Standard Test Method for Measuring the Toxicity of Sediment-Associated Contaminants with Freshwater Invertebrates
|Publication Date:||1 December 2004|
|ICS Code (Microbiology of water):||07.100.20|
This test method covers procedures for testing freshwater organisms in the laboratory to evaluate the toxicity of contaminants associated with whole sediments. Sediments may be collected from the field or spiked with compounds in the laboratory.
Test methods are described for two toxicity test organisms, the amphipod Hyalella azteca ( H. azteca) (see 13.1.2) and the midge Chironomus dilutus (formerly known as C. tentans; Shobanov et al. 1999.(1) (see 14.1.2). The toxicity tests are conducted for 10 days in 300-mL chambers containing 100 mL of sediment and 175 mL of overlying water. Overlying water is renewed daily and test organisms are fed during the toxicity tests. Endpoints for the 10-day toxicity tests are survival and growth. These test methods describe procedures for testing freshwater sediments; however, estuarine sediments (up to 15 ppt salinity) can also be tested with H. azteca. In addition to the 10-day toxicity test method outlined in 13.1.2 and 14.1.2, general procedures are also described for conducting 10-day sediment toxicity tests with H. azteca (see 13.1.2) and C. dilutus (see 14.1.2).
Guidance for conducting sediment toxicity tests is outlined in Annex A1 for Chironomus riparius, in Annex A2 for Daphnia magna and Ceriodaphnia dubia, in Annex A3 for Hexagenia spp., in Annex A4 for Tubifex tubifex, and in Annex A5 for the Diporeia spp. Guidance is also provided in Annex A6 for conducting long-term sediment toxicity tests with H. azteca by measuring effects on survival, growth, and reproduction. Guidance is also provided in Annex A7 for conducting long-term sediment toxicity tests with C. dilutus by measuring effects on survival, growth, emergence, and reproduction. 1.6 outlines the data that will be needed before test methods are developed from the guidance outlined in Annex A1 to Annex A7 for these test organisms. General procedures described in Sections 17 for sediment testing with H. azteca and C. dilutus are also applicable for sediment testing with the test organisms described in Annex A1 to Annex A7.
Procedures outlined in this test method are based primarily on procedures described in the United States Environmental Protection Agency (USEPA) (2-9 )(Footnote 2), Test Method E 1367, and Guides E 1391, E 1525 and E 1688.
Additional research and methods development are now in progress to: (1) evaluate additional test organisms, (2) further evaluate the use of formulated sediment, (3) refine sediment dilution procedures, (4) refine sediment toxicity identification evaluation (TIE) procedures (10), (5) refine sediment spiking procedures, (6) develop in situ toxicity tests to assess sediment toxicity and bioaccumulation under field conditions, (7) evaluate relative sensitivities of endpoints measured in tests, (8) develop methods for new species, (9) evaluate relationships between toxicity and bioaccumulation, and (10) produce additional data on confirmation of responses in laboratory tests with natural populations of benthic organisms. Some issues that may be considered in interpretation of test results are the subject of continuing research including the influence of feeding on bioavailability, nutritional requirements of the test organisms, and additional performance criteria for organism health. See Section 6 for additional detail. This information will be described in future editions of this standard.
The USEPA (2) and Guide E 1688 also describes 28-day bioaccumulation methods for the oligochaete Lumbriculus variegatus.
Results of tests, even those with the same species, using procedures different from those described in the test method may not be comparable and using these different procedures may alter bioavailability. Comparison of results obtained using modified versions of these procedures might provide useful information concerning new concepts and procedures for conducting sediment tests with aquatic organisms. If tests are conducted with procedures different from those described in this test method, additional tests are required to determine comparability of results. General procedures described in this test method might be useful for conducting tests with other aquatic organisms; however, modifications may be necessary.
