toxicity identification evaluation of metal-contaminated sediments using an artificial pore water...

509

Environmental Toxicology and Chemistry, Vol. 18, No. 3, pp. 509–518, 1999q 1999 SETAC

Printed in the USA0730-7268/99 $9.00 1 .00

TOXICITY IDENTIFICATION EVALUATION OF METAL-CONTAMINATED SEDIMENTSUSING AN ARTIFICIAL PORE WATER CONTAINING

DISSOLVED ORGANIC CARBONS

ANNE M. BOUCHER and MARY C. WATZIN*School of Natural Resources, Aiken Center, University of Vermont, Burlington, Vermont 05405, USA

(Received 21 January 1998; Accepted 16 June 1998)

Abstract—Recent investigations of sediment-associated pollutants in Lake Champlain indicated significant contamination with As,Mn, and Ni in Outer Malletts Bay, Vermont, USA. Ceriodaphnia dubia exposed to sediment pore water from several sites in OuterMalletts Bay showed repeatable, acute mortality at only one site. A toxicity identification evaluation (TIE) was conducted on porewater to determine the contaminants causing mortality at this site. Unlike most TIE applications, the dilution water used in thesetests was formulated to match the hardness, alkalinity, pH, conductivity, and dissolved organic carbon content of the pore water.Results from phase I of the TIE indicated that divalent metals may be responsible for toxicity. Phase II results revealed levels ofMn above LC50 values. Spiking experiments employed in phase III confirmed Mn as the principal toxicant in sediment pore water.The formulated pore water worked well and helped ensure that toxicant behavior was influenced primarily by each TIE manipulationand not by physical and chemical differences between the dilution and site water. Although the Mn toxicity at this site may be theresult of its unique geomorphology, this situation underscores the need to look broadly for potential toxicants when evaluatingcontaminated sites.

Keywords—Arsenic Manganese Pore water Ceriodaphnia dubia Toxicity identification evaluation

INTRODUCTION

Sediments are a sink for many anthropogenic contaminantsentering aquatic systems. The resulting mixture of pollutantscan become a diffuse source of toxic compounds to overlyingwater and to organisms living in or on the sediments [1].Naturally occurring toxicants such as ammonia [2] and hy-drogen sulfide [3] can also contribute to the toxicity of sedi-ments. Determining which contaminants are the cause of tox-icity in sediments can be difficult.

The sediments in Outer Malletts Bay (Vermont, USA) inLake Champlain are an example of just such a complicatedsituation. In 1993, the sediments of Outer Malletts Bay werefound to be contaminated with As, Mn, and Ni [4]. At severallocations, the concentration of these contaminants exceededproposed sediment quality guidelines [5,6]. Because As con-centrations reached as high as 176 mg/g, initial attention fo-cused on this contaminant. The potential toxicity of bay sed-iments was investigated using acute Ceriodaphnia dubia testsin interstitial (pore) water extracted from surficial sediments.Although a number of sites from throughout the bay weretested, only one site in the deepest part of the bay (site OMB-7; 448349510N, longitude 738189050W; Fig. 1) was consistentlyacutely toxic to C. dubia (unpublished data). However, thissite was not the location of the highest As concentration mea-sured in the bay.

The objective of this study was to identify the agent oragents of toxicity at site OMB-7 in Outer Malletts Bay. Themethod used was toxicity identification evaluation (TIE), aprocedure developed by the U.S. Environmental ProtectionAgency (U.S. EPA) in the late 1980s to determine the toxic

* To whom correspondence may be addressed([email protected]).

agents in complex effluents, but now widely applied to ambientwaters and sediments [7–10].

Guidelines describing the three parts of a TIE [11–13] havebeen published. Phase I involves characterizing the class orclasses of contaminants contributing to toxicity. In phase II,specific toxicants that could be responsible for the observedtoxicity are identified. Finally, in phase III, various approachesare used to confirm that the toxicants identified in phase II arethe cause of the toxicity observed in phase I. Because TIEprocedures were originally designed for effluent samples,aqueous samples are needed to conduct TIEs on contaminatedsediments. Evidence has shown that pore water extracted bycentrifugation provides a conservative estimate of bulk sedi-ment toxicity when metals are the contaminants of concern[14–16]; therefore, sediment pore water was used for this TIE.

In most TIEs, dilution water for toxicity tests is formulatedin the laboratory to match only the hardness of the samplewater. During this TIE, special efforts were made to accountfor the fact that other characteristics of water chemistry suchas alkalinity, pH, and the concentration of dissolved organiccarbon (DOC) also affect the chemistry and subsequent bio-availability of toxicants in pore-water solutions [10,15,17].Consequently, we developed an artificial pore water (APW)that mimicked not only the hardness, pH, and alkalinity of siteOMB-7 pore water but also the DOC content in order to re-move DOC as an additional source of variability in the toxicitytesting.

MATERIALS AND METHODS

Pore water

All utensils used to process and store sediments were madefrom high-density polyethylene and were acid washed andacetone rinsed before use. Sediments were collected with a

510 Environ. Toxicol. Chem. 18, 1999 A.M. Boucher and M.C. Watzin

Fig. 1. Sampling location for Outer Malletts Bay (Vermont, USA) toxicity identification evaluation.

Table 1. Water quality characteristics of site OMB-7 pore water

Characteristic Mean (n 5 9) 61 SD

pHPretestPosttest

HardnessAlkalinityConductivityDissolved organic carbon

7.078.24100 mg CaCO3/L115 mg CaCO3/L251 mS/m34 mg/L

0.580.76

11.07.0

22.0—

Ponar dredge at site OMB-7 (Fig. 1) up to a week beforebeginning TIE experiments. Sediments were collected on threeoccasions: November 7, 1994, March 11, 1995, and July 26,1995, avoiding periods of oxygen depletion in the bottom wa-ters. Sediments were placed into buckets and stirred by handuntil homogenized. The sediments were transferred to 5-L con-tainers that were kept cool during transportation to the labo-ratory. Sediments were stored at 48C in the dark for no longerthan 6 weeks with minimal head space and lids sealed withparafilm. On the day of use, pore water was extracted fromsediments by centrifugation. Sediment from previously un-opened 5-L containers was packed into 250-ml bottles andcentrifuged at 10,000 g 48C for 30 min. Immediately afterextraction, the pore water was decanted from the sedimentpellet and allowed to come to room temperature (25 6 28C)in an open container.

