US Geological Survey

USGS Staff – Published Research

University of Nebraska - Lincoln Year 

Environmental contaminants in

freshwater fish and their risk to

piscivorous wildlife based on a national

monitoring program

Jo Ellen Hinck∗ Christopher J. Schmitt†

Kimberly A. Chojnacki‡ Donald E. Tillitt∗∗

∗Columbia Environmental Research Center, jhinck@usgs.gov

†Columbia Environmental Research Center

‡Columbia Environmental Research Center

∗∗Columbia Environmental Research Center

This paper is posted at DigitalCommons@University of Nebraska - Lincoln.

http://digitalcommons.unl.edu/usgsstaffpub/83
Environmental contaminants in freshwater fish

and their risk to piscivorous wildlife based

on a national monitoring program

Jo Ellen Hinck & Christopher J. Schmitt &

Kimberly A. Chojnacki & Donald E. Tillitt

Received: 18 November 2007 /Accepted: 10 April 2008 / Published online: 2 July 2008

# Springer Science + Business Media B.V. 2008

Abstract Organochlorine chemical residues and elemental concentrations were measured in piscivorous

and benthivorous fish at 111 sites from large U.S. river

basins. Potential contaminant sources such as urban

and agricultural runoff, industrial discharges, mine

drainage, and irrigation varied among the sampling

sites. Our objectives were to provide summary statistics for chemical contaminants and to determine if

contaminant concentrations in the fish were a risk to

wildlife that forage at these sites. Concentrations of

dieldrin, total DDT, total PCBs, toxaphene, TCDDEQ, cadmium, chromium, mercury, lead, selenium,

and zinc exceeded toxicity thresholds to protect fish

and piscivorous wildlife in samples from at least one

site; most exceedences were for total PCBs, mercury,

and zinc. Chemical concentrations in fish from the

Mississippi River Basin exceeded the greatest number

of toxicity thresholds. Screening level wildlife risk

analysis models were developed for bald eagle and

mink using no adverse effect levels (NOAELs), which

were derived from adult dietary exposure or tissue

concentration studies and based primarily on reproductive endpoints. No effect hazard concentrations

(NEHC) were calculated by comparing the NOAEL to

the food ingestion rate (dietary-based NOAEL) or

biomagnification factor (tissue-based NOAEL) of each

receptor. Piscivorous wildlife may be at risk from a

contaminant if the measured concentration in fish

exceeds the NEHC. Concentrations of most organochlorine residues and elemental contaminants represented no to low risk to bald eagle and mink at most

sites. The risk associated with pentachloroanisole,

aldrin, Dacthal, methoxychlor, mirex, and toxaphene

was unknown because NOAELs for these contaminants were not available for bald eagle or mink. Risk

differed among modeled species and sites. Our

screening level analysis indicates that the greatest risk

to piscivorous wildlife was from total DDT, total

PCBs, TCDD-EQ, mercury, and selenium. Bald eagles

were at greater risk to total DDT and total PCBs than

mink, whereas risks of TCDD-EQ, mercury, and

selenium were greater to mink than bald eagle.

Keywords Biomonitoring . Ecological risk

assessment . Toxicity . Dietary exposure .

Biomagnification

Introduction

The bioaccumulation of contaminants by fish represents a

potentially significant risk to piscivorous wildlife. Persistent, accumulative substances such as DDT (as DDE),

Environ Monit Assess (2009) 152:469–494

DOI 10.1007/s10661-008-0331-5

J. E. Hinck (*) : C. J. Schmitt : K. A. Chojnacki :

D. E. Tillitt

U.S. Geological Survey (USGS),

Columbia Environmental Research Center (CERC),

4200 New Haven Rd.,

Columbia, MO 65201, USA

e-mail: jhinck@usgs.gov
polychlorinated biphenyls (PCBs), mercury, and selenium can accumulate in fish and pose a threat to piscivorous wildlife. Fish often comprise a substantial dietary

portion of species such as mink (Mustela vison), river

otter (Lutra canadensis), osprey (Pandion haliaetus),

and bald eagle (Haliaeetus leucocephalus), and therefore represent a significant route of contaminant

exposure. In addition, wildlife typically consume most

or all of the fish, including internal organs, in which

concentrations of many contaminants are generally

greater than those in the muscle tissue (fillet) typically

used to assess human health risks.

The Large River Monitoring Network (LRMN) of

the Biomonitoring of Environmental Status and

Trends (BEST) Project was initiated to document

spatial trends in accumulative chemical contaminants,

health indicators, and reproductive biomarkers. From

1995 through 2004, the BEST-LRMN Project measured concentrations of organochlorine pesticides,

PCBs, and elemental contaminants in whole-body

composite samples of piscivorous and benthivorous

fish from nine U.S. river basins (Schmitt 2002; 2004;

Schmitt et al. 2005; Hinck et al. 2006a, b; 2007,

2008). Hinck et al. (2006b) determined that avian and

mammalian piscivorous wildlife were at risk to

mercury, but not organochlorine pesticides or PCBs,

in northern pike (Esox lucius), longnose sucker

(Catostomus catostomus), and burbot (Lota lota) from

the Yukon River Basin. However, the risk of chemical

contaminants to piscivorous wildlife that live and

forage at BEST-LRMN sites in the conterminous

United States has not been evaluated. Therefore, this

paper reports a wildlife risk analysis that was

conducted to determine the risk of these contaminants

to avian (bald eagle) and mammalian (mink) receptors

that consume the large adult fish such as those

collected by the BEST-LRMN Project and have a

broad geographic range. Unlike many ecological risk

assessments that estimate fish contaminant concentrations from sediment or soil contaminant concentrations, contaminant concentrations in whole-body

fish samples from the BEST-LRMN Project were

used in the analysis to give a more accurate estimate

of risk to piscivorous wildlife.

The BEST-LRMN Project was preceded by the

National Contaminant Biomonitoring Program

(NCBP). The NCBP measured organochlorine residues and elemental contaminants in whole-body fish

composite samples from 1976 through 1986 at 107

locations across the United States (Schmitt et al.

1999). The NCBP demonstrated that concentrations

of organochlorine pesticides, PCBs, and certain

elemental contaminants such as mercury were generally decreasing in fish and provided descriptive

statistics for concentration distributions in fish from

across the United States. The geometric means and

85th percentile concentrations from the NCBP have

been previously used as benchmarks for national

trends of contaminant concentrations in whole-body

fish (Schmitt et al. 1999). Some of the BEST-LRMN

sites were co-located at historical NCBP sites, and

trend analysis for contaminant concentrations were

included in individual basin reports (Schmitt 2002;

2004; Schmitt et al. 2005; Hinck et al. 2006a; 2007).

This paper provides an example of how contaminant concentrations in fish from a routine monitoring

program can be used to screen for risk of contaminant

exposure to piscivorous wildlife. This approach

follows the procedures developed by the U.S.

Environmental Protection Agency (USEPA) to conduct screening-level ecological risk assessments

(USEPA 1998). Although these assessments are

usually applied locally to assess specific sites, they

have been used in regional assessments. For example,

Yeardley et al. (1998) examined the risk of mercury to

wildlife in northeastern U.S. lakes, and Hinck et al.

(2006b) determined the risk of contaminant exposure

to avian and mammalian receptors in the Yukon River

Basin. Our objectives were (1) to provide summary

statistics for organochlorine chemical residues and

elemental contaminants measured by the BESTLRMN Project and (2) to screen for contaminants

that represent a risk to bald eagle and mink and

identify the areas where the risk to these receptors

was greatest. The wildlife risk analysis data from

Hinck et al. (2006b) were updated in this paper to

provide results for all BEST-LRMN samples. These

results were then compared to site specific studies to

determine if piscivorous wildlife had previously been

identified as at risk to contaminant exposure.

Materials and methods

Detailed sampling, field, and laboratory procedures

have been previously described (Schmitt 2002; 2004;

Schmitt et al. 2005; Hinck et al. 2006a, b; 2007;

2008). A summary is presented here.

470 Environ Monit Assess (2009) 152:469–494
Sampling and field procedures

Fish were collected from 111 sites in nine river basins

(Table 1; Fig. 1). Sites were located in the Apalachicola River Basin (n=3), Colorado River Basin (n=

14), Columbia River Basin (n=16), Mobile River

Basin (n=4), Mississippi River Basin (n=48), Pee

Dee River Basin (n=3), Rio Grande Basin (n=10),

Savannah River Basin (n=3), and Yukon River Basin

(n=10; Table 1). Of these, 51 were NCBP sites that

had been sampled historically (Schmitt et al. 1999).

