Behavior of Organic Contaminants in Soils, and Consequences for Ecotoxicity Tests

ERDC TN-DOER-R3

September 2004

The use of this grass species is limited to the southern part of North America as far north as

Kentucky. Earthworms did not exhibit metal toxicity in the range studied, and a correlation was

found between bioaccumulation and substrate Zn levels. Thus, earthworm bioaccumulation was

considered as a valid Tier III assay for Zn (Best et al. 2003).

Behavior of Organic Contaminants in Soils, and Consequences for Ecotoxicity

Tests: Organic compounds differ from metals with respect to their behavior in the soil (e.g.

organic contaminants are biodegradable, while metals are not) and toxic action. Their interaction

with plants differs from that of metals. Moreover, ecotoxicological research on organic contami-

nants is far scarcer than on metals. The bioavailability of organic contaminants can change sig-

nificantly over time due to complexation, degradation, or loss by volatilization, runoff, or

leaching.

Hydrophobic organic contaminants (HOC) in sediments/soils and dredged material can exist in

any of the following four states: solids (soil and organic matter particles), fluid (water), gas

(within the sediment/soil), and biota (Brusseau 1997). Partitioning of the contaminant over these

phases depends on the contaminant characteristics and the processes affecting the transitions

from one phase to the other. The most important phase transition processes are the solubilization

(transition of pure contaminant into the water phase), volatilization (transition from the water to

the gas phase), and sorption (binding to the solid phase; Beck et al. 1993). The most immobile

HOCs are hydrophobic (low solubility in water) and lipophilic. These contaminants are usually

strongly bound to organic material and sediment/soil particles (Tenner et al. 1997). These char-

acteristics decrease bioavailability and toxicity for plants (Shimp et al. 1993), soil macrofauna

(Belfroid et al. 1995), surface fauna, and microorganisms. Bioavailability of HOC decreases with

aging (Alexander 1995). Therefore, short-term laboratory experiments in which spiked substrates

are tested may lead to overestimates of the ecotoxicological risks of aged contaminants. The

aging process results in adsorption of HOC in micro- or nanopores and/or the soil matrix. This

means that strongly bound, non-extractable, or difficult to extract HOC residues can be present in

the soil long after HOC pollution took place. Although the concentrations of aged HOCs may be

high, ranging from 100 to 1000 mg/kg organic soil matter, desorption processes are extremely

slow, and, therefore, the relationship between total HOC concentration and ecological processes

in the same soil is difficult to establish (De Jonge 1996).

The organic carbon fraction of DM can play an important role in the bioavailability of HOCs

(Fredrickson et al. 2003). A large fraction of the HOC that is extractable from sediments with

organic solvents can be non-available to biota, and the proportion of the non-biologically avail-

able fraction to total extractable HOC levels can vary greatly from sediment to sediment. For

instance, Talley et al. (2002) showed that Soxhlet-extractable PAH concentrations in DM from a

CDF in Milwaukee Harbor averaged 115 mg kg-1, but only 46 mg kg-1, i.e. less than half, was

bioavailable. These results and others have led to questioning the validity of the equilibrium par-

titioning theory developed by DiToro et al. (1991), that describes bioaccumulation of HOCs

from sediments by a simple constant partitioning process between sediment and organic carbon

and biomass (Kraaij et al. 2002). The quality of organic carbon may also have a large effect on

the bioavailability of HOCs. For instance, in several studies the non-bioavailable fraction of

solvent-extractable HOCs was associated with pitch and coke-derived particles (Paine et al.

1996; Talley et al. 2002).