Selection of Toxicity Testing Organisms:
The choice of a test organism has a major influence on the relevance, success, and interpretation of a test. Furthermore, no one organism is best suited for all sediments. The following criteria were considered when selecting test organisms to be described in this standard (Table 1 and Guide E 1525). A test organism should: (1) have a toxicological data base demonstrating relative sensitivity and discrimination to a range of chemicals of concern in sediment, (2) have a database for interlaboratory comparisons of procedures (for example, round-robin studies), (3) be in contact with sediment [e.g., water column vs benthic organisms], (4) be readily available through culture or from field collection, (5) be easily maintained in the laboratory, (6) be easily identified, (7) be ecologically or economically important, (8) have a broad geographical distribution, be indigenous (either present or historical) to the site being evaluated, or have a niche similar to organisms of concern, (for example, similar feeding guild or behavior to the indigenous organisms), (9) be tolerant of a broad range of sediment physico-chemical characteristics (for example, grain size), and (10) be compatible with selected exposure methods and endpoints. The method should also be (11) peer reviewed and (12) confirmed with responses with natural populations of benthic organisms (see 1.6.8).
Of the criteria outlined in Table 1, a data base demonstrating relative sensitivity to contaminants, contact with sediment, ease of culture in the laboratory, interlaboratory comparisons, tolerance of varying sediment physico-chemical characteristics, and confirmation with responses of natural benthos populations were the primary criteria used for selecting H. azteca and C. dilutus to be described as test methods in the current version of this standard (see Sections 13 and 14). Procedures for conducting sediment tests with organisms in accordance with Annex A1 to Annex A7 do not currently meet all the required selection criteria listed in Table 1. A similar data base must be developed before these or other test organisms can be included as standard test methods instead of as guidance in future versions of these this method.
An important consideration in the selection of specific species for test method development is the existence of information concerning relative sensitivity of the organisms both to single chemicals and complex mixtures. A number of studies have evaluated the sensitivity of H. azteca, C. dilutus, and L. variegatus, relative to one another, as well as other commonly tested freshwater species. For example, Ankley et al (11) found H. azteca to be as, or slightly more, sensitive than Ceriodaphnia dubia to a variety of sediment elutriate and pore-water samples. In that study, L. variegatus were less sensitive to the samples than either the amphipod or the cladoceran. West et al (12) found the rank sensitivity of the three species to the lethal effects of copper in sediments from the Keweenaw Waterway, MI was (from greatest to least): H. azteca > C. dilutus > L. variegatus. In short-term (48 to 96 h) exposures, L. variegatus generally was less sensitive than H. azteca, C. dubia, or Pimephales promelas to cadmium, nickel, zinc, copper, and lead (13). Of the latter three species, no one species was consistently the most sensitive to the five metals.
In a study of contaminated Great Lakes sediment, H. azteca, C. dilutus, and C. riparius were among the most sensitive and discriminatory of 24 organisms tested (14-17). Kemble et al (18) found the rank sensitivity of four species to metal-contaminated sediments from the Clark Fork River, MT to be (from greatest to least): H. azteca > C. riparius > Oncorhynchus mykiss (rainbow trout) > Daphnia magna. Relative sensitivity of the three endpoints evaluated in the H. azteca test with Clark Fork River sediments was (from greatest to least): length > sexual maturation > survival.
In 10-day water-only and whole-sediment tests, Hyalella azteca and C. dilutus were more sensitive than D. magna to fluoranthene-spiked sediment (19).
Ten-day, water-only tests also have been conducted with a number of
chemicals using H. azteca, C.
dilutus, and L. variegatus ((19) and Table 2). These tests all were
flowthrough exposures using a
soft natural water (Lake Superior) with measured chemical
concentrations that, other than the
absence of sediment, were conducted under conditions (for example,
feeding) similar to those being described for the standard 10-day
sediment test in 13.1.2. In
general, H. azteca was more sensitive to copper, zinc, cadmium,
nickel, and lead than either C.