Each site was sampled once. Sampling sites were

located on the mainstem and large tributaries of the

river basins with known access points without regard

to contaminant sources; however, the sites did

represent a range of contaminant sources (e.g.,

chemical manufacturing, agricultural, and urban

areas). Most fish were captured by electrofishing;

hook-and-line, gill nets, and fyke nets were used only

in the Yukon River Basin. Piscivorous species

included black bass (Micropterus spp.), white bass

(Morone spp.), catfish (Ictaluridae), northern pike,

northern pikeminnow (Ptychocheilus oregonensis),

burbot, trout (Salmonidae), pikeperch (Sander spp.),

and goldeye (Hiodon alosoides). Benthivorous species included common carp (Cyprinus carpio) and

sucker (Catistomidae; Table 2).

Laboratory analyses

Fish were shipped to the laboratory frozen on dry ice

and stored at −20°C until prepared for analysis.

Individual whole fish were partly thawed, cut into

pieces, and ground to a fine texture. The ground subsamples were then grouped to create a single

homogenous composite sample for each site, species,

and gender combination. A total of 409 composite

samples were included in the chemical analysis.

One sub-sample (10 g) of each composite sample

was solvent-extracted and analyzed gravimetrically

for lipid content and by high-resolution capillary gas

chromatography with electron capture detection for

organochlorine pesticide residues and total PCBs

(sum of individual congeners) after size exclusion

and adsorption column cleanup. Pentachlorobenzene,

pentachloroanisole, aldrin, Dacthal, heptachlor, endosulfan I, endosulfan II, endosulfan sulfate, and

methoxychlor were not measured in Columbia, Rio

Grande, and Mississippi River Basin samples. Quality

assurance (QA) measures for the organochlorine

pesticide and PCB analyses included the analysis of

blanks, triplicate analyses, and matrix spikes. The

detection limit for each compound was calculated by

adding the average procedural blank concentration to

three times the procedural blank standard deviation

(Keith 1991) and varied among study years with

detection limits changing by two orders of magnitude

for some organochlorine residues (Table 3). Concentrations were converted from ng/g wet weight (ww) to

μg/g ww for statistical and wildlife risk analyses.

A second sub-sample of each fish composite

sample (100 g) was freeze–dried for elemental

analysis. Percent moisture was determined as weight

lost during lyophilization. For samples from the

Mississippi, Rio Grande, and Columbia River Basins,

one portion of the dried material was acid-digested

and analyzed by atomic absorption spectroscopy and

inductively coupled plasma emission spectroscopy.

For all other samples, three sub-samples were

analyzed for elemental contaminants. One portion of

the dried material was digested in nitric acid and

analyzed by inductively coupled plasma mass spectroscopy for cadmium, copper, chromium, nickel,

lead, and zinc. A second portion was dry-ashed

(magnesium nitrate-nitric acid-HCl) and analyzed by

hydride generation atomic absorption spectroscopy

for arsenic and selenium. A third portion was

analyzed directly for total mercury using thermal

combustion, amalgamation, and atomic absorption

spectroscopy. QA measures for the elemental analyses

included the analysis of reagent blanks, duplicate

samples, certified reference materials, and fortified

samples. Dry-weight (dw) detection limits were

determined individually for each element in each

sample. Elemental concentrations (including detection

limits) were converted from μg/g dw to μg/g ww for

statistical analysis and reporting using the moisture

content of each sample (Table 4).

A third sub-sample (10 g) was solvent-extracted

and subjected to reactive cleanup for use in the H4IIE

bioassay (Tillitt et al. 1991; Whyte et al. 2004).

Concentrations of 2,3,7,8-tetrachlorodibenzo-p-dioxin

equivalents (TCDD-EQ; pg/g ww) were determined

by slope ratio assay as modified by Ankley et al.

(1991). QA measures for the H4IIE bioassay included

analysis of duplicate samples and reference materials.

Concentrations were converted to μg/g ww for the

wildlife risk analysis.

Environ Monit Assess (2009) 152:469–494 471
Table 1 Freshwater fish collection sites of the BEST-LRMN Project

