DDT, Chlordane, Toxaphene and PCB Residues in Newport Bay and Watershed: Assessment of Hazard to Wildlife and Human Health - Reviews of Environmental Contamination and Toxicology - page 143

values used are outdated, because they are not based on the superior slow-stir
technique (de Bruijn et al. 1989 ). Second, the high variability in the outdated K oc
values is not seen in those derived by the slow-stir method, precluding the necessity
of using the lower 95th and 99th percentile of the K oc values. Multiplying the K oc
value for chlordane published by US EPA Region IX ( 2002 ) and the SARWQCB
staff (SARWQCB 2006 ) by the chronic marine CTR standard gives a single data
point of 65 ppb chlordane in sediment containing 1% OC.
The last two data points in Table 30 are based on spiked sediment bioassays
(McLeese and Metcalfe 1980 ; McLeese et al. 1982 ). These two bioassays are used to
assess toxicity primarily from the water column rather than from sediment. The fi rst
study was done with sand shrimp ( Crangon septemspinosa ) and involved adding an
unreported amount of chlordane to a beaker, drying off the solvent, adding water and
coarse sand (0.28% OC; 0.5-2 mm diameter particles). The sand was allowed to
settle and the shrimp were added. The authors concluded that chlordane dissolved in
the water phase was the primary cause of toxicity. Chlordane bound to sediments
contributed little to toxicity. For these reasons, and the fact that the chlordane moved
from water to sediment, this bioassay is primarily a water bioassay. The same is true
of the second study (McLeese et al. 1982 ). The difference between the two studies is
that in the second study the organism was a polychaete worm ( Nereis virens ) and the
sediment was sandy silt that contained 2% OC. In a true sediment bioassay, all of the
chlordane would be picked up off of the glass by the sediment. The sediment would
then be transferred to a clean container and equilibrated with water. Samples of water
and sediment would be analyzed periodically until an equilibrium was reached.
The test organism would be added only after equilibrium was achieved.
Sediment LC 50 s for these two data points can be estimated using equilibrium
partitioning. If one applies the K oc published in the U.S. EPA Region IX ( 2002 ) and
SARWQCB staff reports ( 2006 ) and the water only LC 50 s reported by McLeese
et al. ( 1982 ), the estimated sediment LC 50 s are 9,000 ppb for sand shrimp and
7,100,000 ppb for the polychaete worm.
The remaining 8 data points in Table 33 from Long and Morgan ( 1990 ) are based on
the presence of chlordane (along with potentially hundreds of other chemicals) in toxic
sediments. None of these eight data points provide dose-response information.
The fl aws in the ERM data set preclude the use of the ERM value as an indication
of the threshold for benthic toxicity due to chlordane in sediments. The threshold
appears to be orders of magnitude greater than the ERM. This conclusion is further
supported by other bioassay data.

4.2.3 Equilibrium Partition Estimates of Toxicity Thresholds

Let us look further at what is known about toxicity thresholds for chlordane to
amphipods and other benthic organisms to gain an understanding of whether the
levels of chlordane in sediments in Newport Bay are high enough to cause toxicity
to these organisms. Cardwell et al. ( 1977 ) studied the chronic toxicity of chlordane
in the amphipod, Hyallela azteca (Table 31 ).