Review of Ecological Risk Assessment - Ashland Lakefront Site, Ashland, WI

• Summary
• Background
• Adherence to the U.S. Environmental Protection Agency's Guidelines for Ecological Risk Assessment
• Problem Formulation
  • Source of contaminants
  • Contaminant fate and transport
  • Conceptual model
  • Assessment endpoints
• Analysis
  • Exposure characterization
  • Ecological effects characterization
• Risk Characterization
  • Causality
• Probable Impacts on Organisms
  • Nature and intensity of effects
  • Spatial scale
  • Potential for natural recovery
• References
• PDF format of this entire document

By:
Christopher Marwood, Ph.D.
Post-doctoral Research Fellow
Department of Zoology
Miami University
Oxford, Ohio, U.S.A.

For:
Technical Outreach Services for Communities (TOSC)
Michigan State University.

Conducted at the request of Ashland/Bayfield County League of Women Voters and the Lake Superior Alliance

Date: May 14, 2001

Summary

This report is an evaluation of two ecological risk assessments (ERA) of contaminated sediments in Lake Superior offshore from Ashland, WI, performed by Short Elliott Hendrickson (SEH) for the Wisconsin Department of Natural Resources, and Dames & Moore (D&M) for Northern States Power. Both ERAs used the same data, collected for the SEH assessment, but arrived at different conclusions. The TOSC program was asked to provide an independent review of the SEH and D&M ERAs, with particular focus on the findings and methodology. In this report, the sources of discrepancies between the two ERAs are outlined. The merit of the conclusions offered by each assessment is examined in the context of the scientific literature pertaining to the contaminants and organisms at the Ashland site. The probable impacts on biota, considering the available data regarding the level of contamination in the sediment and water, are discussed.

The major differences and sources of disagreement between the SEH and D&M risk assessments are outlined below.

1. ERA style - The assessment prepared by SEH more closely adhered to the structure suggested by the U.S. EPA's Guidelines for Ecological Risk Assessment. In the D&M assessment, important sections of the analysis and risk characterization phases were missing.
2. Measurement endpoints - Different models were used to compare sediment PAH concentrations with expected ecological impacts. SEH used Sediment Effects Concentrations derived from Hyalella azteca 28-day chronic exposures. D&M used an equilibrium-partitioning model to predict pore water PAH concentrations, and then a quantitative structure activity relationship (QSAR) to estimate an EC25, a concentration they claim is protective of populations.
3. Analysis - SEH characterized exposure by comparing individual sediment PAH concentrations to Sediment Effects Concentrations, then calculated a toxic unit from the summed values. D&M attempted to calculate a hazard quotient based on average sediment concentrations divided by the estimated EC25; however, D&M made several deviations from the model they used, as well as calculation errors. The D&M calculations, when performed as specified by the model from which they were derived, yield a result very similar to that of the SEH analysis.
4. Risk characterization - SEH used a triad sediment approach to evaluate sediment PAH concentrations, a benthic community survey, and laboratory sediment exposures. D&M considered the results of the benthic survey and laboratory exposures inconclusive, and used only sediment PAH concentrations in its risk characterization.
5. Phototoxicity - Because PAHs are known to exhibit enhanced toxicity in sunlight, phototoxicity experiments using contaminated sediment were conducted for the SEH assessment. D&M did not use these results in its assessment.
6. Conclusions - SEH concluded a high risk of detrimental ecological effects to the benthic communities in much of the sediment adjacent to the Ashland waterfront. D&M disputed the SEH conclusions, and predicted a smaller area of contamination requiring remediation.

Based on our evaluation of the two ERAs, it is the opinion of TOSC that the D&M assessment contained errors in calculations in the risk analysis that render its risk characterization invalid, and therefore, the D&M conclusions concerning PAH impacts on aquatic organisms cannot be considered accurate. No major flaws were found with the SEH analysis. Although there was variability between samples taken from the site, the benthic community survey and laboratory experiments in the SEH document demonstrated evidence of ecological impacts. The conclusions of the SEH risk assessment regarding the likely impacts on benthic communities are valid, and recommendations regarding clean up criteria necessary to ameliorate ecological effects are appropriate.