Basin site no. River Nearest city or feature Latitude longitude

Apalachicola River Basin

330 Chattahoochee Omaha, GA 32°13′19.80′′ N, 84°55′35.10′′ W

331 Flint Albany, GA 31°34′34.86′′ N, 84°08′49.80′′ W

332 Apalachicola Blountstown, FL 30°25′58.20′′ N, 85°01′17.10′′ W

Colorado River Basin

311 Yampa Lay, CO 40°25′38.00′′ N, 107°51′24.00′′ W

312 Green Ouray NWR, UT 40°08′31.00′′ N, 109°39′17.00′′ W

313 Green San Rafael, UT 38°45′56.00′′ N, 110°05’16.00’’W

314 Gunnison Delta, CO 38°45′59.58′′ N, 108°02′30.30′′ W

315 Colorado Loma, CO 39°09′39.00′′ N, 108°48′28.56′′ W

316 Colorado Gold Bar Canyon, UT 38°34′02.00′′ N, 109°38′51.00′′ W

317 San Juan Hogback Diversion, NM 36°44′41.00′′ N, 108°41′28.00′′ W

319 Colorado South Cove, AZ 36°05′23.70′′ N, 114°06′12.30′′ W

320 Colorado Willow Beach, AZ 35°52′33.12′′ N, 114°39′53.10′′ W

321 Colorado Needles, CA 34°43′44.64′′ N, 114°20′12.96′′ W

322 Colorado Imperial Dam, AZ 32°54′05.94′′ N, 114°28′09.48′′ W

323 Gila Hayden, AZ 33°01′22.14′′ N, 110°44′16.32′′ W

324 Gila Phoenix, AZ 33°22′33.42′′ N, 112°18′19.20′′ W

325 Gila Arlington, AZ 33°19′06.92′′ N, 112°40′26.46′′ W

Columbia River Basin

41 Snake Hagerman, ID 42°47′36.21′′ N, 114°56′18.10′′ W

42 Salmon Riggins, ID 45°35′43.42′′ N, 116°16′55.00′′ W

43 Snake Lewiston, ID 46°24′54.28′′ N, 117°02′03.49′′ W

44 Yakima Granger, WA 46°20′49.31′′ N, 120°12′27.03′′ W

45 Willamette Oregon City, OR 45°19′03.47′′ N, 122°39′57.50′′ W

46 Columbia Cascade Locks, OR 45°41′23.11′′ N, 121°51′00.41′′ W

96 Snake Ice Harbor Dam, WA 46°41′51.68′′ N, 118°53′07.88′′ W

97 Columbia Pasco, WA 46°31′49.22′′ N, 119°16′42.07′′ W

98 Columbia Grand Coulee, WA 47°57′44.85′′ N, 118°58′53.84′′ W

117 Flathead Creston, MT 48°09′01.09′′ N, 114°11′29.71′′ W

501 Columbia Beaver Army Terminal, OR 46°10′57.86′′ N, 123°04′13.87′′ W

502 Columbia Warrendale, OR 45°38′00.82′′ N, 121°58′42.57′′ W

503 Columbia Vernita Bridge, WA 46°37′28.40′′ N, 119°51′31.45′′ W

504 Columbia Northport, WA 48°58′21.70′′ N, 117°38′48.92′′ W

505 Willamette Portland, OR 45°33′04.51′′ N, 122°41′43.74′′ W

506 Columbia Vancouver, WA 45°35′44.21′′ N, 122°32′13.61′′ W

Mobile River Basin

326 Tombigbee Lavaca, AL 32°15′53.60′′ N, 88°00′44.21′′ W

327 Coosa Childersburg, AL 33°19′57.76′′ N, 86°21′55.87′′ W

328 Alabama Eureka Landing, AL 31°23′14.06′′ N, 87°42′42.19′′ W

329 Mobile Bucks, AL 31°03′15.85′′ N, 87°59′48.07′′ W

Mississippi River Basin

15 Mississippi Luling, LA 29°59′53.2′′ N, 90°25′31.1′′ W

23 Kanawha Winfield, WV 38°29′06.0′′ N, 81°48′57.6′′ W

24 Ohio Marietta, OH 39°24′36.8′′ N, 81°26′26.3′′ W

25 Cumberland Clarksville, OH 36°32′28.6′′ N, 87°22′04.7′′ W

26 Illinois Beardstown, IL 40°07′50.6′′ N, 90°20′45.6′′ W

27 Mississippi Guttenburg, IA 42°43′37.2′′ N, 91°01′30.0′′ W

28 Arkansas Pine Bluff, AR 34°16′27.0′′ N, 94°57′12.0′′ W

472 Environ Monit Assess (2009) 152:469–494
Table 1 (continued)

Basin site no. River Nearest city or feature Latitude longitude

29 Arkansas Keystone Res., OK 36°07′54.0′′ N, 96°20′47.0′′ W

30 White Devall’s Bluff, AR 34°47′01.0′′ N, 91°26′28.0′′ W

31 Missouri Nebraska City, NE 40°40′15.9′′ N, 95°49′44.6′′ W

32 Missouri Garrison Dam, ND 47°28′27.3′′ N, 101°26′15.5′′ W

67 Allegheny Natrona, PA 40°39′54.0′′ N, 79°41′24.0′′ W

68 Wabash New Harmony, IN 38°11′58.4′′ N, 87°58′36.0′′ W

70 Ohio Metropolis, IL 37°07′40.8′′ N, 88°39′25.2′′ W

71 Tennessee Savannah, TN 35°12′52.0′′ N, 88°18′36.0′′ W

72 Wisconsin Woodman, WI 43°05′42.0′′ N, 90°48′57.6′′ W

73 Des Moines Keosauqua, IA 40°44′52.8′′ N, 91°59′38.4′′ W

74 Mississippi Little Falls, MN 45°58′48.0′′ N, 94°22′00.0′′ W

75 Mississippi Cape Girardeau, MO 37°18′36.0′′ N, 89°31′01.2′′ W

76 Mississippi Memphis, TN 38°08′30.3′′ N, 90°03′36.6′′ W

77 Arkansas John Martin Res., CO 38°03′55.0′′ N, 102°56′02.0′′ W

78 Verdigris Oolagah, OK 36°31′16.0′′ N, 95°33′37.0′′ W

79 Canadian Eufaula, OK 35°16′43.0′′ N, 95°34′39.0′′ W

80 Yazoo Redwood, MS 32°24′36.0′′ N, 90°55′27.0′′ W

81 Red Alexandria, LA 31°20′48.0′′ N, 92°27′37.0′′ W

82 Red Lake Texoma, TX/OK 33°52′08.0′′ N, 96°47′04.0′′ W

83 Missouri Hermann, MO 38°42′24.1′′ N, 91°26′17.5′′ W

84 Big Horn Hardin, MT 45°52′12.2′′ N, 107°34′34.0′′ W

85 Yellowstone Sidney, NE 47°34′46.8′′ N, 104°13′10.7′′ W

86 James Olivet, SD 43°13′45.0′′ N, 97°41′05.0′′ W

89 Platte Louisville, NE 40°59′33.1′′ N, 96°12′30.9′′ W

90 Kansas Bonner Springs, KS 39°02′47.0′′ N, 94°47′05.0′′ W

111 Mississippi Lake City, MN 44°22′49.8′′ N, 92°07′33.0′′ W

112 Mississippi Dubuque, IA 42°26′27.6′′ N, 90°35′06.0′ W

201 Big Sunflower Anguilla, MS 32°58′18.0′′ N, 90°46′40.0′′ W

202 Bogue Phalia Leland, MS 33°24′22.0′′ N, 90°50′26.0′′ W

203 Steele Bayou Rolling Fork, MS 32°54′71.0′′ N, 90°57′10.0′′W

204 Tensas Tendal, LA 32°25′56.0′′ N, 91°21′57.0′′ W

205 S. Skunk Oskaloosa, IA 41°21′19.0′′ N, 92°39′31.0′′ W

206 Iowa Morengo, IA 41°50′23.0′′ N, 92°11′54.0′′ W

207 Cache Cotton Plant, AR 35°02′32.0′′ N, 91°19′12.0′′ W

208 Cache Egypt, AR 35°51′23.0′′ N, 90°56′15.0′′ W

209 S. Fork Iowa New Providence, IA 42°19′26.0′′ N, 93°10′10.0′′ W

210 Iowa Rowan, IA 42°45′36.0′′ N, 93°37′23.0′′ W

211 Cedar St. Charles City, IA 43°03′45.0′′ N, 92°40′23.0′′ W

212 Little R. Ditch Moorehouse, MO 36°50′03.0′′ N, 89°43′48.0′′ W

213 Wolf LaGrange, TN 35°01′57.0′′ N, 89°14′48.0′′ W

400a Leetown Res. Kearneysville, WV 39°21′2.15′′ N, 77°55′32.69′′ W

Pee Dee River Basin

336 Pee Dee Rockingham, NC 34°53′22.14′′ N, 79°51′24.89′′ W

337 Pee Dee Pee Dee, SC 34°21′23.22′′ N, 79°41′35.19′′ W

338 Pee Dee Bucksport, SC 33°42′18.09′′ N, 79°11′24.00′′ W

Rio Grande Basin

16 Rio Grande Mission, TX 26°09′28.74′′ N, 98°20′02.82′′ W

63 Rio Grande Elephant Butte Res., NM 33°12′48.55′′ N, 107°13′27.26′′ W

64 Rio Grande Alamosa, CO 37°25′06.42′′ N, 37°23′38.57′′ W

65 Pecos Red Bluff Lake, TX 32°00′00.00′′ N, 104°58′30.00′′ W

Environ Monit Assess (2009) 152:469–494 473
Summary statistics including the number of

detected concentrations, unweighted geometric

means, median, and maximum concentrations were

computed to examine distributions of each chemical

contaminant for the entire dataset. The 85th percentile

also was computed to compare to historical NCBP

concentrations (Schmitt et al. 1999). Organochlorine

residues and elemental contaminant concentrations

were compared to toxicity thresholds based on wholebody fish concentrations and wildlife toxicity values

from the scientific literature. Criteria for inclusion of a

toxicity threshold in our evaluation were that the

threshold was based on a whole-body fish concentration and associated with reproductive performance,

growth, or survival. Wildlife toxicity values were

used when whole-body toxicity threshold were not

available. Censored values (i.e., concentrations less

than the detection limit) were replaced by one half the

detection limit in all figures.

Screening level wildlife risk analysis

Risk to piscivorous wildlife was evaluated with

models based on adult dietary exposure or tissue

concentrations. Dietary-based toxicity reference values (TRVs) based on no observed adverse effect

levels (NOAELs) from the scientific literature were

used for most contaminants (Linder et al. 1980; Hill

and Camardese 1986; Sample et al. 1996; Tillitt et al.

1996; USEPA 1995). An adjustment factor for differences in body size was applied when the mammal

NOAEL TRV was not based on a mink study; a

scaling factor of 1 was used for interspecies extrapolation among birds (Sample et al. 1996). Consensus

TRVs adopted by the USEPA were used for arsenic,

cadmium, chromium, copper, nickel, and lead

(USEPA 1995; 2005a, b, c, d; 2007a, b). Tissuebased TRVs developed from NOAEL egg concentrations were used for total DDT, total PCB, and

TCDD-EQ in the bald eagle model because developing embryos are more sensitive to these contaminants

than adult birds (Elliott and Harris 2001). The TRV

for TCDD-EQ was based on a toxic equivalent that

was developed using consensus toxic equivalent

factors (van den Berg et al. 1998; Elliott and Harris

2001). A tissue-based TRV was also used for mercury

in the bald eagle model (Giesy et al. 1995) after the

use of a dietary-based TRV (0.0064 mg/kg/d; Sample

Table 1 (continued)