Dilution water

The artificial pore water used as dilution water was for-mulated to match the pH, hardness, alkalinity, conductivity,and DOC content of the site pore water (Table 1). First, re-constituted water was made by adding MgSO4·7H2O, CaSO4,NaHCO3, and KCl to type III water from ROPureSTt waterpurification system (Barnstead-Thermolyne, Dubuque, IA,USA). In an effort to create a standard dilution medium, so-dium humate (Aldrich Chemical, Milwaukee, WI, USA) was

used to approximate the DOC component of site OMB-7 porewater. Humic acid is the compound accounting for the majorityof DOC in pore water [18]. The amount of humic acid usedin the APW formula was determined spectrophotometrically.Humic acid absorbs light at a wavelength of 336 nm [19]. AShimadzu Recording Spectrophotometer (Model UV-160A,Shimadzu, Tokuyo, Japan) was used to measure the absorbenceof site OMB-7 pore water. A series of humic acid solutionsof known concentrations was prepared by serial dilution of aconcentrated aqueous solution of sodium humate (0.5 g/L) withtype III water. The absorbances of each of the known con-centrations were measured and a standard curve was prepared.This curve was used to determine the concentration of humicacid that corresponded to the mean level of absorbance of siteOMB-7 pore water. The mean humic acid concentration was

TIE with artificial pore water Environ. Toxicol. Chem. 18, 1999 511

found to be 34 mg/L in the pore water. The APW was com-pleted by adding 13.6 ml of stock humic acid solution (2.5g/L) to each liter of reconstituted water. The final APW hada hardness of 75 to 100 mg CaCO3/L and a pH of 8 to 8.3.

Toxicity tests

The cladoceran C. dubia was used for all toxicity testsassociated with the TIE. Organisms were reared in moderatelyhard reconstituted water at 25 6 28C and exposed to a 16:8h light:dark photoperiod. Mass cultures were fed concentratedSelanastrum capricornutum and a yeast–Cerophyll–TroutChowt (Ralston-Purina, St. Louis, MO, USA) solution daily[20]. Water in the mass cultures was changed at least twiceweekly.

Neonates less than 24 h old were obtained for use in toxicitytests by isolating adults that were ready to reproduce frommass cultures the day before an experiment. Individuals wererandomly selected from brood chambers and distributed among30-ml plastic cups containing 10 ml of test water. Five or-ganisms were placed into each cup and fed 66 ml of the yeast–Cerophyll–Trout Chow mixture at the beginning of each test.Toxicity tests were performed in a dilution series. Exposureto the test water lasted for 48 h. The survival of the testorganisms was checked at 24 and 48 h.

Statistics

Acute LC50 values and 95% confidence intervals were cal-culated by probit analysis or the trimmed Spearman–Karbermethod for all manipulations in which partial mortalities oc-cured in two or more intermediate dilutions [21]. In some casesthe dilution series used did not properly bracket the LC50concentration so partial mortalities did not occur. In such cases,95% confidence intervals could not be calculated.

Toxicity identification evaluation

Initial toxicity. Initial toxicity tests were conducted in con-junction with phase I manipulations on November 14, 1994,and July 26, 1995. The November test was performed usingmoderately hard reconstituted water as the dilution water,whereas the July toxicity test was performed using APW asthe dilution water.

Phase I: Characterization. Simple chemical and physicalmanipulations of the pore water were used to remove distinctgroups of compounds during phase I [11]. Immediately fol-lowing each manipulation, a toxicity test was performed, andLC50 values were calculated for manipulated and unmanip-ulated samples. The results were evaluated by visually com-paring LC50 values for alterations in toxicity. If toxicity waschanged as a result of a manipulation, that class of compoundswas implicated as causing toxicity.

To determine whether nonpolar organic compounds werecausing toxicity, pore water was passed through a solid-phaseextraction chromatography column containing an octadecyl(C18) sorbent (J.T. Baker Chemical Company, Phillipsburg, NJ,USA). The column was conditioned with HPLC-grade meth-anol, type III water, and three aliquots of moderately hardreconstituted water with the pH adjusted to 7 (the approximatepH of the pore water). The effluent from the last aliquot ofreconstituted water was used as the column toxicity blank.Aliquots of 100% pore water were then introduced to the col-umn. Effluent from the first 25-ml aliquot was collected anddiscarded. Effluent from the third aliquot was used for toxicity

testing. Effluents from the second and fourth aliquots wereused as procedure controls.

To determine whether inorganic cations could be causingtoxicity, a chelating agent (ethylenediaminetetraacetate ligand[EDTA]) was added to the pore water. A 0.031 M stock so-lution of EDTA was prepared and specific volumes were addedto 10-ml aliquots of test water. Three volumes (0.2 ml, 0.1 ml,and 0.05 ml) of EDTA stock solution were used, each addedto a full dilution series of pore water.

To test whether small changes in pH over a biologicallytolerable range influenced toxicity, samples of pore water wereadjusted to pH 6.5, 7.5, and 8.5 with 1.0, 0.1, and 0.01 N HCland NaOH. To control the pH for the duration of the test,zwitterionic hydrogen ion buffers were added to the test so-lutions. To maintain the pH at 6.5, 3 g/L of 2-(N-morpho-line)ethanesulfonic acid was used. To maintain the pH at 7.5,4 g/L of 2-(N-morpholine)propanesulfonic acid was used. Thehighest pH aliquot was adjusted to pH 8.5, but no buffer wasadded. The pH of the pore water solution rose to almost thislevel during the toxicity test; therefore, no buffer was neededto maintain the pH.

The pH adjustment test was repeated with the addition of0.1 ml of EDTA stock solution to the test cups. This combi-nation of tests was used to distinguish between metals andammonia as possible toxicants.

