviewing the manuscript and providing insightful comments.

REFERENCES Adams, W. J., Kimerle, R. A. & Mosher, R. G. (1985). Aquatic safety assessment of chemicals sorbed to sediments. In Cardwell, R. D., Purdy, R. & Bahner, R. C. (eds) Aquatic Toxicology and Hazard Assessment: Seventh Symposium. Special Technical Publication 854, American Society for Testing and Materials, Philadelphia, PA, USA, pp. 429453. APHA (1985). Standard Methods for the Examination of Water and Wastewater, 16th Edn. APHA, AWWA, and WPCF. Amer. Public Health Assoc., Washington, DC, USA.

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ASTM (1990). Standard Guide for Conducting IO-day Static Sediment Toxicity Tests with Marine and Estuarine Amphipodr, E 1367-90. American Society for Testing and Materials, Philadelphia, PA, USA, 24 pp. Bierman, V. J. (1990). Equilibrium partitioning and biomagnification of organic chemicals in benthic animals. Environ. Sci. Technol., 23, 1407-1412. Buchanan, J. B. & Kain, J. M. (1971). Measurement of the physical and chemical environment. Sediments. In Holme, N. A. & MacIntyre, A. D. (eds) Methods for Studying the Marine Benthos. IBP Handbook, No. 16. Blackwell Scientific, pp. 3(X51. Cardwell, R. D. & Meador, J. P. (1989). Tributyltin in the environment: an overview and key issues. In Ocean’s 89 Proceedings, Vol. 2. Organotin Symposium, Institute of Electrical and Electronics Engineers, Piscataway, NJ and Marine Technology Society, Washington, DC, USA, pp. 537544. Chiou, C. T., Kile, D. E., Brinton, T. L, Makolm, R. L. & Leenheer, J. A. (1987). A Comparison of Water Solubility Enhancements of Organic Solutes by Aquatic Humic Materials and Commercial Humic Acids. Environ. Sci. Technol., 21, 1231-I 234. Di Toro, D. M., Zarba, C. S., Hansen, D. J., Berry, W. J., Swartz, R. C., Cowan, C. E., Pavlou, S. P., Allen, H. E., Thomas, N. A. & Paquin, P. R. (1991). Technical basis for establishing sediment quality criteria for nonionic organic chemicals using equilibrium partitioning. Environ. Toxicol. Chem., 10, 1541-I 583. Dowson, P. H., Bubb, J. M. & Lester, J. N. (1993). Temporal distribution of organotins in the aquatic environment. Mar. Pollut. Bull., 26, 487-494. Emerson, K., Russo, R. C., Lund, R. E. & Thurston, R. V. (1975). Aqueous ammonia equilibrium calculations: effect of pH and temperature. J. Fish. Res. Bd. Canada, 32, 2379-2383. Espourteille, F. A., Greaves, J. & Huggett, R. J. (1993). Measurement of tributyltin contamination of sediments and Crassostrea virginica in the southern Chesapeake bay. Environ. Toxicol. Chem., 12, 305-314. Fent, K. (1996). Ecotoxicology of organotin compounds. Critical Rev. Toxicol., 26, l-1 1I. Heard, C. S., Walker, W. W. & Hawkins, W. E. (1989). Aquatic toxicological effects of organotins: an overview. In Ocean’s 89 Proceedings, Vol. 2. Organotin Symposium, Institute of Electrical and Electronics Engineers, Piscataway, NJ and Marine Technology Society, Washington, DC, USA, pp. 554563. Hedges, J. L. & Stern, J. (1984). Carbon and nitrogen determination of carbonate containing solids. Limnol. Oceanogr., 29, 657663. He&es, S. E. & Allen, C. P. (1983). Lipid quantification of freshwater invertebrates: method modification for microquantitation. Can. J. Fish. Aquatic Sri., 40, 1315-I 3 17. Karickhoff, S. W. (1980). Sorption kinetics of hydrophobic pollutants in natural sediments. In Baker, R. A. (ed.) Contaminants and Sediments, Vol2. Analysis, Chemistry, Biology. Ann Arbor Science, Ann Arbor, MI, USA, pp. 193-205. Karickhoff, S. W. & Morris, K. R. (1985). Impact of tubificid oligochaetes on pollutant transport in bottom sediments. Environ. Sci. Technol., 19, 51-56. Kerr, D. & Meador, J. P. (1996). Modeling dose-response with generalized linear models. Environ. Tox. Chem., 15, 395-401. Krahn, M. M., Moore, L. K., Bogar, R. G., Wigren, C. A., Chart, S-L. & Brown, D. W. (1988). High-performance liquid chromatographic method for isolating organic contaminants from tissue and sediment extracts. J. Chromatogr., 437, 161-175. Krone, C. A., Brown, D. W., Burrows, D. G., Bogar, R. G., Chan, S-L. & Varanasi, U. (1989a). A method of analysis of butyltin species and measurement of butyltins in sediment and English sole livers from Puget Sound. Mar. Environ. Res., 27, I-18. Krone, C. A., Brown, D. W., Burrows, D. G., Chan, S.-L. 8~ Varanasi, U. (19896). Butyltins in sediment from marinas and waterways in Puget Sound, Washington State, USA. Mar. Pollut. Bull., 20, 528-53 1. Krone, C. A., Stein, J. E. & Varanasi, U. (1996). Butyltin contamination of sediments and benthic fish from the East, Gulf and Pacific coasts of the United States. Environ. Monit. Assess., 40,7589. Landrum, P. F. & Robbins, J. A. (1990). Bioavailability of sediment-associated contaminants to benthic invertebrates. In Baudo, R., Giesy, J. P. & Muntau, H. (eds) Sediments: Chemistry and Toxicity of In-Place Pollutants. Lewis Publishers, Ann Arbor, MI, USA, pp. 237-263.