dilutus or L. variegatus. Chironomus dilutus and H. azteca exhibited a
similar sensitivity to
several of the pesticides tested. Lumbriculus variegatus was not
tested with several of the
pesticides; however, in other studies with whole sediments
Using the data from Table 2, sensitivity of H. azteca, C. dilutus, and
L. variegatus can be
evaluated relative to other freshwater species. For this analysis,
acute and chronic toxicity data
from water quality criteria (WQC) documents for copper, zinc, cadmium,
nickel, lead, DDT, dieldrin,
and chlorpyrifos, and toxicity information from the AQUIRE data base
Measurable concentrations of ammonia are common in the pore water of many sediments and have been found to be a common cause of toxicity in pore water (21 22, 23). Acute toxicity of ammonia to H. azteca, C. dilutus, and L. variegatus has been evaluated in several studies. As has been found for many other aquatic organisms, the toxicity of ammonia to C. dilutus and L. variegatus has been shown to be dependent on pH. Four-day LC50 values for L. variegatus in water-column (no sediment) exposures ranged from 390 to 6.6 mg/L total ammonia as pH was increased from 6.3 to 8.6 Schubauer-Berigan et al.(24). For C. dilutus, 4-day LC50 values ranged from 370 to 82 mg/L total ammonia over a similar pH range (Schubauer-Berigan et al.) (24). Ankley et al. (25) reported that the toxicity of ammonia to H. azteca (also in water-only exposures) showed differing degrees of pH-dependence in different test waters. In soft reconstituted water, toxicity was not pH dependent, with 4-day LC50 values of about 20 mg/L at pH ranging from 6.5 to 8.5. In contrast, ammonia toxicity in hard reconstituted water exhibited substantial pH dependence with LC50 values decreasing from >200 to 35 mg/L total ammonia over the same pH range. Borgmann and Borgmann (26) later showed that the variation in ammonia toxicity across these waters could be attributed to differences in sodium and potassium content, which appear to influence the toxicity of ammonia to H. azteca.
Although these studies provide benchmark concentrations that may be of concern in sediment pore waters, additional studies by Whiteman et al. (27) indicated that the relationship between water-only LC50 values and those measured in sediment exposures differs among organisms. In sediment exposures, the 10-day LC50 for L. variegatus and C. dilutus occurred when sediment pore water reached about 150 % of the LC50 determined from water-only exposures. However, experiments with H. azteca showed that the 10-day LC50 was not reached until pore water concentrations were nearly 10 the water-only LC50, at which time the ammonia concentration in the overlying water was equal to the water-only LC50. The authors attribute this discrepancy to avoidance of sediment by H. azteca. Thus, it appears that water-only LC50 values may provide suitable screening values for potential ammonia toxicity, higher concentrations may be necessary to actually induce ammonia toxicity in sediment exposures, particularly for H. azteca. Further, these data underscore the importance of measuring the pH of pore water when ammonia toxicity may be of concern. Ankley Schubauer-Bergian (28) and Besser et al. (29) describe procedures for conducting toxicity identification evaluations (TIEs) for pore-water or whole-sediment samples to determine if ammonia is contributing to the toxicity of sediment samples.
Relative species sensitivity frequently varies among chemicals; consequently, a battery of tests including organisms representing different trophic levels may be needed to assess sediment quality (14, 17, 30-33). For example, Reish (34) reported the relative toxicity of six metals (arsenic, cadmium, chromium, copper, mercury, and zinc) to crustaceans, polychaetes, pelecypods, and fishes and concluded that no one species or group of test organisms was the most sensitive to all of the metals.
Sensitivity of a species to chemicals is also dependent on the duration of the exposure and the endpoints evaluated. Annex A6 and Annex A7 describe results of studies which demonstrate the utility of measuring sublethal endpoints in sediment toxicity tests with the amphipod H. azteca and the midge C. dilutus.
The sensitivity of an organism to chemicals should be balanced with the concept of discrimination (14). The response of a test organism should provide discrimination between different levels of contamination. However, insensitive organisms may be preferred for determining bioaccumulation. The use of indigenous organisms that are ecologically important and easily collected is often very straightforward; however, indigenous organisms at a site may be insensitive to the chemicals of concern. Indigenous organisms might be more important for evaluation of bioaccumulation (9). See Guides E 1525, E 1688, and E 1850 for additional detail on selection of test organisms.
Sensitivity of an organism is related to route of exposure and biochemical sensitivity to chemicals. Sediment-dwelling organisms can receive a dose from three primary sources: interstitial water, sediment particles, and overlying water. Food type, feeding rate, assimilation efficiency, and clearance rate will control the dose of chemicals from sediment (Guide E 1688). Benthic invertebrates often selectively consume different particle sizes (35) or particles with higher organic carbon concentrations which may have higher chemical concentrations. Detrital feeders may receive most of their body burden directly from sediment ingestion. In amphipods (36) and clams (37) uptake through the gut can exceed uptake across the gills for certain hydrophobic compounds. Organisms in direct contact with sediment may also accumulate chemicals by direct adsorption to the body wall or by absorption through the integument (38).