Basin site no. River Nearest city or feature Latitude longitude

511 Arroyo Colorado Harlingen, TX 26°11′44.28′′ N, 97°36′20.52′′ W

512 Rio Grande Brownsville, TX 25°52′12.96′′ N, 97°27′06.30′′ W

513 Rio Grande Below Falcon Dam, TX 20°08′06.66′′ N, 99°08′06.42′′ W

514 Rio Grande Below Amistad Res., TX 29°26′49.06′′ N, 101°03′10.58′′ W

515 Rio Grande Foster Ranch, Langtry, TX 29°46′40.91′′ N, 101°45′13.22′′ W

516 Rio Grande El Paso, TX 31°47′55.00′′ N, 31°48′15.00′′ W

Savannah River Basin

333 Savannah Augusta, GA 33°22′00.18′′ N, 81°56′46.44′′ W

334 Savannah Sylvania, GA 33°01′16.86′′ N, 81°31′04.50′′ W

335 Savannah Port Wentworth, GA 32°13′26.34′′ N, 81°08′47.04′′ W

Yukon River Basin

301 Yukon Eagle, AK 65°22′28.56′′ N, 142°30′20.88′′ W

302 Porcupine Fish Hook Bend, AK 67°12′50.94′′ N, 142°09′58.56′′ W

303 Porcupine Fort Yukon, AK 66°35′21.06′′ N, 145°19′59.40′′ W

304 Ray Near the Bridge, AK 65°53′01.26′′ N, 149°48′00.72′′ W

305 Tanana Fairbanks, AK 64°45′02.52′′ N, 148°03′15.84′′ W

306 Tolovana Nenana, AK 64°56′56.40′′ N, 149°44′13.44′′ W

307 Yukon Tanana, AK 65°08′09.72′′ N, 152°24′54.72′′ W

308 Yukon Galena, AK 64°41′56.22′′ N, 156°58′36.30′′ W

309 Innoko Innoko NWR, AK 63°38′33.66′′ N, 158°00′43.56′′ W

310 Yukon Kotlik, AK 63°02′12.18′′ N, 163°34′22.56′′ W

a

Reference site in original study (Schmitt 2002)

474 Environ Monit Assess (2009) 152:469–494
Fig. 1 Freshwater fish collection sites in the Yukon River Basin

(YRB), Columbia River Basin (CRB), Colorado River Basin (CORB),

Rio Grande Basin (RGB), Mississippi River Basin (MRB), Mobile

River Basin (MORB), Apalachicola River Basin (ARB), Savannah

River Basin (SRB), and Pee Dee River Basin (PRB). See Table 1 for

specific site information

Table 2 Benthivorous and piscivorous species collected from each basin

River Basin Benthivores Piscivores

Yukon Longnose sucker Northern pike, burbot

Columbia Common carp, largescale sucker Largemouth bass, smallmouth bass, northern pikeminnow, walleye,

rainbow trout

Colorado Common carp, white sucker Largemouth bass, smallmouth bass, channel catfish, brown trout

Rio Grande Common carp Largemouth bass, smallmouth bass, spotted bass, white bass,

northern pike, channel catfish

Mississippi Common carp, white sucker,

smallmouth buffalo

Largemouth bass, smallmouth bass, spotted bass, striped bass,

white bass, northern pike, goldeye, sauger, brown trout

Mobile Common carp Largemouth bass

Apalachicola Common carp Largemouth bass

Savannah Common carp, redhorse sucker Largemouth bass

Pee Dee Common carp Largemouth bass

Environ Monit Assess (2009) 152:469–494 475
et al. 1996) was found to be too conservative,

resulting in risk at most sites (n=105). Models

assuming a diet of 100% fish were used to screen

for risk. Site-specific data for other exposure pathways such as water and sediment were not available

for all sites and were not considered in this risk

evaluation; exposure from these other pathways was

expected to be low and therefore would contribute

little to the overall risk. A no effects hazard

concentration (NEHC) was calculated for each species modeled using the equation NEHC (μg/g)=

[dietary-based NOAEL TRV (mg/kg body weight/

day)]/food ingestion rate (kg food/kg body weight/

day)] or [tissue-based NOAEL TRV (mg/kg egg)/

biomagnification factor (BMF)fish→egg]. The bald

eagle-specific BMFs were 28 for total PCBs, 19 for

TCDD-EQ, 22 for total DDT, and one for mercury

(Giesy et al. 1995). Piscivorous wildlife may be at

risk from a contaminant if the measured concentration

in the fish composite sample exceeds the NEHC. Risk

Table 3 Detection limits (μg/g wet weight unless otherwise noted) for organochlorine residues

Analyte Mississippi River

Basin (1995)

Rio Grande and

Columbia River

Basins (1997)

Yukon River

Basin (2002)

Colorado River

Basin (2003)

Mobile, Apalachicola,

Savannah, and Pee Dee

River Basins (2004)

Pentachlorobenzene NM NM 0.00005−0.00006 0.00007 0.00007

Hexachlorobenzene 0.01 0.01 0.0021−0.0025 0.00026 NA

Pentachloroanisole NM NM 0.00023−0.00025 0.00013 0.00010

α-HCH 0.01 0.01 0.00018−0.00020 0.00008 0.00010

β-HCH 0.01 0.01 0.00017−0.00019 0.00019 0.00030

γ-HCH 0.01 0.01 0.00039−0.00044 0.00051 0.00010

δ-HCH 0.01 0.01 0.00010−0.00011 0.00005 0.00010

Aldrin NM NM NA 0.00027 0.00009

Dieldrin 0.01 0.01 0.00016 0.00015 0.00008

Endrin 0.01 0.01 0.00011−0.00012 0.00010 0.00025

Dacthal NM NM 0.00029−0.00032 0.00050 0.00013

Heptachlor NM NM 0.00019−0.00021 0.00019 0.00010

Heptachlor epoxide 0.01 0.01 0.00019−0.00021 0.00001 0.00010

Oxychlordane 0.01 0.01 0.00008−0.00009 0.00008 0.00010

cis-chlordane 0.01 0.01 0.00019−0.00022 0.00004 0.00028

trans-chlordane 0.01 0.01 0.00031−0.00035 0.00023 NA

cis-nonachlor 0.01 0.01 0.00008−0.00009 0.00010 NA

trans-nonachlor 0.01 0.01 0.00021−0.00023 NA NA

o,p’-DDE 0.01 0.01 0.00015−0.00016 0.00008 0.00081

o,p’-DDD 0.01 0.01 0.00061−0.00070 0.00055 0.00010

o,p’-DDT 0.01 0.01 0.00012−0.00013 0.00053 0.00010

p,p’-DDE 0.01 0.01 0.00039−0.00042 NA NA

p,p’-DDD 0.01 0.01 0.00038−0.00043 0.00067 NA

p,p’-DDT 0.01 0.01 0.0014−0.0016 0.0014 0.00047

Endosulfan I NM NM 0.00034−0.00039 0.00021 0.00010

Endosulfan II NM NM 0.00014−0.00019 0.00017 0.00010

Endosulfan sulfate NM NM 0.00079−0.00090 0.00005 0.00014

Methoxychlor NM NM 0.0017−0.0019 0.00035 0.00010

Mirex 0.01 0.01 0.00001 0.00005 0.00010

PCB 0.05 0.03 0.02 0.048 NA

Toxaphene 0.05 0.03 0.011 0.024 0.010

TCDD-EQ (pg/g) 0.2−1.0 0.4−1.3 1.7 0.1−0.5 1.0−2.6

Study year is shown in parentheses

HCH Hexachlorocyclohexane, PCB polychlorinated biphenyl, NM analyte not measured, NA all concentrations were greater than the

detection limit

476 Environ Monit Assess (2009) 152:469–494
was categorized as none (<NEHC), low (1–10×

NEHC), or high (>10× NEHC). Model species

considered in the risk analysis included the bald eagle

as an avian receptor and mink as a mammalian

receptor. Both of these species have distributions that

span most of the United States and consume the large,

adult fish such as those collected by the BEST-LRMN

Project. Some contaminants we evaluated have been

associated with the death of bald eagles and mink. For

example, reproductive impairment of mink has been

directly related to PCBs in Great Lakes fish (Aulerich

and Ringer 1977; Heaton et al. 1995), and mercury

has been implicated in increased mink mortality

(Wren et al. 1987). Bald eagle survival and reproduction have been negatively affected by dieldrin, p,p′-

DDE, and PCBs in the environment (Reichel et al.

1984; Wiemeyer et al. 1984). Benthivorous and

piscivorous fish were examined separately because fish

in higher trophic levels (piscivorous fish) generally have

higher concentrations of accumulative contaminants

such as total PCBs and mercury, but benthivorous fish

species are the preferred prey in some bald eagle

populations (Haywood and Ohmart 1986).