Background

In 1998, Short Elliott Hendrickson, Inc. (SEH), under contract from Wisconsin DNR, completed an ecological risk assessment (ERA) for sediments offshore from the Northern States Power (NSP) property in Ashland, Wisconsin, contaminated with polycyclic aromatic hydrocarbons (PAHs) and volatile organic compounds (VOCs). SEH found:

• The contaminants in sediment offshore from the Ashland site are present at very high concentrations, but are distributed in a heterogeneous (patchy) manner. Concentrations are similar to those found at other contaminated sites where toxicity to aquatic and benthic organisms has been demonstrated;
• There was degradation of benthic communities in contaminated areas, compared to less contaminated areas;
• Laboratory toxicity studies in which various aquatic organisms were exposed to sediment or overlying water indicated growth and survival of most organisms was impaired at sampling locations with high concentrations of PAHs.

Based on the chemical concentrations, the laboratory toxicity tests and the benthic survey, SEH concluded there is a high probability of adverse effects in aquatic organisms from contaminated sediments at the Ashland site. NSP has questioned the methods used in the SEH study and the study's conclusion that nearly 10 acres of sediments should be dredged. According to WI-DNR and NSP representatives, the ecological risk assessment is key to selecting a remedy for the sediment contamination related to the site, the most ecologically significant area of contamination, but the findings of the assessment remain in dispute.

The TOSC program was asked to provide an independent review of the SEH ecological risk assessment, with particular focus on the study's findings and methodology. In addition, project partners asked that we address the major areas of disagreement between the SEH and Dames and Moore reports. The objective of this review was to address questions of basic science and engineering-there may be questions over policy or economics (e.g., future use of the Ashland site) that our review does not address.

The scientific merits of the ecological risk assessments conducted by SEH and D&M were evaluated by comparing and assessing the two assessments with respect to:

• Adherence to the Environmental Protection Agency's guidelines for conducting ecological risk assessments
• Accuracy and appropriate use of data from the scientific literature, and their relevance to the Ashland site
• Likely source of contaminants at Ashland
• Assumptions regarding exposure of organisms to contaminants
• Appropriate choice of biological endpoints
• Appropriate use of toxicological models and statistics
• Likelihood of long-term effects on biota
• Validity of conclusions, given the scientific evidence presented

Adherence to the U.S. Environmental Protection Agency's Guidelines for Ecological Risk Assessment

The currently accepted procedure for conducting ecological risk assessments is detailed in the U.S. EPA's Guidelines for Ecological Risk Assessment (U.S. Environmental Protection Agency 1998). The EPA Guidelines provides a framework that promotes scientifically sound decisions regarding the appropriate choice of endpoints and models, as well as adherence to risk management goals. The Guidelines divides the process into three distinct phases: problem formulation, analysis, and risk characterization. Within each stage, there are recommended techniques for formulating conceptual models describing how organisms might be exposed to contaminants, the appropriate choice of endpoints to measure exposure and effects of contaminants, and statistical techniques for analyzing data gathered during the assessment.

In both the SEH and D&M ERAs, there is a statement that the assessment was conducted in accordance with EPA Guidelines. Indeed, both documents are similarly structured with respect to the major sections of the risk assessment process. Both documents include problem formulation, analysis, and risk characterization phases. However, the assessments differ in the detail provided. Because the D&M assessment relied on data gathered by SEH for its assessment, in the D&M document certain aspects normally considered essential to a comprehensive ERA were not specifically addressed. For example, the D&M document lacks an assessment of the study design and data quality. Additionally, the D&M ERA did not address alternate exposure endpoints except the analytical chemistry values used to estimate exposure to PAHs. The SEH ERA addressed these issues, and provided much more detail with regard to the choice of effects and exposure endpoints, the conceptual model, the work plan, and sources of uncertainty in the analysis.

Several of the major differences between the two assessments arose because the two assessments used different data sets as input to the analysis phase. SEH conducted a preliminary (or screening level) study in which contaminant concentrations in sediment collected in 1996 were compared to sediment quality guidelines. Subsequently, concentrations of contaminants in sediment collected in 1998, together with lab tests and the benthic diversity survey were used to conduct the baseline assessment. In contrast, D&M used both the 1996 chemistry data (for which there was no organic carbon content data) as well as the 1998 data in its risk assessment, and referred to the lab experiments and benthic survey conducted by SEH as "verification studies," essentially excluding them from the assessment, as those data were not used in the risk decision.