Sodium thiosulfate (Na2S2O3) was added to the pore waterto determine whether oxidants were causing toxicity. This testalso provided additional information on metal contaminationbecause the thiosulfate ion chelates some metallic cations [22].Two dilution matrices were set up in which three volumes (0.2ml, 0.1 ml, and 0.05 ml) of stock Na2S2O3 solution (20.5 g/L)were added to 10-ml volumes of pore water at five dilutionsand to an APW control. In one of the matrices, before theintroduction of Na2S2O3, the pore water samples were treatedwith 2 ml of deionized water saturated with SO2 gas to reduceoxidants and allow thiosulfate to chelate metals that might becausing toxicity.

To test whether large changes in pH influenced the toxicityin the sample, separate aliquots were changed to pH 3 and pH11 with dropwise additions of 1.0, 0.1, and 0.01 N HCl andNaOH. The aliquots were allowed to stand for at least 2 h andthen adjusted back to the ambient pH of the sample. As acontrol, an aliquot at ambient pH was allowed to stand for thesame amount of time as the other aliquots.

To separate toxicants associated with filterable materialsfrom the pore water, subsamples of the pH 3, pH 11, andambient aliquots were filtered using positive pressure through1.0-mm glass fiber filters.

Subsamples of the pH 3, pH 11, and ambient aliquots wereaerated for 1 h to determine whether volatile or oxidizablecompounds were causing toxicity. The pH of the samples waschecked after 30 min and adjusted if it had drifted. Care wastaken in removing the samples from the aeration vessels toprevent any sublated compounds present on the upper wallsof the containers from redissolving in the solutions.

Phase II: Identification. On three occasions (November1994, March 1995, and July 1995), pore water was extracted,preserved with HNO3, and analyzed for total metals (by in-ductively coupled plasma spectroscopy, methods 200.7 [23]and 6010 [24]) and total As (by graphite furnace atomic ad-sorption spectroscopy, methods 206.2 [23], 3020, and 7060[24]). Samples were kept at 48C in the dark until analyses wereperformed. Suspect metal toxicants were identified by com-


Fig. 2. Phase I toxicity identification evaluation results with site OMB-7 pore water. (a) EDTA addition; (b) pH change; (c) pH adjustment; (d)pH adjustment with EDTA addition; (e) Na2S2O3 with SO2 addition. ‘‘Baseline’’ indicates results of toxicity tests performed on unmanipulatedpore water. Results of each manipulation are compared to its corresponding baseline test. In (b), pH i refers to the ambient pH of the pore water.Bars represent 6 one standard error. § 5 standard error not calculable, §§ 5 LC50 is estimated.

paring metal concentrations in site pore water to U.S. EPAwater quality criteria [25], reported LC50 values, and LC50values determined in the laboratory in APW.

Toxic units were used to express the toxicity of site OMB-7 pore water and the toxicity of each suspected toxicant. Wholepore-water toxic units were calculated by dividing 100% bythe 48-h LC50 value for the pore water. Toxic units for eachsuspect metal were calculated by dividing the concentrationdetected in site OMB-7 pore water by the laboratory LC50value. Comparisons between the toxic units for each metal andthe whole pore-water toxic units were made to determine whatportion of the whole pore-water toxicity could be attributedto each metal.

Ethylenediaminetetraacetate ligand strongly chelates di-valent metals on a stoichiometric basis (1 mol EDTA to 1 molmetal [10]). The concentrations of the suspect metals foundin site OMB-7 pore water were compared to the amount ofEDTA needed to remove toxicity to help determine whichmetal or metals might have contributed most to the toxicityof the pore water.

Phase III: Confirmation. Four confirmation experimentswere conducted with pore water from a reference site(448349510N, longitude 738179050W) outside Outer MallettsBay and APW spiked with the suspect toxicants. In experiment1, the suspect toxicants were spiked at ambient levels intoseparate aliquots of APW and reference site pore water. Acutetoxicity tests were performed to determine which suspects in-duced toxicity. The LC50 values were calculated and comparedto the LC50 values in site OMB-7 pore water determined inphase I. In experiment 2, the suspect toxicants were spikedinto separate aliquots of site OMB-7 pore water to achievetwice the ambient levels of each toxicant. Toxicity tests wereperformed to determine how the toxicity of the samples re-sponded to doubling the concentrations of the suspected toxicagents. In experiment 3, pore water from site OMB-7 wasdiluted to 50% with APW. The suspect toxicants were added

back individually to aliquots of this mixture to achieve 100%of the concentration of each toxicant found in undiluted porewater. Toxicity tests were performed to determine which spikesreturned the level of toxicity in the diluted samples to thatfound in undiluted pore water. In experiment 4, phase I ma-nipulations that removed toxicity from site OMB-7 pore waterwere performed on spiked APW and reference site pore waterto see which of the suspected toxicants would exhibit the samepatterns of response as site OMB-7 pore water. These phaseI manipulations included pH changes to 3 and 11, addition ofEDTA, addition of Na2S2O3 and SO2, adjustment of pH from6.5 to 8.5, and combined pH adjustment and EDTA addition.The LC50 values were used to compare the patterns of re-sponse of each toxicant to each other toxicant and to the resultsof the phase I manipulations performed on site OMB-7 porewater.

RESULTS AND DISCUSSION

Initial toxicity

In initial toxicity tests, pore water from site OMB-7 sedi-ment was found to be acutely toxic to C. dubia. The LC50values were similar in the November and July samples: 45%pore water on November 14, 1994, and 43% pore water onJuly 26, 1995. No mortality occurred in artificial pore wateror moderately hard reconstituted water controls associated withthese initial toxicity screenings.