Results

Environmental contaminant concentrations

Organochlorine concentrations were less than the

detection limits in most samples (Table 5). Transnonachlor, p,p′-DDE, p,p′-DDD, total PCBs, and

TCDD-EQ were the most frequently detected compounds. Higher detection limits for the Mississippi,

Rio Grande, and Columbia River Basins than for

those other basins sampled influenced mean and

median concentrations of many organochlorine residues (Table 5). Mean concentrations ranged from

0.00007 to 0.08089 μg/g, and median concentration

ranged from 0.00007 to 0.0723 μg/g (most <0.01 μg/

g). Mean and median concentrations were greatest for

total chlordanes, p,p′-DDE, total DDT, total PCBs,

and toxaphene (Table 5). The 85th percentile ranged

from 0.00021 μg/g for pentachlorobenzene and

heptachlor to 0.450 μg/g for total PCBs and were

>0.02 μg/g for trans-nonachlor, total chlordanes, p,p′-

DDE, p,p′-DDD, total DDT, total PCBs, and toxaphene (Table 5). Maximum concentrations ranged

from 0.00111 μg/g for aldrin to 11.37 μg/g for total

DDT but were <0.08 μg/g for most organochlorine

residues (Table 5). The median TCDD-EQ concentration was 0.0000017 μg/g (Table 5). Mean concentrations of select organochlorine pesticides and total

PCBs were less than those reported in fish from the

historical NCBP (Table 5). Mean concentrations of

Dacthal and heptachlor were tenfold lower than those

measured in NCBP samples, and mean concentrations

of chlordane components (oxychlordane, cis- and

trans-chlordanes and nonachlors) also were lower in

BEST-LRMN samples (0.002−0.005 μg/g) than

NCBP samples (0.007−0.011 μg/g; Table 5). Mean

p,p′-homolog concentrations (0.003–0.045 μg/g) were

lower than those reported in NCBP samples (0.010

−0.070 μg/g; Table 5). Concentrations of total PCBs

and toxaphene also were lower in BEST-LRMN

samples compared to those in NCBP samples.

Table 4 Detection limits (μg/g wet weight) for elemental contaminants

Analyte Mississippi River

Basin (1995)

Rio Grande and Columbia

River Basins (1997)

Yukon River

Basin (2002)

Colorado River

Basin (2003)

Mobile, Apalachicola, Savannah,

and Pee Dee River Basins (2004)

Arsenic 0.11−0.27 0.03−0.32 NA NA 0.01

Cadmium 0.02−0.05 0.02−0.06 0.02−0.03 0.02−0.03 0.01

Chromium 0.16 NA 0.22−0.27 0.24−0.35 NA

Copper NA NA NA NA NA

Mercury 0.04−0.05 0.02−0.06 NA NA NA

Nickel 0.16−0.26 0.22−0.33 0.21−0.27 0.24−0.35 0.01

Lead 0.01−0.04 0.03−0.14 0.21−0.27 0.24−0.35 0.01

Selenium 0.12−0.18 0.22−0.27 NA NA NA

Zinc NA NA NA NA NA

Study year is shown in parentheses

NA All concentrations were greater than the detection limit

Environ Monit Assess (2009) 152:469–494 477
Elemental contaminant concentrations exceeded detection limits in most samples (Table 6). Mean concentrations ranged from 0.03 μg/g for cadmium to

35.2 μg/g for zinc, and median concentrations ranged

from 0.04 to 36.0 μg/g (Table 6). The 85th percentile

was <1.0 μg/g for arsenic, cadmium, mercury, nickel,

lead, and selenium and >1.0 μg/g for chromium,

copper, and zinc. Maximum concentrations ranged

from 0.51 μg/g for cadmium to 150 μg/g for zinc but

were <10 μg/g for most elemental contaminants

(Table 6). Mean concentrations of arsenic, cadmium,

copper, mercury, lead, selenium, and zinc were greater

Table 5 Organochlorine chemical residue concentrations (μg/g wet weight) in whole-body fish composite samples from 1995 to 2004

Analyte 1995−2004 NCBP

No.>DL/n Mean Median 85th Maximum Mean

Lipid (%) 409/409 4.89 4.96 8.99 18.10 N/A

Pentachlorobenzene 56/135 0.00007 0.00007 0.00021 0.00119 N/A

Hexachlorobenzene 110/409 0.00277 0.01 0.01 0.068 <0.01

Pentachloroanisole 113/135 0.00056 0.00057 0.00258 0.021 <0.01

α-HCH 62/409 0.00146 0.01 0.01 0.01 0.005

β-HCH 55/409 0.00192 0.01 0.01 0.01 N/A

γ-HCH 40/409 0.00178 0.01 0.01 0.01 <0.01

δ-HCH 54/409 0.00134 0.01 0.01 0.01 N/A

Aldrin 38/135 0.00012 0.00019 0.00027 0.00111 N/A

Dieldrin 198/409 0.00411 0.01 0.017 0.250 0.008

Endrin 76/409 0.00215 0.01 0.01 0.71 0.005

Dacthal 46/135 0.00025 0.00032 0.00070 0.00928 0.007

Heptachlor 14/135 0.00009 0.00019 0.00021 0.00382 0.007

Heptachlor epoxide 115/409 0.00189 0.01 0.01 0.075 N/A

Oxychlordane 116/409 0.00229 0.01 0.01 0.028 0.007

cis-chlordane 175/409 0.00374 0.01 0.013 0.120 0.011

trans-chlordane 150/409 0.00307 0.01 0.01 0.35 0.008

cis-nonachlor 172/409 0.00351 0.01 0.01 0.047 0.009

trans-nonachlor 229/409 0.00548 0.01 0.02 0.31 0.013

ΣChlordane N/A 0.0236 0.03 0.057 0.529 N/A

o,p’-DDE 89/409 0.00238 0.01 0.01 0.022 N/A

o,p’-DDD 99/409 0.00283 0.01 0.01 0.34 N/A

o,p’-DDT 67/409 0.00189 0.01 0.01 0.24 N/A

p,p’-DDE 382/409 0.0447 0.05 0.29 8.30 0.070

p,p’-DDD 250/409 0.00907 0.01 0.04 2.80 0.019

p,p’-DDT 88/409 0.00332 0.01 0.01 0.31 0.010

ΣDDT N/A 0.0809 0.07 0.38 11.37 0.110a

Endosulfan I 13/135 0.00011 0.00021 0.00037 0.00390 N/A

Endosulfan II 47/135 0.00015 0.00017 0.00050 0.05400 N/A

Endosulfan sulfate 80/135 0.00054 0.00079 0.00220 0.07900 N/A

Methoxychlor 15/135 0.00021 0.00035 0.00180 0.00960 N/A

Mirex 118/409 0.00183 0.01 0.01 0.08 0.006

PCB 256/409 0.084 0.067 0.450 3.30 0.335

Toxaphene 83/409 0.03 0.05 0.05 0.83 0.066

TCDD-EQ 255/409 0.0000015 0.0000017 0.000007 0.0000680 N/A

Number of samples with a concentration greater than the detection limit (DL), geometric mean, median, and 85th percentile

concentrations are presented. The geometric mean from 1986 National Contaminant Biomonitoring Program (NCBP) samples also is

provided for comparison (Schmitt et al. 1999). ΣChlordane is the sum of heptachlor epoxide, oxychlordane, cis-chlordane, transchlordane, cis-nonachlor, and trans-nonachlor, and ΣDDT is the sum of o,p’- and p,p′-DDE, DDD, and DDT. Censored values were

represented by one half the detection limit in the summed values

N/A not applicable

478 Environ Monit Assess (2009) 152:469–494
than those reported in fish from the historical NCBP

(Table 6). Most differences in mean concentrations

between BEST-LRMN samples and NCBP samples

were <0.04 μg/g, but mean differences for selenium

(0.17 μg/g) and zinc (14.0 μg/g) were greater (Table 6).

The 85th percentile concentrations of arsenic, cadmium, mercury, lead, selenium, and zinc also were

greater than those reported in fish from the NCBP

(Table 6). Differences in mean concentrations of

arsenic, cadmium, mercury, and lead were <0.10 μg/g,

but differences in copper (0.44 μg/g), selenium

(0.32 μg/g), and zinc (54.8 μg/g) were greater (Table 6).