Some discrepancies in document structure between the ERAs exist simply because D&M used data acquired by SEH during its screening level assessment, rather than gather additional data. Many of the sections absent from the assessment correspond to activities not performed by D&M during its assessment (the study design and selection of measurement endpoints, for example). The analysis phase of the D&M document, in particular, does not include details regarding exposure and effects characterizations. Because the SEH assessment detailed these sections, reiteration in the D&M document was not strictly necessary. While a more comprehensive assessment decreases uncertainty during the risk characterization and decision phases, the EPA framework for risk assessment is flexible, in the sense that not all risk assessments require all activities identified by the Guidelines.

Problem Formulation

During the problem formulation phase of risk assessment, the nature of the stressor (contaminant) is described, assessment endpoints (exposure, ecosystem characteristics, and effect) are chosen, a conceptual model of exposure is formed, and an analysis plan is laid out. The two risk assessments disagreed on a number of points during this phase.

Source of contaminants

The source of the contaminants detected in sediments offshore from the Ashland Lakefront Property was a point of conflict between the SEH and D&M assessments. In 1996, SEH collected offshore sediments at 80 locations and from several depths and quantified the concentrations of PAHs, VOCs, metals and others. The SEH document stated that the source of the contaminants was not certain, but based on the high concentrations of PAHs and VOCs in the sediments, the most likely source was the former manufactured gas plant (MGP) on the NSP property adjacent to the lakefront. D&M contended the evidence points to other sources of contaminants.

There is convincing evidence supporting the hypothesis of contamination by waste from the former MGP. The high concentration of PAHs detected in sediments (including the non-aqueous phase liquid, or NAPL layer) is consistent with the coal tar waste byproducts from MGPs. The very high concentrations and substantial depth to which PAHs penetrate the sediment (several meters) suggest contamination of the sediments occurred over a substantial length of time, such as during the operation of an MGP over many years. SEH provided evidence of an open sewer terminating offshore, through which wastes may have been directly deposited to sediments. Alternatively, the hydrogeology of the area indicates possible groundwater migration of waste materials from the former MGP to the lakefront.

D&M pointed out the Ashland sediments contain many lower molecular weight compounds, and suggested the PAH profile is more indicative of "fuel oils" than MGP wastes. While a PAH profile may be used as a "fingerprint" to broadly classify the source of PAHs in relatively clean sediment, the PAH profile of weathered MGP waste is virtually indistinguishable from fuel oils or creosote, based on molecular weight (Su et al. 2000). In the absence of an alternate source of PAH contamination such as fuel oil, it is most likely the source of PAHs and VOCs in the offshore sediments at the Ashland site is waste from the former manufactured gas plant.

For the purposes of the risk assessment, both SEH and D&M documents assumed PAHs were the primary contaminants. However, it was suggested by D&M that other compounds not measured by SEH during its analysis may contaminate the sediments, and may have contributed to the observed toxicity in the laboratory studies. Phenols, dioxins, furans, PCBs and pesticides were specifically identified as possible contaminants. However, sediments historically contaminated with PAHs by MGPs and creosote operations do not typically contain appreciable quantities of dioxins, furans, or PCBs. Although these chemicals are widespread contaminants in aquatic environments, they do not usually co-occur with PAHs. Chlorinated hydrocarbons have been found in some sediments historically contaminated by wood preservative operations in which both PAHs and pentachlorophenol were used (McKee et al. 1990). However, these classes of chemicals are rarely found at concentrations capable of inducing acute toxicity (Malins et al. 1985; Ozretich et al. 2000). In the absence of any credible source for these other compounds, the primary contaminants inducing toxicity in Ashland sediments are most likely those associated with MGPs or creosote waste contamination, specifically PAHs, VOCs, and heterocyclics (Mueller et al. 1989; Padma et al. 1998).