Phase I: Characterization

Toxicity was altered by the following five phase I manip-ulations (Fig. 2): adding 0.1 ml or more of EDTA completelyremoved toxicity; changing the pH of the pore water to 11and back to ambient pH also completely removed toxicity;toxicity was reduced when the pH of the pore water was main-tained at 6.5 but not affected by keeping the pH at 7.5 or 8.5;toxicity was completely removed when EDTA was added to


Table 2. Results of analyses of site OMB-7 pore water for total metalsand total arsenic

Element

Sampling Date

November 7, 1994a

(mg/L)March 11, 1995b

(mg/L)July 26, 1995c

(mg/L)

AlAsBaBeCdCrCoCuFePbMnNiSeAgVZn

5,000260290,5,518

,50,25

33,00037

26,000,40133

,10,50

47

2,900198352,2,5

615

28,700,50

38,900,15

,100,10,10

15

2,380135235,1,5

,5014

23,600,50

23,30016

,100,10,10

28

a Extracted on November 15, 1994; analyzed on December 7, 1994.b Extracted on March 15, 1995; analyzed on April 14, 1995.c Extracted on July 26, 1995; analyzed on July 29, 1995.

Table 3. The median lethal concentration (LC50) for 48-h exposurefor suspect toxicants with Ceriodaphnia dubia determined under

laboratory conditions (Lab) and reported in the literature (Lit)

Toxicant

Lab

(mg/L) 95% CI

Lit

(mg/L) Reference

As(III)As(V)NiMn

1,5402,540

2309,100

1,380–1,720a

180–2908,250–10,030

1,448b

1312,700

[17]

[27]c

a The 95% confidence interval could not be calculated because of lackof partial mortality in test dilutions.

b No values reported in the literature.c J.R. Hockett, personal communication.

test cups in which the pH was maintained at 6.5, 7.5, and 8.5,respectively; and adding Na2S2O3 and SO2-saturated water tothe pore water reduced toxicity. The water quality character-istics of unmanipulated pore water remained fairly constantthroughout the phase I testing period.

The removal of toxicity by EDTA implicates metallic cat-ions as the cause of toxicity in site OMB-7 pore water [11].The removal of toxicity as a result of the pH change to 11does not implicate a specific class of compounds; however, itdoes reveal a characteristic of the agent responsible for toxicityat the site. Any contaminant confirmed as being a toxic agentmust necessarily respond in a similar manner to changes inpH. Decreased toxicity at pH 6.5 in comparison to toxicity atpH 8.5 is a characteristic of ammonia and some metals [7].Ammonia is found in two forms over the range of the pHchange test. Ionized ammonia ( ) is the predominant spe-1NH4

cies at lower pHs, whereas the concentration of unionizedammonia (NH3) increases as pH increases [26]. Ionized am-monia is less toxic to C. dubia than unionized ammonia. Thistrend in pH-dependent toxicity is also exhibited by three di-valent metals, Zn21, Ni21, and Cd21 [27,28].

The pH adjustment test was performed with the additionof EDTA to determine whether the toxicant exhibiting thedecreased toxicity at pH 6.5 was a metal, a group of metals,or ammonia. The addition of EDTA to the pH adjustment testresulted in complete survival at each pH (Fig. 2), suggestingthat the toxic agent or agents are metals, not ammonia. Theconcentrations of total ammonia in the pore water during thecourse of each of these tests (EDTA alone, pH adjustmentalone, EDTA and pH adjustment in combination) were between4.0 and 5.5 mg/L. These concentrations are not sufficientlyhigh to cause acute toxicity to C. dubia according to publishedtoxicity studies [7,15] and are well below the LC50 for C.dubia determined in this laboratory (35 mg/L). Therefore, theresults of the combined test discounted ammonia as a suspectedtoxicant and added another piece of evidence supporting ametal as the toxic agent. The reduction in toxicity when sodiumthiosulfate with SO2 gas was introduced into the pore wateralso suggested a metal as the toxic agent.

Because no reductions in toxicity were found with octa-decyl sorbent, sodium thiosulfate alone, filtration or aeration,nonpolar organic compounds, oxidants such as chlorine, par-ticulate matter, and volatile or oxidizable compounds were alldiscounted as potential toxicants.

Phase II: Identification

The concentrations of metals detected in site OMB-7 porewater (Table 2) were generally consistent across the separateanalyses. The only exceptions were that approximately twiceas much Al and Zn were found in the November pore watersamples than in the other two samples.

No metals exceeded the acute criteria for the protection ofaquatic life established by the U.S. EPA [25]. Total As ex-ceeded the criterion for chronic toxicity (190 mg/L) in two outof three analyses. Lead was found to exceed the chronic cri-terion (3 mg/L) in the November sample. Manganese standsout in these analyses as being present at high levels. However,neither acute nor chronic criteria have been established forMn.

Manganese and Ni in pore water exceeded LC50 valuesreported for C. dubia in the literature. Nickel exceeded theacute LC50 value (13 mg/L at pH 8–8.5, hardness 5 300 mgCaCO3/L [28]) in the July sample. Manganese exceeded the

LC50 value (12,700 mg/L, hardness 5 50 mg CaCO3/L; J.R.Hockett, personal communication) in all analyses. Arsenic canexist in two oxidation states under the conditions typicallyfound in sediment pore water. Byrd (unpublished data) iden-tified both species in site OMB-7 pore water and found thatAs(III) was about 50% of the total As concentration, whereasthe other 50% was As(V). Therefore, we assumed that the totalAs concentrations were equally divided between As(III) andAs(V). Estimated As(III) concentrations in the pore water didnot exceed the published LC50 value (1,448 mg/L [17]).

The acute LC50 values for As(III), As(V), Mn, and Ni inAPW were determined at a hardness of 75 to 100 mg CaCO3/Land a pH of 8 to 8.3. The LC50s for 48-h exposure in thelaboratory and those reported in the literature are listed in Table3. The laboratory LC50 for As(III) was slightly lower thanthat for As(V). This is consistent with reports that As(III) ismore toxic than As(V) [17,29]. The laboratory LC50 for Mnwas comparable to the reported value (J.R. Hockett, personalcommunication). The LC50 value determined for Ni in APWwas an order of magnitude higher than the value reported invery hard reconstituted water at the same pH [28].