Toxicity thresholds

Concentrations in whole-body fish samples exceeded

literature-based toxicity thresholds for dieldrin, total

chlordane, total DDT, total PCBs, toxaphene, TCDDEQ, cadmium, chromium, mercury, lead, selenium,

and zinc (Table 7). Concentrations of hexachlorobenzene, arsenic, and copper did not exceed toxicity

thresholds for fish and wildlife at any sampling location.

Toxicity thresholds meeting our criteria were not found

for pentachlorobenzene, pentachloroanisole, α-hexachlorocyclohexane (HCH), β-HCH, δ-HCH, aldrin,

endrin, Dacthal, endosulfans, methoxychlor, mirex, or

nickel (Table 7). Concentrations of some contaminants

differed between trophic levels. Concentrations of

dieldrin, endrin, total chlordane, total DDT, total PCBs,

toxaphene, TCDD-EQ, chromium, and selenium generally were similar in benthivorous and piscivorous

fish (Fig. 2). Concentrations of endrin, total DDT, and

toxaphene in Mississippi River Basin samples were

greater in benthivores compared to piscivores (Fig. 2).

Concentrations of cadmium, lead, and zinc were

greater in benthivores than in piscivores, but mercury

was greater in piscivorous fish than in benthivorous

fish (Fig. 2).

The greatest concentrations and therefore the

greatest exceedences of toxicity thresholds were

found in Mississippi River Basin samples. Toxicity

thresholds for dieldrin and total chlordane were

exceeded by only a few samples from the Mississippi

River Basin (Fig. 2); carp samples from Sites 76 and

206 exceeded thresholds for both of these contaminants. Total DDT concentrations exceeded toxicity

thresholds in 126 samples representing all basins

except the Yukon, Savannah, and Pee Dee River

Basins and were greatest in samples from the

Columbia, Colorado, Rio Grande, and Mississippi

River Basins (Fig. 2). A total of 172 samples

representing all basins except the Rio Grande and

Yukon River Basins exceeded protective thresholds

for total PCBs. Total PCB concentrations were greatest in samples from the Columbia, Colorado, Mississippi, Mobile, and Apalachicola River Basins

(Fig. 2). Toxaphene concentrations exceeded toxicity

thresholds in 42 samples representing all basins

except the Savannah River Basin but were generally

greatest in samples from the Columbia, Rio Grande,

and Mississippi River Basins (Fig. 2). TCDD-EQ

concentrations exceeded toxicity thresholds for fish

Table 6 Elemental contaminant concentrations (μg/g wet weight) in whole-body fish composite samples from 1995 to 2004

Analyte 1995−2004 NCBP

No.>DL/n Mean Median 85th Maximum Mean 85th

Moisture (%) 409/409 72.2 72.6 75.2 79.9 N/A N/A

Arsenic 233/409 0.11 0.19 0.27 1.95 0.08 0.24

Cadmium 180/409 0.03 0.04 0.09 0.51 0.01 0.04

Chromium 393/409 0.76 0.60 2.38 71.8 N/A N/A

Copper 409/409 0.80 0.86 1.26 3.92 0.79 1.7

Mercury 409/409 0.13 0.13 0.27 0.78 0.09 0.18

Nickel 274/409 0.30 0.30 0.73 5.59 N/A N/A

Lead 233/409 0.07 0.10 0.27 9.29 0.06 0.21

Selenium 398/409 0.59 0.53 0.98 4.66 0.42 0.66

Zinc 409/409 35.2 36.0 76.0 150 21.2 31.7

Number of samples with a concentration greater than the detection limit (DL), geometric mean, median, and 85th percentile concentrations

are presented. The geometric mean and 85th percentiles from 1986 NCBP samples are provided for comparison (Schmitt et al. 1999)

N/A Not applicable

Environ Monit Assess (2009) 152:469–494 479
and wildlife in 108 samples representing all basins

except the Yukon and Savannah River Basins and

were greatest in samples from the Columbia, Mississippi, and Mobile River Basins (Fig. 2). Cadmium

concentrations were generally greater in carp and

other benthivores than in piscivores and exceeded

toxicity thresholds in 17 samples representing the

Columbia, Colorado, and Mississippi River Basins

(Fig. 2). Conversely, mercury concentrations were

consistently greater in piscivores and exceeded

toxicity thresholds in 278 samples representing all

basins (Fig. 2). Protective toxicity thresholds for lead

were exceeded by 12 samples representing the

Columbia, Rio Grande, Mississippi, and Apalachicola

River Basins, and concentrations were greatest in

benthivore samples from the Columbia River Basin

(Fig. 2). Selenium concentrations exceeded toxicity

thresholds in 101 samples representing all basins except

the Savannah and Pee Dee River Basins and were

greatest in samples from the Colorado, Rio Grande, and

Mississippi River Basins (Fig. 2). Zinc concentrations

were generally greater in carp and other benthivores

than piscivores, but concentrations in northern pike

were greater than longnose sucker from the Yukon

River Basin. Zinc exceeded toxicity thresholds in 198

samples representing all basins (Fig. 2).

Screening level wildlife risk analysis

Concentrations of dieldrin, endrin, total DDT, total

PCBs, and TCDD-EQ in whole-body fish samples

represented a risk to piscivorous wildlife at one or

more sites. No risk of dieldrin and endrin exposure to

bald eagle and mink was found at most sites. Dieldrin

concentrations at two Mississippi River Basin sites

(Sites 76 and 206; 0.15−0.25 μg/g) exceeded the

NEHC for mink (Figs. 2 and 3), and endrin concentrations in samples from Site 76 (0.22−0.71 μg/g) also

Table 7 Toxicity thresholds for fish and wildlife and the number of samples and sites that exceed these concentrationsa

Contaminant Toxicity threshold

Conc. (μg/g) Reference No. of samplesb No. of sites b

Hexachlorobenzene 0.33 Newell et al. 1987 0 0

γ-HCH 0.1 Newell et al. 1987 0 0

Dieldrin 0.12−5.65 Newell et al. 1987; Shubat & Curtis 1986;

Peakall 1996

5 2

ΣChlordane 0.30 Eisler 1990 5 2

ΣDDT 0.15−3.0 Anderson et al. 1975; Blus 1996;

Beckvar et al. 2005

126 42

PCB 0.11−0.48 Hornshaw et al. 1983; Newell et al. 1987 172 54

Toxaphene 0.035−0.90 Mayer et al. 1975 42 18

TCDD-EQ 0.0000044−0.00003 Nosek et al. 1992; Heaton et al. 1995;

Tillitt et al. 1996; Walker et al. 1996

108 49

Arsenic 2.2−11.6 Gilderhus 1966; McGeachy and Dixon 1992 0 0

Cadmium 0.23−15.6 Spehar 1976; Eisler 1985; Hansen et al. 2002 17 12

Chromium 1.0 Eisler 1986 121 59

Copper 11.1−42.0 Stouthart et al. 1996 0 0

Mercury 0.1−0.3 Barr 1986; Yeardley et al. 1998;

Beckvar et al. 2005;

278 99

Lead 0.4−8.8 Holcombe et al. 1976 12 9

Selenium 0.75−1.0 Lemly 1996; Hamilton 2004 101 43

Zinc 40−64 Spehar 1976 198 97

Toxicity thresholds for dieldrin, toxaphene, arsenic, cadmium, lead, selenium, and zinc were based on whole-body fish concentrations. Toxicity thresholds for hexachlorobenzene, γ-HCH, ΣChlordane, ΣDDT, PCB, TCDD-EQ, and chromium were based

on wildlife toxicity values. Toxicity thresholds were not available for pentachlorobenzene, pentachloroanisole, α-HCH, β-HCH, δ-HCH,

aldrin, endrin, Dacthal, endosulfans, methoxychlor, mirex, and nickel. ΣChlordane is the sum of heptachlor epoxide, oxychlordane, cischlordane, trans-chlordane, cis-nonachlor, and trans-nonachlor, and ΣDDT is the sum of o,p’- and p,p′-DDE, DDD, and DDT

a

Number of samples=409 for all contaminants except for TCDD-EQ (n=408). Number of sites=111

b The number of samples or sites to exceed lowest toxicity threshold available for a contaminant

480 Environ Monit Assess (2009) 152:469–494
YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Endrin (

μg/g)

0.00001

0.0001

0.001

0.01

0.1

1

10

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Total DDT (

μg/g)

0.0001

0.001

0.01

0.1

1

10

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Dieldrin (

μg/g)