240

J. P. Meador et al.

Landrum, P. F., Nihart, S. R., Eadie, B. J. & Gardner, W. S. (1984). Reverse-phase separation method for determining pollutant binding to Aldrich humic acid and dissolved organic carbon of natural waters. Environ. Sci. Technol., 18, 187-192. Landrum, P. F., Reinhold, M. D., Nihart, S. R. & Eadie, B. J. (1985). Predicting the bioavailability of organic xenobiotics to Pontoporeia hoyi in the presence of humic and fulvic materials and natural dissolved organic matter. Environ. Toxicol. Chem., 4, 459467. Langston, W. J. & Burt, G. R. (1991). Bioavailability and effects of sediment-bound TBT in deposit feeding clams, Scrobicularia plana. Mar. Environ. Res., 32, 61-77. Laughlin, R. B., Guard, H. E. & Coleman, W. M., III (1986). Tributyltin in seawater: speciation and octanol-water partition coefficient. Environ. Sci. Technol., 20, 201-204. Maguire, R. J. (1987). Environmental aspects of tributyltin. Appl. Organometallic Chem., 1, 475498. McCarthy, J. F., Roberson, L. E. & Burrus, L. W. (1989). Association of benzo(a)pyrene with dissolved organic matter: prediction of Kdom from structural and chemical properties of the organic matter. Chemosphere, 19, 19 1l-1 920. Meador, J. P. (1986). An analysis of photobehavior of Daphniu magna exposed to tributyltin. In Ocean’s 86 Proceedings, Vol. 4. Organotin Symposium, Institute of Electrical and Electronics Engineers, Piscataway, NJ and Marine Technology Society, Washington, DC, USA, pp. 12131218. Meador, J. P. (1993). The effect of laboratory holding on the toxicity response of marine infaunal amphipods to cadmium and tributyltin. J. Exper. Mar. Biol. Ecol., 174, 2277242. Meador, J. P., Varanasi, U. & Krone, C. A. (1993). Differential sensitivity of marine infaunal amphipods to tributyltin. Mar. Biol., 116, 231-239. Meador, J. P., Clark, R. C., Robisch, P. A., Ernest, D. W., Landahl, J. T., Varanasi, U., Chan, S-L. & McCain, B. (1994). National Benthic Surveillance Project: Pac$ic Coast. Trace Element Analyses of Cycles I to V (1984-1988). NOAA Technical Memorandum NMFS-NWFSC-16. U.S. Dept. of Commerce, National Oceanic and Atmospheric Admin, National Marine Fisheries Service, 206 pp. Meador, J. P., Stein, J. E., Reichert, W. L. & Varanasi, U. (1995). A review of bioaccumulation of polycyclic aromatic hydrocarbons by marine organisms. Rev. Environ. Contam. Tox., 143, 79165. Meador, J. P. (1996). Comparative toxicokinetics of tributyltin in five marine species and its utility in predicting bioaccumulation and acute toxicity. Aquat. Tox. (in