Three factors could have contributed to the difference intoxicity of Ni to C. dubia in reconstituted water and APW:pH, hardness, and the organic carbon content of the water. The


Table 4. The concentrations of potential toxicants detected in site OMB-7 pore water on the threesampling dates, toxicity attributed to each potential toxicant in toxic units (TU), and TU calculated for

site OMB-7 whole pore water

Toxicant

November 1994

(mg/L) TUa

March 1995

(mg/L) TUa

July 1995

(mg/L) TUa

MnNiTotal Asb

26,000,40260

2.9,0.17

38,900,15198

4.3,0.066

23,30016

135

2.60.071

As(III)As(V)

Total TUc

Whole pore-water TUd

0.170.103.12.2

0.1290.0784.52.9

0.08770.05312.72.3

a TU 5 concentration of metal detected in pore water/laboratory LC50 value for that metal.b Assuming As(III) and As(V) each account for one half of the total As concentration.c Total TU 5 sum of TU attributed to each toxicant in the pore water.d Whole pore-water TU 5 100%/whole pore-water LC50 value for site OMB-7.

Table 5. Comparison of whole pore-water toxic units and toxic units (TU) theoretically attributed toMn based on total Mn detected in site OMB-7 pore water on three dates (calculations following

Schubauer-Berigan et al. [10])

Datecollected

Mn(mg/L)

Wholepore-water

TUTheoretical

Mn TUaNon-Mn

TUbPredicted

Mn LC50c

November 1994March 1995July 1995

26,00038,90023,300

2.22.92.3

2.94.32.6

20.721.420.3

11.8211.7910.01

a Toxic units associated with Mn, assuming complete bioavailability. Theoretical Mn TU 5 [total Mn]/dilution water LC50.

b Non-Mn TU 5 whole pore-water TU 2 theoretical Mn TU.c Predicted Mn LC50 5 [total Mn]/whole pore-water TU (mg/L).

pH was between 8 and 8.5 in both the published tests in re-constituted water [28] and our tests in APW; therefore, pHcannot be responsible for the difference in toxicity of Ni. Thepublished LC50 value for Ni in reconstituted water was de-termined in very hard water [28], whereas our value was de-termined in softer APW, meaning that Ni toxicity was greaterin harder water. This observation contradicts what would beexpected from our understanding of the effects of water hard-ness on Ni toxicity. Generally, Ni toxicity is reduced as hard-ness increases [30]. The third factor is the organic carboncontent of the water. It may be a stronger influence than hard-ness on the toxicity of Ni. The published LC50 was derivedin water with no DOC [28]. When humic acid was present inour tests with APW, the toxicity of Ni was reduced. The large,amorphous organic carbon molecules comprising humic acidsmay have chelated some of the Ni introduced into APW, there-by limiting the availability of the metal to test organisms [31].

Using the laboratory and literature LC50 values as well aswater quality criteria, three possible toxicants were identified.Manganese concentrations in the site pore water exceeded boththe laboratory and literature LC50 values. Nickel concentra-tions in the site pore water were an order of magnitude lowerthan the APW LC50 value but were slightly above the literatureLC50 value. Arsenic remained on the list of suspects eventhough the individual As species were present at concentrationsbelow laboratory LC50 values because none of the phase Imanipulations directly altered or inhibited As toxicity [11] inthe pore water. Thus, As, Mn, and Ni were investigated in allTIE confirmation work.

An analysis of toxic units associated with whole pore waterand each suspect toxicant (Table 4) suggests that As and Ni

contributed negligibly to the toxicity observed in site OMB-7 pore water on the three occasions that their concentrationwas measured. The toxic units for Mn exceeded those for thewhole pore water, suggesting that Mn was the major contrib-utor to toxicity in the pore water.

The difference between the toxic units theoretically attri-buted to Mn and the whole pore-water toxic units is estimatedto be between 20.3 and 21.4 toxic units (Table 5). Becausethe difference is small, Mn can be assumed to account for alltoxicity observed in the pore water [10]. The LC50 value forMn in the pore water was calculated by dividing the measuredMn concentration in pore water by the whole pore-water toxicunits (Table 5). The predicted LC50 values for Mn for eachof the three dates using this approach were similar to both thelaboratory Mn LC50 value (9,100 mg/L) and the reported LC50value (12,700 mg/L). These results are further evidence im-plicating Mn as the toxic agent at site OMB-7.

Comparing the concentrations of EDTA added during twophase I experiments with the pore-water concentrations of Mnand Ni added more evidence implicating Mn. The amounts ofEDTA used were 6.2 mmol, 3.1 mmol, and 1.6 mmol in 10 mlof pore water. The two higher additions resulted in the removalof toxicity from site OMB-7 pore water. The amount of Ni in10 ml of the pore water in July was 0.0027 mmol. The amountof the lowest EDTA addition was three orders of magnitudegreater than that necessary to remove any toxicity resultingfrom Ni, making it extremely unlikely that Ni was the toxicagent. The concentrations of Mn in both tests fell between thetwo highest concentrations of EDTA added to the pore water(Table 6), but in all cases, the addition of 3.1 mmol EDTAwas sufficient to remove toxicity from the samples. Some Mn


Table 6. Comparison of the concentration of Mn in each test replicateto the calculated amount of EDTA remaining after the addition of 3.1

mmol of EDTA. The LC50 for Mn is 9,100 mg/L

EDTA treatment

Pretreatment[Mn]

(mmol)

Posttreatment[Mn]

(mmol) (mg/L)

November 1994July 1995Artificial pore water and

reference spikes

4.74.2

4.6

1.61.1

1.5

8,8006,000

8,200

Fig. 3. Phase III matching toxicity test results with site OMB-7 pore water. (a) As(III); (b) As(V); (c) Mn; (d) Ni. ‘‘Baseline’’ indicates resultsof toxicity tests performed on unmanipulated pore water. ‘‘100% add.’’ indicates the results of toxicity tests performed on pore water aliquotsin which suspect toxicant concentrations were doubled. ‘‘50% dil./50% add.’’ indicates dilution of pore water to 50% and addition of toxicantsback to the original concentration. Bars represent 6 one standard error. § 5 Standard error not calculable.

possibly was bound up by natural chelating agents such asorganic carbon in the samples; therefore, less EDTA was need-ed to remove toxicity than predicted by stoichiometric cal-culations. However, if all the Mn is assumed to be uncom-plexed and available, the concentration of Mn not chelated byEDTA is the difference between the total Mn concentrationand the concentration of EDTA added to the pore water. Thisdifference (Table 6), in all cases, was below the Mn LC50value (9,100 mg/L) determined with APW in this laboratory.Therefore, the amount of Mn left in the pore water, whetherit was complexed by other substances or not, was below thatnecessary to cause acute toxicity to the test organisms.