Further evidence supporting PAHs as the primary contaminants comes from the toxicity of the sediment itself. The toxicity observed in the laboratory experiments with Ashland sediment is consistent with the known toxicity of PAHs. For example, exposure to sediment spiked with 1.24 mg gTOC?1 fluoranthene caused 75% mortality in fathead minnows (Schlueter et al. 2000); a total PAH concentration of 1.0 mg gTOC-1 caused approximately the same mortality in fathead minnows exposed to Ashland sediment elutriate (SEH, Figure 16). Other studies have found similar mortality associated with sediments contaminated with mixtures of PAHs from coal tars and creosote (Roberts et al. 1989; Sved et al. 1997; Tagatz et al. 1983). The demonstration of UV-enhanced phototoxicity of Ashland sediments is also strong evidence pointing to PAHs as the predominant toxicant in the sediments.

The presence of other contaminants and their contribution to toxicity, if any, in the Ashland sediments is merely speculative, unless the concentration of each purported chemical (and other environmental contaminants not specifically mentioned) is quantified by chemical analysis. However, this approach is neither practical nor appropriate for environmental risk assessments, in which the integration of all available information regarding the contaminants, including known past and present sources, is favored (U.S. Environmental Protection Agency 1998). The very high levels of PAHs in Ashland sediment, the absence (or low levels) of other contaminants, and the absence of any other credible source of contaminants, suggest PAHs are the predominant contaminant and the most likely source of observed toxicity in Ashland sediments.

The SEH and D&M ERAs disagreed as to the depth of PAH contamination in Ashland sediments, and by inference, the PAH concentrations to which organisms at the surface are exposed. SEH measured concentrations of PAHs (as well as other contaminants) at several discrete depths in sediment cores. There was substantial variability in concentrations from location to location in the bay likely due to the patchy distribution of wood chips. The D&M ERA contended PAH concentrations in the upper six inches in which organisms reside are lower than in deeper sediments, and therefore pose less of a threat to organisms near the surface. SEH interpreted its data to indicate PAH levels are greatest near the surface and decrease with depth. Examination of data from sampling locations for which PAHs were measured both in the upper four feet and in deeper sediments confirms SEH's assertion: concentrations in the 0-4 ft cores exceed those in deeper cores more often. Samples in which PAH concentrations were measured in the upper six inches alone as well as at greater depths in cores from the same sampling location (there were only three) gave less clear results. The heterogeneous distribution of PAHs in the bay makes it impossible to state unequivocally whether PAHs increase or decrease with depth in Ashland sediments. However, given the high concentrations of PAHs measured in the upper six inches alone (SEH Table 5), the probability of contact between organisms residing in surface sediment layers with PAHs is high.

Contaminant fate and transport

Both ERAs identified fate and transport mechanisms for PAHs in aquatic sediments, as well as numerous species likely at risk from exposure to PAHs. The two documents disagreed on a few points regarding exposure of benthic invertebrates to PAHs, uptake of PAHs by aquatic organisms, and natural degradation rates of PAHs.

The D&M ERA suggested some amphipods and chironomids do not come in contact with the sediment itself or the pore water expected to contain the highest concentrations of PAHs, but rather remain above the surface of the sediment in the overlying water where PAH levels are lower. SEH assumed in its ERA that organisms found in Ashland samples penetrate the sediment during feeding activities and burrowing actions, and are exposed to high concentrations of PAHs. The vast majority of the scientific literature supports the SEH position that benthic organisms are likely to be exposed to high PAH levels.

SEH position that benthic organisms are likely to be exposed to high PAH levels. PAHs tend to partition rapidly to sediments (Bestari et al. 1998). However, some desorption from sediment to the aqueous phase can occur, especially from heavily contaminated sediments (Zhang et al. 2000). As pointed out in both ERAs, water circulation and bioturbation can enhance the rate of desorption. The overall contribution of simple dissolution of PAHs to the aqueous phase to exposure of organisms is likely small, as shown by the undetectable aqueous PAH concentrations in the water overlying the Ashland sediments.

The bulk of PAHs enter the water column associated with sediment solids. Chironomids have been shown in laboratory experiments to elevate levels of sediment-associated PAHs in overlying water by disturbing the sediment during burrowing activities (Ciarelli et al. 2000; Clements et al. 1994). D&M stated in its ERA that because PAHs in the water column remain bound to sediments, they are less bioavailable to organisms. However, some organisms such as mollusks filter indiscriminately on suspended particles, and can accumulate high levels of PAHs from suspended sediment (DeLeon et al. 1988; Gewurtz et al. 2000). In addition, some proportion of particles is phytoplankton, which can accumulate PAHs, and are selectively consumed by zooplankton.