0.00001

0.0001

0.001

0.01

0.1

1

10

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Total PCBs (

μg/g)

0.001

0.01

0.1

1

10

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

TCDD-EQ (

μg/g)

0.00000001

0.0000001

0.000001

0.00001

0.0001

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Total chlordane (

μg/g)

0.0001

0.001

0.01

0.1

1

10

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Toxaphene (

μg/g)

0.001

0.01

0.1

1

10

100

Fig. 2 Organochlorine residue and elemental concentrations

(all μg/g wet-weight) in whole-body composite samples of

benthivorous (black circles) and piscivorous (white circles)

fish from the Yukon River Basin (YRB), Columbia River

Basin (CRB), Colorado River Basin (CORB), Rio Grande Basin

(RGB), Mississippi River Basin (MRB), Mobile River Basin

(MORB), Apalachicola River Basin (ARB), Savannah River

Basin (SRB), and Pee Dee River Basin (PRB). Literature-based

toxicity thresholds are represented by dotted lines or boxes (see

Table 7 for reference). No effect hazard concentrations

(NEHCs) for bald eagle (solid line) and mink (dashed line)

are presented for each contaminant (see Table 8 for references

and text for calculation)

Environ Monit Assess (2009) 152:469–494 481
exceeded the NEHC for bald eagle and mink (Fig. 2).

Total DDT concentrations did not represent a risk to

bald eagle at sites from the Yukon, Mobile, Savannah,

and Pee Dee River Basins (Fig. 3). Total DDT

concentrations exceeded the NEHC for bald eagle in

samples from the Columbia River Basin (34 samples

from 12 sites; 0.27−1.41 μg/g), Colorado River Basin

(nine samples from three sites; 0.28−2.80 μg/g), Rio

Grande Basin (12 samples from four sites; 0.30

−1.78 μg/g), Mississippi River Basin (36 samples

from 16 sites; 0.27−11.37 μg/g), and Apalachicola

River Basin (one sample from Site 330; 0.33 μg/g;

Figs. 2 and 3). The risk of total DDT to mink was

lower with few NEHC exceedences. Carp from Sites

201, 202, 203, and 204 in the Mississippi River Basin

(5.49−11.37 μg/g) were the only samples with total

DDT concentrations that exceeded the NEHC for

mink (Figs. 2 and 3). Total PCB concentrations did

not represent a risk to piscivorous wildlife at sites in

the Yukon, Rio Grande, Savannah, and Pee Dee River

Basins (Fig. 3). Total PCB concentrations exceeded

the NEHC for bald eagle in samples from the

Columbia River Basin (two samples from two sites;

0.75−1.3 μg/g), Colorado River Basin (six sample

from two sites; 0.87−2.1 μg/g), Mississippi River

Basin (20 samples from nine sites; 0.75−3.3 μg/g),

Mobile River Basin (four samples from Site 327;

0.95−2.7 μg/g), and Apalachicola River Basin (three YRB CRB CORB RGB MRB MORB ARB SRB PRB Chromium (

μg/g)

0.01

0.1

1

10

100

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Mercury (

μg/g)

0.01

0.1

1

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Lead (

μg/g)

0.001

0.01

0.1

1

10

100

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Cadmium (

μg/g)

0.001

0.01

0.1

1

10

100

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Selenium (

μg/g)

0.1

0.3

1

3

10

YRB

CRB

CORB

RGB

MRB

MORB

ARB

SRBPRB

Zinc (

μg/g)

10

100

1000

Fig. 2 (continued)

482 Environ Monit Assess (2009) 152:469–494
samples from Site 330; 0.94−1.3 μg/g; Figs. 2 and 3).

The risk of total PCB exposure was lower to mink

than bald eagle (Fig. 3). Concentrations of total PCBs

exceeded the NEHC for mink in samples from the

Columbia River Basin (one sample from Site 503;

1.3 μg/g), Colorado River Basin (four samples from

two sites; 1.1−2.1 μg/g), Mississippi River Basin (12

samples from five sites; 1.1−3.3 μg/g), Mobile River

Basin (three samples from Site 327; 1.5−2.7 μg/g),

and Apalachicola River Basin (one sample from Site

330; 1.3 μg/g; Figs. 2 and 3). No risk of TCDD-EQ

exposure to bald eagle and mink was found at sites in

the Yukon or Savannah River Basins (Fig. 3). TCDDEQ concentrations exceeded the NEHC for bald eagle

in samples from the Columbia River Basin (one

sample from Site 96; 0.000043 μg/g), Mississippi

River Basin (29 samples from 18 sites; 0.000017−

0.000068 μg/g), and Mobile River Basin (one sample

from Site 327; 0.000034 μg/g; Figs. 2 and 3). The

risk of TCDD-EQ exposure was greater to mink

than bald eagle (Fig. 3). TCDD-EQ concentrations

exceeded the NEHC for mink in samples from the

Columbia River Basin (38 samples from 15 sites;

0.000002−0.000043 μg/g), Colorado River Basin (six

samples from three sites; 0.000002−0.000006 μg/g),

Rio Grande Basin (25 samples from 10 sites;

0.000002−0.000006 μg/g), Mississippi River Basin

(106 samples from 42 sites; 0.000002−0.000068 μg/g),

Mobile River Basin (eight samples from four sites;

0.000002−0.000034 μg/g), Apalachicola River Basin

(one sample from Site 330; 0.0000046 μg/g), and Pee

Dee River Basin (one sample from Site 337;

0.000014 μg/g; Figs. 2 and 3). A NOAEL TRV was not

available for pentachlorobenzene, pentachloroanisole,

aldrin, Dacthal, methoxychlor, mirex, or toxaphene in

bald eagle and hexachlorobenzene, pentachloroanisole, Dacthal, and mirex in mink (Table 8); therefore,

the risk of these contaminants is unknown.

Chromium, mercury, selenium, and zinc concentrations in whole-body fish samples represented a risk

to piscivorous wildlife at one or more sites. Chromium represented a risk to piscivorous wildlife only in

the Rio Grande Basin, where concentrations were

comparatively high; they exceeded the NEHC for

bald eagle, mink, or both in Rio Grande samples from

Sites 63 (71.8 μg/g), 511 (34.0 μg/g), 512 (18.3 μg/g),

513 (21.1−70.2 μg/g), and 515 (21.8 μg/g; Figs. 2

and 4). No risk of mercury exposure to bald eagle was

associated with samples from the Colorado, Rio

Grande, or Mississippi River Basins (Fig. 2). Mercury

concentrations exceeded the NEHC for bald eagle in

the Yukon River Basin (two samples from two sites;

0.56−0.65 μg/g), Columbia River Basin (one sample

from Site 43; 0.61 μg/g), Mobile River Basin (five

samples from three sites; 0.52−0.78 μg/g), and

Apalachicola-Flint-Chattahoochee River Basin (two

samples from Site 332; 0.51−0.65 μg/g, Savannah

River Basin (two samples from Site 334; 0.63−

0.67 μg/g), and Pee Dee River Basin (two samples

from Site 338; 0.65−0.78 μg/g; Figs. 2 and 4). For

mink, samples from the Yukon River Basin (five

samples from three sites; 0.37−0.65 μg/g), Columbia

River Basin (six samples from four sites; 0.38−

0.61 μg/g), Rio Grande Basin (two samples from

two sites; 0.40−0.46 μg/g), Mississippi River Basin

(four samples from three sites; 0.39−0.45 μg/g),

Mobile River Basin (seven samples from four sites;

0.40−0.78 μg/g), Apalachicola River Basin (two

samples from Site 332; 0.51−0.65 μg/g), Savannah

River Basin (four samples from three sites; 0.43−

0.67 μg/g), and Pee Dee River Basin (four samples

from two sites; 0.37−0.78 μg/g) exceeded the NEHC

for mercury (Figs. 2 and 4). Selenium did not represent

a risk to piscivorous wildlife at any site in the Yukon,

Columbia, Apalachicola, Savannah, or Pee Dee River

Basins (Figs. 2 and 4). Selenium concentrations

exceeded the NEHC for bald eagle at one site in the

Mississippi River Basin (Site 77; 4.11−4.66 μg/g) and

for mink in samples from the Colorado River Basin

(34 samples from 12 sites; 1.29−2.95 μg/g), Rio

Grande Basin (six samples from three sites; 1.24−

1.87 μg/g), Mississippi River Basin (ten samples

from three sites; 1.18−4.66 μg/g), and the Mobile

River Basin (one sample from Site 329; 1.29 μg/g;

Figs. 2 and 4). No risk of zinc exposure to piscivorous wildlife was found at most sites; zinc concentrations in one Mississippi River Basin carp sample

(Site 79; 150 μg/g) represented a risk to bald eagle

(Fig. 2). Concentrations of arsenic, cadmium, copper,

nickel, and lead did not represent a risk to wildlife at

any site.