Phase III: Confirmation

Because a range of values for each metal was found in thethree analyses of pore water, we selected one value for thespiking experiments in phase III of the TIE. Two of the threeanalyses detected Mn in the mid-20,000 mg/L range and thethird at almost 40,000 mg/L. Manganese (as MnSO4) wasspiked at 25,000 mg/L to provide a conservative approximationof Mn contamination in the pore water but one that still ex-ceeded LC50 values. Nickel (as NiCl2) was spiked at 20 mg/L,

a concentration comparable to that detected in the July porewater sample and exceeding the literature acute LC50 value[28]. The total As concentration that was targeted was 200mg/L, a rough average of the concentrations detected in thethree analyses. On the basis of our assumption that As(III) andAs(V) were present in the pore water in equal amounts, spikingexperiments were conducted with these As species togetherand separately using 100 mg/L As(III) (from As2O3) and 100mg/L As(V) (from NaHAsO3).

In experiment 1, aliquots of APW spiked individually withambient concentrations of each of the suspect toxicants showedtoxicity only in the Mn treatments (mean 48-h LC50 556.27%, n 5 8). The Ni, As(III), As(V), and As(III) 1 As(V)aliquots resulted in no toxicity (100% survival after 48 h).The results of the reference pore-water spiking experimentswere similar to those of the APW spiking experiments; Mnwas the only toxicant that caused toxicity (mean 48-h LC505 56.87%, n 5 3). The level of toxicity induced by Mn inreference water and APW was somewhat lower than the meanLC50 of site OMB-7 pore water, which was 40.00% (n 5 8).However, because toxicity developed only in the Mn-spikedpore water, this experiment supported the identification of Mnas the agent of toxicity at site OMB-7.

In experiment 2, neither As(III) nor As(V) added separatelyto aliquots of site OMB-7 pore water at concentrations of 200mg/L significantly altered the baseline LC50 (Fig. 3). Adding25,000 mg/L of Mn to site OMB-7 pore water essentially dou-bled the toxicity of the sample. The Ni addition (20 mg/L)resulted in slightly increased toxicity, but the toxicity was notdoubled. These results suggested Mn as the principal agent oftoxicity but also that Ni may be a secondary contributor totoxicity.

In experiment 3, the baseline toxicity was not matchedwhen site OMB-7 pore water was diluted to one half of its


Fig. 4. Phase III results of toxicity tests performed on artificial pore water spiked with Mn (25,000 mg/L). (a) EDTA addition; (b) pH change;(c) pH adjustment; (d) Na2S2O3 with SO2 addition. In (b), pH i refers to the ambient pH of the pore water. Bars represent 6 one standard error.

original concentration and then As(III) and As(V) were addedback, separately, to match their original concentrations (Fig.3). Toxicity was removed in the As(III) aliquot. Some toxicitywas observed in the pore water spiked with As(V), but not atlevels that matched the original toxicity of the sample. Ad-ditions of Mn and Ni to separate aliquots of the 50% site OMB-7–50% APW mixture resulted in toxicities similar to those inthe unmanipulated pore water. These results suggested bothMn and Ni as toxic agents, but indicated that As is not likelyresponsible for the toxicity at the site.

In experiment 4, phase I manipulations were performed todetermine whether pH alterations or interactions with chelatingagents could affect the toxicity of the suspect toxicants to C.dubia. None of the manipulations resulted in any detectableincrease in toxicity, when using As(III), As(V), or Ni spikes.In contrast, phase I manipulations on Mn-spiked APW resultedin patterns of toxicant behavior similar to site OMB-7 porewater (Fig. 4). For example, EDTA completely removed tox-icity. Toxicity was also removed by changing the pH to 11but remained unchanged in the aliquot adjusted to pH 3. ThepH adjustment resulted in the removal of toxicity at pH 6.5and progressively greater toxicity at pH 7.5 and pH 8.5. Theaddition of Na2S2O3 and SO2-saturated water did not alter Mntoxicity. Toxicity was also induced by Mn spiked into referencepore water and the toxicity of Mn-spiked reference pore waterresponded in a similar manner to phase I manipulations as itdid to Mn-spiked APW (Fig. 5). All these results are consistentwith the identification of Mn as the toxic agent in site OMB-7 pore water.

CONCLUSIONS

TIE methods

A number of conclusions about TIE procedures based onthis application can be made. First, using artificial pore water

as a dilution water worked well in this series of experiments.The behavior of potential toxicants in pore water extractedfrom a reference site was similar to their behavior in artificialpore water. Matching as many water quality characteristics ofsite pore water as possible increases the chance that toxicantbehavior is influenced primarily by each TIE manipulation andnot by physical and chemical differences between the dilutionand test site waters. Additionally, by using APW, TIE exper-iments are easier to conduct than if pore water is extractedfrom a reference site and used as dilution water or in spikingexperiments in phase III. The extraction of pore water fromsediments is time consuming and labor intensive. Artificialpore water is easy to prepare and, therefore, is available inlarge quantities. Because it is very difficult to find referencesites where the pore water can be verified as uncontaminated,the use of APW both expedites the performance of the TIEexperiments and ensures a source of uncontaminated dilutionwater.

Second, performing chemical analyses for metals in thesediment pore water on several separate dates helps to providerepresentative toxicant concentrations. Concentrations of met-als and As varied in the pore water throughout the samplingand testing period. Having a number of data points to compareallowed us to generate representative values for the contam-inant concentrations based on the magnitude of the fluctuationswe observed.