press). Moore, D. W., Dillon, T. M. & Suede], B. C. (1991). Chronic toxicity of tributyltin to the marine polychaete worm Neanthes arenaceodentata. Aquat. Tax., 21, 181-198. Nebeker, A. V., Schuytema, G. S., Griffis, W. L., Barbitta, J. A. & Carey, L. A. (1989). Effect of sediment organic carbon on survival of Hyalellu uzteca exposed to DDT and endrin. Environ. Touicol. Chem., 8, 705-718. Okamato, K. (1991). Biological reference materials for metal speciation: NIES fish tissue reference material for organotin compounds, In Subramanian, K. S., Iyengar, G. V. & Okamoto, K. (eds) Biological Truce Element Research: Multidisciplinary Perspectives. ACS Symposium Series no. 445. Amer. Chem. Sot., Washington, DC, USA, pp. 257-264. Ozretich, R. J., Smith, L. M. & Roberts, F. A. (1995). Reversed-phase separation of estuarine interstitial water fractions and the consequences of C is retention of organic matter. Environ. Toxicol. Chem., 14, 1261-1272. Stephan, C. E. (1977). Methods for calculating an LC5e. In Mayer, F. L. & Hamelink, J. L. (eds) Aquatic Toxicology and Hazard Evaluation, First Symposium. Special Technical Publication STP 634, American Society for Testing and Materials, Philadelphia, PA, USA, pp. 65-84. Swartz, R. C., Schults, D. W., Dewitt, T. H., Ditsworth, G. R. & Lamberson, J. 0. (1990). Toxicity of fluoranthene in sediment to marine amphipods: a test of the equilibrium partitioning approach to sediment quality criteria. Environ. Toxicol. Chem., 9, 1071~1080. Tas, J. W. (1993). Fate and Effects of Triorganotins in the Aqueous Environment: Bioconcentration Kinetics, Lethal Body Burdens, Sorption and Physico-Chemical Properties. Ph.D. Dissertation, University of Utrecht. US Congress. (1988). Organotin antifouling paint control act of 1988. H.R. 2210-3. One Hundredth Congress of the United States of America, Second Session. Washington, DC, USA.

Toxicity of tributyltin to infaunal invertebrates

241

Unger, M. A., MacIntyre, W. G. & Huggett, R. J. (1988). Sorption behavior of tributyltin on estuarine and freshwater sediments. Environ. Toxicol. Chem., 7, 907-915. Valkirs, A. O., Davidson, B., Kear, L. L. & Fransham, R. L. (1991). Long-term monitoring of tributyltin in San Diego Bay, California. Mar. Environ. Res., 32, 151-167. Wade, T. L., Garcia-Roma, B. & Brooks, J. M. (1990). Butyltins in sediments and bivalves from U.S. coastal areas. Chemosphere, 20, 647-662. Wu, S-C. & Gschwend, P. M. (1986). Sorption kinetics of hydrophobic organic compounds to natural sediments and soils. Environ. Sci. Technol., 20, 717-725.