Despite the potential complexities in estimating the dose that an animal receives from sediment, the toxicity and bioaccumulation of many chemicals in sediment such as chlordecone, fluoranthene, organochlorines, and metals have been correlated with either the concentration of these chemicals in interstitial water or in the case of nonionic organic chemicals, concentrations of an organic-carbon basis (39, 40). The relative importance of whole sediment and interstitial water routes of exposure depends on the test organism and the specific contaminant (35, 38). Because benthic communities contain a diversity of organisms, many combinations of exposure routes may be important. Therefore, behavior and feeding habits of a test organism can influence its ability to accumulate contaminants from sediment and should be considered when selecting test organisms for sediment testing.
The response of H. azteca and C. dilutus in laboratory toxicity studies has been compared to the response of natural populations of benthic organisms to potentially contaminated sediments.
Chironomids were not found in sediment samples that decreased the growth of C. dilutus by 30 % or more in 10-day laboratory toxicity tests (41). Wentsel et al (42-44) reported a correlation between effects on C. dilutus in laboratory tests and the abundance of C. dilutus in metal-contaminated sediments.
Canfield et al. (45,46,47) evaluated the composition of benthic invertebrate communities in sediments for the following areas: (1) three Great Lakes Areas of Concern (AOC; Buffalo River, NY: Indiana Harbor, IN: Saginaw River, MI), (2) the upper Mississippi River, and (3) the Clark Fork River located in Montana. Results of these benthic community assessments were compared to sediment chemistry and toxicity (28-day sediment exposures with H. azteca which monitored effects on survival, growth, and sexual maturation). Good concordance was evident between measures of laboratory toxicity, sediment contamination, and benthic invertebrate community composition in extremely contaminated samples. However, in moderately contaminated samples, less concordance was observed between the composition of the benthic community and either laboratory toxicity test results or sediment contaminant concentration. Laboratory sediment toxicity tests better identified chemical contamination in sediments compared to many of the commonly used measures of benthic invertebrate community composition. Benthic measures may reflect other factors such as habitat alteration in addition to responding to contaminants. Canfield et al. (45, 46, 47) identified the need to better evaluate non-contaminant factors (i.e., TOC, grain size, water depth, habitat alteration) in order to better interpret the response of benthic invertebrates to sediment contamination.
Results from laboratory sediment toxicity tests were compared to colonization of artificial substrates exposed in situ to Great Lakes sediment (14) Burton et al. (17) Survival or growth of H. azteca and C. dilutus in 10-28-day laboratory exposures were negatively correlated to percent chironomids and percent tolerant taxa colonizing artificial substrates in the field. Schlekat et al (48) reported general good agreement between sediment toxicity tests with H. azteca and benthic community responses in the Anacostia River in Washington, DC.
Sediment toxicity with amphipods in 10-day toxicity tests, field contamination, and field abundance of benthic amphipods were examined along a sediment contamination gradient of DDT (48). Survival of Eohaustorius estuarius, Rhepoxynius abronius, and H. azteca in laboratory toxicity tests was positively correlated to abundance of amphipods in the field and negatively correlated to DDT concentrations. The threshold for 10-day sediment toxicity in laboratory studies was about 300 µg DDT (+metabolites)/g organic carbon. The threshold for abundance of amphipods in the field was about 100 µg DDT (+metabolites)/g organic carbon. Therefore, correlations between toxicity, contamination, and field populations indicate that short-term sediment toxicity tests can provide reliable evidence of biologically adverse sediment contamination in the field, but may be underprotective of sublethal effects.
Limitations - While some safety considerations are included in this standard, it is beyond the scope of this standard to encompass all safety requirements necessary to conduct sediment tests.
This standard is arranged as follows:
This standard does not purport to address all of the safety concerns, if any, associated with its use. It is the responsibility of the user of this standard to establish appropriate safety and health practices and determine the applicability of regulatory limitations prior to use. Specific hazard statements are given in Section 8.
Footnote * - A Summary of Changes section appears at the end of this standard.
Footnote 2 - The boldface numbers in parentheses refer to the list of references at the end of this standard.