8281 - Laserfiche WebLink

Home Browse Search

COMMENTARY 2Jt caveats necessary to addressing disadvantages with the ap- proach. Fausch et al. (1990) repeated for each of the ap- proaches the necessity of appropriate reference sites, and for the more data intensive methods (the IBI consists of 12 community attributes), the setting of a priori values, such as the expected fish community for a relatively unperturbed stream in specific ecoregion, by a competent fish ecol- ogist/biologist/ichthyologist. They acknowledge that most of the methods have been developed in the Midwest, and problems with the methods were encountered when applied to other ecoregions. The Ohio Environmental Protection Agency has expended substantial effort in applying the IBI method to water quality monitoring of that state (OEPA, 1988). Canton and Van Derveer (1997) and Van Derveer and Canton (1997) did not use any of these methods to substantiate that healthy fish communities occurred in streams in southeastern Colorado. Furthermore, their claim of no biological effects in Col- orado streams cannot be confirmed from the information given or referenced in their articles. Their statement seemed to have fallen into the null fallacy trap: (1) There is no evidence for adverse effects, versus (2) There is evidence for no adverse effects (J. Skorupa, USFWS, personal commun- ication, 1995), The null fallacy occurs when statement 1 (a null finding) is given equal weight as statement 2 (a positive finding). What often is overlooked is that a null finding usually implies a lack of positive evidence in both direc- tions-for effects or for absence of effects. Contrary to the claim of Canton and Van Derveer (1997) that there are no biological effects, there is evidence that fish populations in several Colorado, Utah, and New Mexico streams with elevated selenium concentrations are not healthy. Findings in several NIWQP investigations indicate that selenium concentrations are elevated sufficiently in water, bottom sediment, and biota in the middle Green, upper Colorado, Gunnison, Mancos, and San Juan rivers to cause adverse effects in fish and wildlife (Butler et al., 1989, 1991, 1994, 1995; Stephens et al. 1988, 1992; Peltz and Waddell, 1991; Blanchard et aI., 1993; Thomas et aI., 1997). These rivers have several endangered fish including Colora- do squawfish (Ptychocheilus lucius), razorback sucker (Xyrauchen texanus), bony tail (Gila elegans), and humpback chub (Gila cypha). Toxic effects from selenium and other in organics are important factors in the decline of these species and inhibition of their recovery (Hamilton, 1995, 1998; Buhl and Hamilton, 1993; Hamilton and Waddell, 1994; Waddell and May, 1995; Buhl, 1997; Hamilton and Buhl, 1997a, b; Hamilton et al., 1996; 2000; Stephens and Waddell, 1998). Moreover, the Recovery Program for the Endangered Fish Species of the Upper Colorado River Basin dropped one proposed floodplain restoration site for endangered fish because elevated selenium concentrations were present in water and biota, put six sites on temporary hold pending outcome of ongoing selenium research with endangered fish, and have three other sites on temporary hold due to concerns about other inorganic contaminants (P. Nelson, USFWS, written communication, 1997; Holley and Weston, 1995; Stephens et al., 1995; Archuleta and Holley, 1996). 4.4. Water/Diet Versus Sediment Uptake Canton and Van Derveer (1997) state that "The chronic toxicity threat posed by Se is primarily from dietary uptake, which is a result of its propensity to cycle through the sediment, where it enters the benthic-detrital food web and ultimately causes reproductive impairment in fish and wild- life through dietary uptake." The authors of this commen- tary agree that the detrital pathway is one of the most important pathways by which selenium is recycled. How- ever, waterborne uptake is responsible for the entry of selenium into the aquatic food web (Lemly and Smith, 1987). This critical point is not mentioned by Canton and Van Derveer (1997). Uptake of selenium by bacteria, algae, and aquatic invertebrates rapidly removes selenium from the water column and concomitantly reduces water concen- trations, which can give a false impression of low water- borne selenium concentrations precluding adverse effects. For example, Besser et al. (1993) reported that the biocon- centration factor for algal accumulation of selenium from water in absence of sediment was 5300-15,700 for sele- nomethionine, 1440-1600 for selenite, and 428 for selenate, and for daphnids these bioconcentration factors were 30,300-229,000, 570-3600, and 293, respectively. Canton and Van Derveer (1997) have presented an incomplete dis- cussion that does not address the importance of waterborne selenium in food web residue dynamics, and do not present data to support their underlying assumption that selenium bioaccumulation and toxicity are more strongly linked to the sediment pathway than the aqueous pathway. Canton and Van Derveer (1997) further state that "Partic- ulate Se, typically measured as either sediment, detrital, or suspended particulate Se, has been recently identified as a better predictor of adverse biological effects." However, one of the key references they cited in support of this statement (Presser et aI., 1994) does not promote sediment, detrital, or suspended particulate selenium as the best pre- dictor, but rather stresses the importance of defining sel- enium contamination based on an ecosystem level evaluation that includes the collection of a variety of food chain organisms (T. Presser, USGS, personal communica- tion, 1997). Presser and Piper (1998) have further illustrated this point in a recent publication in which they apply a mass balance approach to data available for systems impacted by selenium in California. Their definition of mass balance has been expanded beyond that documented in water-quality loads because of the nonconservative behavior of selenium. and consequently they recommend including a biotic