PAHs may also be transferred to higher trophic levels by direct ingestion by benthic invertebrates. The literature suggests most benthic organisms accumulate PAHs to some degree when exposed to PAH-contaminated sediments. Chironomus riparius exposed in laboratory microcosms to spiked sediments accumulated benzo(a)pyrene or fluoranthene to very high concentrations (Clements et al. 1994). Numerous field studies have demonstrated uptake by isopods (van Hattum et al. 1998), amphipods (Gewurtz et al. 2000; Landrum et al. 1991), mussels (Metcalfe et al. 1997), and to varying degrees by fish (Burkhard and Lukasewycz 2000; Djomo et al. 1996). Organisms such as chironomids that are in close contact with sediment (e.g., during feeding) accumulate higher PAH concentrations than organisms with less contact (Gewurtz et al. 2000). Transfer of PAHs from sediment to benthos to fish through consumption has been demonstrated in laboratory studies (Clements et al. 1994). Therefore, the possibility for transfer of PAHs to higher trophic levels exists. Fish for which benthic invertebrates constitute a large fraction of their diet are at greater risk of exposure to PAHs.

Accumulation and transfer of PAHs to higher trophic levels is primarily a function of the ability of the organism to metabolize PAHs by the mixed function oxygenase (MFO) enzymes. Organisms with a well-developed MFO system, such as fish, rapidly metabolize PAHs; those with poor MFO systems, such as bivalve mollusks, accumulate PAHs (Albers 1995; Elder and Dresler 1988). As most aquatic organisms at higher trophic levels have well developed MFO systems, increases in accumulation through trophic levels do not occur.

Contrary to the suggestion in the D&M ERA, the resuspension of particles in the water column by the burrowing action of benthic invertebrates does not result in a significant degradation of PAHs. While the transfer of PAHs from sediment to biota may lower the sediment concentration somewhat, provided no new input occurs, it is incorrect to refer to this uptake by organisms as "bioremediation." Metabolic degradation of parent PAHs (i.e., unsubstituted rings) by some bacteria occurs under certain conditions (Lantz et al. 1997; McNally et al. 1998). However, the anoxic state of sediments, the recalcitrant chemical nature of higher molecular weight PAHs, as well as toxicity to bacteria (McConkey et al. 1997) slows bacterial degradation and inhibits remediation of contaminated sediments (Mueller et al. 1989). Fate studies have demonstrated that natural (i.e., not manipulated) bacterial degradation diminishes PAH sediment concentrations very slowly (Bestari et al. 1998). Moreover, given the very high concentrations of PAHs in Ashland sediments after many years, it must be assumed that natural degradation rates are low, and will not result in any significant attenuation of PAHs in the Ashland sediments in the near future.

On the subject of PAH biotransformation, D&M tended to minimize the impacts of PAH metabolites. D&M stated in its ERA that metabolism of PAHs usually results in detoxification, and imply that the cytochrome P450 monooxygenase system in fish efficiently removes PAHs before they enter the blood stream. This statement is incorrect. In most fish, the P450 enzyme system is present in gills and gut tissues, but is concentrated primarily in the liver. D&M stated that fish are deficient in epoxide hydrolase, one of the enzymes involved in carcinogenesis of PAH metabolites, citing a reference to a study of rainbow trout. While apparently true for rainbow trout, this statement does not hold for many other types of fish. Observations of DNA adducts and tumors in fish (including rainbow trout) in the lab and collected from PAH-contaminated sites indicate a strong correlation between PAHs and carcinogenicity (Couch and Harshbarger 1985; Ericson et al. 1999; Metcalfe et al. 1988).