Third, evidence from these experiments supports the con-cept that EDTA binds metals in a 1:1 molar ratio [10]. Toxicitywas not mitigated in site OMB-7 pore water until the EDTAconcentration was sufficiently high to reduce the Mn concen-tration to a level below that toxic to C. dubia. This conclusionassumes that the remaining Mn was bioavailable. Alternative-ly, some of the Mn may have been complexed by inorganic


Fig. 5. Phase III results of toxicity tests performed on reference sediment pore water spiked with Mn (25,000 mg/L). (a) EDTA addition; (b)pH change; (c) pH adjustment; (d) pH adjustment with EDTA. In (b), pH i refers to the ambient pH of the pore water. Bars represent 6 onestandard error. § 5 standard error not calculable, §§ 5 LC50 is estimated.

or organic carbon ligands present in the pore water. The re-maining toxic portion of Mn was available for complexationby EDTA.

TIE findings and implications

The evidence collected in this TIE overwhelmingly impli-cated Mn as the primary agent of toxicity in site OMB-7 porewater. Phase I results implicated divalent metals as the classof compounds responsible for toxicity. Phase II revealed con-centrations of Mn that could be harmful to aquatic life. PhaseIII experiments verified that ambient concentrations of Mn insite OMB-7 pore water were acutely toxic to C. dubia underlaboratory conditions. Toxic unit calculations indicated thatMn concentrations could account for the majority of the tox-icity observed in site OMB-7 pore water. The concentrationsof the other contaminants were not sufficiently high to causetoxicity in spiking experiments performed in the laboratory,whereas Mn spiking proportionally increased the toxicity ofthe sample. The response of APW and reference pore waterspiked with Mn to phase I manipulations was similar to thatobserved when these manipulations were performed on siteOMB-7 pore water during phase I.

Rarely has Mn been determined to be a toxic agent in lakesediments, even though Mn is often a major component offreshwater sediments. The situation in Malletts Bay may beunique. The bedrock in the watershed draining into MallettsBay is enriched with Mn and sediments containing Mn arecarried by the Lamoille River into the bay [32]. Causewaysacross the openings of the bay (Fig. 1) prevent the water ofthe bay from mixing with water in the main lake [33]; there-fore, the sediments settle to the bottom and are focused intothe deepest part of the bay [4]. Thermal stratification regularlyoccurs in the bay in late summer or fall; as a result, oxygen

becomes depleted in the hypolimnion [4]. This most likelyresults in reductive dissolution of manganese oxyhydroxides,which comprise a large proportion of the Mn in the sediments[34,35]. Manganese has been detected in the bottom water andsediment pore water at high levels as stratification and anoxiadevelop [4]. Manganese ions most likely remain in the sedi-ment pore water after turnover because of a lag in the time ittakes oxygen to diffuse from the water column into the sed-iments and the relatively slow oxidation kinetics of Mn(II)[34]. Therefore, Mn levels remain sufficiently high in surfacesediments to cause acute toxicity to C. dubia at all times ofthe year.

The results of this TIE revealed that Mn is responsible forthe acute toxicity in site OMB-7 sediments in the laboratory.The extent to which the Mn contamination at this site con-tributes to toxicity in the water column or has an effect on theecology of the bay is not clear, especially because pore-watertoxicity tests tend to provide conservative estimates of in situtoxicity. At other sites, investigators have found that com-pounds responsible for toxicity in pore water do not alwayscause toxicity in the overlying water column [16,36]; but thesituation in Outer Malletts Bay is undoubtedly complex. Theimplication of Mn as a toxic agent at this site underscores theneed to look broadly for potential contaminants in any hazardassessment.

Acknowledgement—This research was supported by a grant from theU.S. EPA through Lake Champlain Basin Program grant X001830-01-0. We would like to thank James Byrd for assistance with chemicalanalyses for arsenic; Michael Cunningham, Mary Kelly, StephanieMcCusker, and Kevin Twombly for assistance in the laboratory; ErikBrown for assistance in preparing figures; and Melosira Captain Ri-chard Furbush and Allen Cook for assistance in the field.


REFERENCES

1. Baudo R, Muntau H. 1990. Lesser known in-place pollutants anddiffuse source problems. In Baudo R, Giesy JP, Muntau H, eds,Sediments: Chemistry and Toxicity of In-Place Pollutants. CRC,Boca Raton, FL, USA, pp 1–14.

2. Lee GF, Jones RA. 1987. Water quality significance of contam-inants associated with sediments: An overview. In Dickson KL,Maki AW, Brungs WA, eds, Fate and Effects of Sediment-BoundChemicals in Aquatic Systems. Pergamon, New York, NY, USA,pp 1–60.

3. Burgess RM, Scott KJ. 1992. The significance of in-place con-taminated marine sediments on the water column: Processes andeffects. In Burton GA, ed, Sediment Toxicity Assessment. Lewis,Chelsea, MI, USA, pp 129–165.

4. King JW, et al. 1997. Physical and chemical characterization ofMalletts Bay. In McIntosh AW, Watzin MC, eds, Lake ChamplainSediment Toxics Assessment Program, Phase II Report. U.S.Environmental Protection Agency, Boston, MA.

5. Persaud D, Jaagumagi R. 1993. Guidelines for the protection andmanagement of aquatic sediment quality. Final/Technical Report.Ontario Ministry of the Environment, Ottawa, ON, Canada.

6. Smith SL, MacDonald DD, Keenleyside KA, Gaudet CL. 1999.The development and implementation of Canadian sediment qual-ity guidelines. J Aquat Ecosyst Health (in press).

7. Ankley GT, Katko A, Arthur JW. 1990. Identification of ammoniaas an important sediment-associated toxicant in the Lower FoxRiver and Green Bay, Wisconsin. Environ Toxicol Chem 9:313–322.

8. Hoke RA, Giesy JP, Kreis RG Jr. 1992. Sediment pore watertoxicity identification in the Lower Fox River and Green Bay,Wisconsin, using the Microtox assay. Ecotoxicol Environ Saf 23:343–354.

9. Norberg-King TJ, Durhan EJ, Ankley GT, Robert E. 1991. Ap-plication of toxicity identification evaluation procedures to theambient waters of the Colusa Basin Drain, California. EnvironToxicol Chem 10:891–900.

10. Schubauer-Berigan MK, et al. 1993. The behavior and identifi-cation of toxic metals in complex mixtures: Examples from ef-fluent and sediment pore water toxicity identification evaluations.Arch Environ Contam Toxicol 24:298–306.

11. U.S. Environmental Protection Agency. 1991. Methods for aquat-ic toxicity identification evaluations: Phase I toxicity character-ization procedures, 2nd ed. EPA/600/6-91/003. Washington, DC.

12. Durhan EJ, Norberg-King TJ, Burkhard LP, Ankley GT, Luka-sewycz MT, Schubauer-Berigan MK, Thompson JA. 1993. Meth-ods for aquatic toxicity identification evaluations: Phase II tox-icity identification procedures for samples exhibiting acute andchronic toxicity. EPA/600/R-92/080. U.S. Environmental Protec-tion Agency, Washington, DC.

13. Mount DI, Norbert-King TJ, Ankley GT, Burkhard LP, DurhanEJ, Schubauer-Berigan MK, Thompson JA. 1993. Methods foraquatic toxicity identification evaluations: Phase III toxicity con-firmation procedures for samples exhibiting acute and chronictoxicity. EPA/600/R-92/081. U.S. Environmental ProtectionAgency, Washington, DC.

14. Giesy JP, Hoke RA. 1989. Freshwater sediment toxicity bioas-sessment: Rationale for species selection and test design J GreatLakes Res 15:539–569.

15. Schubauer-Berigan MK, Ankley GT. 1991. The contribution ofammonia, metals and nonpolar organic compounds to the toxicityof sediment interstitial water from an Illinois River tributary.Environ Toxicol Chem 10:925–939.

16. Ankley GT, Schubauer-Berigan MK, Dierkes JR. 1991. Predicting

the toxicity of bulk sediments to aquatic organisms with aqueoustest fractions: Pore water vs. elutriate. Environ Toxicol Chem 10:1359–1366.

17. Spehar RL, Fiandt JT, Anderson RL, DeFoe DL. 1980. Compar-ative toxicity of arsenic compounds and their accumulation ininvertebrates and fish. Arch Environ Contam Toxicol 9:53–63.

18. Thurman EM. 1985. Organic Geochemistry of Natural Waters.Martinus Nijhoff/Dr. W. Junk, Dordrecht, Germany.

19. Witters HE, Van Puymbroeck S, Vangenechten JHD, Vander-borght OCJ. 1990. The effect of humic substances on the toxicityof aluminum to adult rainbow trout, Oncorhynchus mykiss (Wal-baum). J Fish Biol 37:43–53.

20. U.S. Environmental Protection Agency. 1991. Methods for mea-suring the acute toxicity of effluent and receiving water to fresh-water and marine organisms, 4th ed. EPA/600/4-90/027. Cincin-nati, OH.

21. Hamilton MA, Russo RC, Thurston RV. 1977. Trimmed Spear-man–Karber method for estimating median lethal concentrationsin toxicity bioassays. Environ Sci Technol 11:714–719.

22. Hockett JR, Mount DR. 1996. Use of metal chelating agents todifferentiate among sources of acute aquatic toxicity. EnvironToxicol Chem 15:1687–1693.

23. U.S. Environmental Protection Agency. 1982. Methods for chem-ical analysis of water and wastes. EPA-600/4/79-020. Cincinnati,OH.

24. U.S. Environmental Protection Agency. 1984. Test methods forevaluating solid waste, 2nd ed. SW-846. Washington, DC.

25. U.S. Environmental Protection Agency. 1987. Quality criteria forwater. EPA/440/5-86/001. Washington, DC.

26. Emerson K, Russo RC, Lund RE, Thurston RV. 1975. Aqueousammonia equilibrium calculations: Effect of pH and temperature.J Fish Res Board Can 32:2379–2383.

27. Campbell PGC, Stokes PM. 1985. Acidification and toxicity ofmetals to aquatic biota. Can J Fish Aquat Sci 42:2034–2049.

28. Schubauer-Berigan MK, Dierkes JR, Monson PD, Ankley GT.1993. pH-dependent toxicity of Cd, Cu, Ni, Pb and Zn to Cer-iodaphnia dubia, Pimephales promelas, Hyalella azteca andLumbriculus variegatus. Environ Toxicol Chem 12:1261–1266.

29. Ferguson JF, Gavis J. 1972. A review of the arsenic cycle innatural waters. Water Res 6:1259–1274.

30. Kszos LA, Stewart AJ, Taylor PA. 1992. An evaluation of nickeltoxicity to Ceriodaphnia dubia and Daphnia magna in a con-taminated stream and in laboratory tests. Environ Toxicol Chem11:1001–1012.

31. Fjeld E, Rognerud S, Steinnes E. 1994. Influence of environ-mental factors on heavy metal concentration in lake sediments insouthern Norway indicated by path analysis. Can J Fish AquatSci 51:1708–1720.

32. Schuck R. 1995. An historical record of arsenic contaminationin the sediments of Arrowhead Mountain Lake, Milton, Vermont.MS thesis. University of Vermont, Burlington, VT, USA.

33. Meyer GE, Gruendling GK. 1979. Limnology of Lake Champlain.Lake Champlain Basin Study Technical Report 30. New EnglandRiver Basins Commission, Burlington, VT, USA.

34. Davison W. 1993. Iron and manganese in lakes. Earth-Sci Rev34:119–163.

35. Balistrieri LS, Murray JW, Paul B. 1992. Biogeochemical cyclingof trace metals in the water column of Lake Sammamish, Wash-ington: Response to seasonally anoxic conditions. Limnol Ocean-ogr 37:529–548.

36. Burgess RM, Schweitzer KA, McKinney RA, Phelps DK. 1993.Contaminated marine sediments: Water column and interstitialtoxic effects. Environ Toxicol Chem 12:127–138.

toxicity identification evaluation of metal-contaminated sediments using an artificial pore water...

Documents