An important aspect of PAH chemistry ignored in both ERAs is photooxidation. There is evidence to suggest that the toxicity of many PAHs is associated with the oxygen-substituted PAHs produced upon exposure to UV light. PAHs taken up into the tissues of aquatic invertebrates and the gills of fish may absorb UVA wavelengths in sunlight and are transformed to oxyPAHs (Mallakin et al. 2000). Some of the oxyPAHs produced in sunlight are many times more toxic than the parent (unsubstituted) PAH (Lampi et al. 2001; Marwood et al. 1999; McConkey et al. 1997). The UVA wavelengths of sunlight responsible for photooxidation of PAHs penetrate to substantial depths in lakes (Williamson et al. 1996). Although UVA penetration in water varies greatly with season and location even within the same lake (Smith et al. 1999), it is possible UVA penetrates to the sediment at Ashland in sufficient quantity to induce photooxidation. This mechanism may play a part in exposure and toxicity of PAHs at Ashland, although the importance relative to other mechanisms is unknown without additional experiments.

A major discrepancy between the ERAs by SEH and D&M is the acknowledgement of the role of UV light in toxicity. SEH identified UV-photoinduced toxicity as a major factor in acute toxicity, and devoted several experiments to determining the relative importance of UV on toxicity of Ashland sediments. In this phase of the risk assessment, D&M did not identify UV as a factor in toxicity, but in the section referred to as "verification studies" D&M attempted to minimize the impacts of UV and discredit the UV studies performed for the SEH assessment. Multiple studies throughout the past decade have demonstrated greatly enhanced PAH toxicity in the presence of UV, both in laboratory studies such as those performed for SEH, and in the field under realistic exposure conditions (Ankley et al. 1995; Diamond et al. 2000; Oris and Giesy 1987). The implications of these studies cannot simply be dismissed, and should have been acknowledged by D&M at this phase of its ERA.

Conceptual model

A conceptual model integrates available information on the stressors, effects, and receptor characteristics into a written and visual representation of the predicted relationships between contaminants and susceptible organisms (U.S. Environmental Protection Agency 1998). The EPA Guidelines indicates the exposure profile should explicitly define the exposure, with respect to intensity, frequency and spatial extent. Numerous mathematical models exist for the estimation of uptake rates by organisms, given the potential daily doses, feeding behavior, frequency of exposure, etc. SEH formed a qualitative conceptual model describing susceptible species and exposure pathways, but not explicitly in mathematical terms. The spatial extent of exposure was defined with reasonable precision given the heterogeneity of the PAH distribution in Ashland sediments. The D&M assessment described a simple exposure pathway but did not identify a conceptual model.

The exposure pathway proposed by D&M included reference to the wood chip layer as a possible source of humic acids, which have been shown to ameliorate PAH toxicity by binding organic compounds in the water column (Gensemer et al. 1998; Oris et al. 1990). Humic material is a complex organic compound produced by the breakdown of plant material. It is unknown whether the wood chips could even be a source of humic acids, as degradation of the chips may be retarded by the presence of the coal tars in the sediments immediately below. The presence of wood chips, presumably persisting over many years from abandoned industrial activities adjacent to the bay, suggests this may be the case. Regardless, it is unlikely that degradation of the wood chips resulted in release of humic acids at a concentration required to diminish bioavailability of PAHs, given the high concentration of PAHs in sediments.

Assessment endpoints

At the end of the problem formulation phase, relevant assessment endpoints are chosen. The selection of assessment endpoints is crucial to the outcome of the ERA. However, the identification of suitable endpoints in complex ecosystems is often a contentious issue when multiple stakeholders disagree on the value of endpoints. For a scientifically defensible ERA, endpoints must be both ecologically relevant and susceptible to known stressors (U.S. Environmental Protection Agency 1998). The SEH document identified as its assessment endpoints the protection of several pertinent groups of organisms from PAH impacts. Assessment endpoints were not identified at this phase in the D&M risk assessment.

In the analysis section of the assessment, D&M pointed out that in ecological risk assessments, protection of populations rather than individual organisms is preferred, and identified "protection of populations of benthic invertebrates" as its only assessment endpoint. While the omission of the term "populations" from the SEH assessment endpoints seems to be the basis of the D&M argument for a re-examination of the data and justification for a second risk assessment, it is an argument based on semantics only. Protection of populations is implied and assumed in the SEH endpoints, since SEH did not identify specific organisms (i.e., key species) requiring protection. It is the opinion of TOSC that the SEH assessment endpoints are appropriate.