Discussion

Overall, mean concentrations of organochlorine residues were less than historical NCBP concentrations;

concentrations of arsenic, cadmium, copper, mercury,

Environ Monit Assess (2009) 152:469–494 483
lead, selenium, and zinc were greater than NCBP

concentrations. Direct comparison of these datasets is

precluded by collection location and species composition differences, which can bias mean concentrations. Concentrations of organochlorine pesticides and

total PCBs were expected to decline because their use

has been banned or regulated in the United States.

However, high concentrations of organochlorine

chemical residues remain evident in fish near point

sources such as former manufacturing facilities and

areas of high historical use. Elemental contaminants

emanating from point sources have also been regulated, as have many pesticides containing potentially

toxic trace elements. However, elemental contaminants also are natural constituents of the earth’s crust;

some are essential trace elements for terrestrial and

aquatic organisms. In spite of regulation, elemental

concentrations in fish remain elevated in some areas

due to irrigated agriculture (Hamilton 2004), mining

(Schmitt et al. 2006), and the combustion of fossil

fuels (Baumann and Gillespie 1986; Yeardley et al.

1998), and both organochlorine and elemental contaminants are distributed globally by atmospheric

transport. Considered independently, concentrations

of most organochlorine residues and elemental contaminants we evaluated represented no or low risk to

bald eagle and mink. Among the contaminants

evaluated, the greatest risk to bald eagle and mink

was associated with total DDT, total PCBs, TCDDEQ, mercury, and selenium. The risk of pentachloroanisole, aldrin, Dacthal, methoxychlor, mirex, or

toxaphene to bald eagle or mink was unknown;

NEHCs could not be calculated because NOAELS

were not available for these compounds.

Dieldrin – Bald eagle Dieldrin – Mink

Total DDT – Bald eagle Total DDT – Mink

Fig. 3 Sites that exceed the no effects hazard concentration (NEHC) for selected organochlorine residues. See Fig. 1 for specific site

locations and Table 8 for NEHC concentrations

484 Environ Monit Assess (2009) 152:469–494
Environmental residues of DDT and its degradation products persist in many areas of historic use,

primarily in agricultural areas and near manufacturing and formulation sites, even though the insecticide has been banned in the United States since

1972. Total DDT concentrations remained high in

whole-body fish samples from multiple sampling

sites in our study and represented a greater risk to

bald eagle than mink. Total DDT concentrations

exceeded the NEHC for bald eagle in at least one

sample from all basins except the Yukon and

Savannah River Basins, and the risk was greatest

at sites in agricultural areas of the lower Columbia,

Mississippi, Rio Grande and Colorado River Basins.

Other studies have reported that DDT concentrations

in these regions continue to represent a risk to avian

wildlife (Grubb et al. 1990; Henny et al. 2004; Buck

et al. 2005). Concentrations of p,p′-DDE were

associated with reproductive effects including eggshell thinning and lower productivity in osprey and

bald eagle populations in the lower Columbia River

Basin (Henny et al. 2004; Buck et al. 2005).

Conversely, DDT and other contaminants did not

affect bald eagle populations nesting along the Salt

and Verde Rivers (Grubb et al. 1990), upstream of

where our study indicated high risk of total DDT

exposure to bald eagles in the Colorado River Basin.

Other studies of bald eagles from our study area were

not found in the scientific literature. Our screening

level risk analysis indicated that bald eagles were at

greatest risk to total DDT in winter nesting areas in

the lower Mississippi River Basin (i.e., Arkansas and

Mississippi). Mink were also at risk to total DDT

exposure in the lower Mississippi River Valley at

sites in watersheds farmed for cotton. We are not

aware of specific studies evaluating total DDT

EXPLANATION

< NEHC

1-10X NEHC

> 10X NEHC

Total PCBs – Bald eagle Total PCBs – Mink

TCDD-EQ – Bald eagle TCDD-EQ – Mink

Fig. 3 (continued)

Environ Monit Assess (2009) 152:469–494 485
Table 8 Toxicity reference value (TRV) for bald eagle and mink based on the no observed adverse effect level from dietary or tissue

studiesa

, no effects hazard concentration (NEHC)b, and the number of samples and sites that exceed the NEHCc

Contaminant Bald eagle Mink

TRV

(mg/kg/d)

NEHC

(μg/g)

No. of

samples

No. of

sites

TRV

(mg/kg/d)

NEHC

(μg/g)

No. of

samples

No. of

sites

PentaCB N/A N/A N/A N/A 7.25d 53.3 0 0

HCB 0.225e 2.25 0 0 N/A N/A N/A N/A

α-HCH 0.56f 5.6 0 0 0.014f 0.10 0 0

β-HCH 0.56f 5.6 0 0 0.31f 2.28 0 0

γ-HCH 2f 20 0 0 6.15f 45.2 0 0

δ-HCH 0.56f 5.6 0 0 0.014f 0.10 0 0

Aldrin N/A N/A N/A N/A 0.154f 1.13 0 0

Dieldrin 0.077f 0.77 0 0 0.015f 0.11 6 4

Endrin 0.010f 0.10 5 2 0.038f 0.28 2 1

ΣChlordane 2.14f 21.4 0 0 1.9f 14.0 0 0

ΣDDT 6g 0.27 92 36 0.62f 4.56 5 4

Endosulfans 10f 100 0 0 0.12f 0.88 0 0

PCB 20g 0.71 35 15 0.137f 1.01 21 10

Methoxychlor N/A N/A N/A N/A 3.1f 22.8 0 0

Toxaphene N/A N/A N/A N/A 6.2f 45.6 0 0

TCDD-EQ 0.000303g 0.000016 31 20 0.00000027h 0.0000019 185 76

Arsenic 2.24i 22.4 0 0 1.04i 7.65 0 0

Cadmium 1.47j 14.7 0 0 0.77j 5.66 0 0

Chromium 2.66k 26.6 4 3 2.40k 17.7 7 5

Copper 4.05l 40.5 0 0 5.60l 41.2 0 0

Mercury 0.5m 0.5 14 9 0.05n 0.37 34 22

Nickel 6.71o 67.1 0 0 1.70o 12.5 0 0

Lead 1.63p 16.3 0 0 4.7p 34.6 0 0

Selenium 0.4f 4.0 3 1 0.154f 1.13 51 19

Zinc 14.49f 145 1 1 123f 904 0 0

TRVs were not available for pentachloroanisole, Dacthal, and mirex for either bald eagle or mink. ΣChlordane is the sum of

heptachlor epoxide, oxychlordane, cis-chlordane, trans-chlordane, cis-nonachlor, and trans-nonachlor, and ΣDDT is the sum of o,p′-

and p,p′-DDE, DDD, and DDT

N/A Not available, pentaCB pentachlorobenzene, HCB hexachlorobenzene

a Most TRVs were derived from adult dietary exposures. TRVs for ΣDDT, PCB, TCDD-EQ, and mercury in bald eagle were based on

egg concentrations

b NEHC=dietary-based no adverse effects level TRV (mg/kg body weight/day)/food ingestion rate (kg food/kg body weight/day).

Food ingestion rate=0.100 kg food/kg body weight/day for bald eagle and 0.136 kg food/kg body weight/day for mink (USEPA

1993). NEHC=tissue-based no adverse effects level TRV (mg/kg egg)/biomagnification factor (BMF). The bald eagle BMFs were 28

for PCB, 19 for TCDD-EQ, 22 for ΣDDT, and 1 for mercury (Giesy et al. 1995)

c

Number of samples=409 for all contaminants except for TCDD-EQ (n=408) and pentachlorobenzene, aldrin, endosulfans, and

methoxychlor (n=135). Number of sites=111 except for pentachlorobenzene, aldrin, endosulfans, and methoxychlor (n=37)

d Linder et al. 1980

e

Hill and Camardese 1986

f

Sample et al. 1996

g Elliott and Harris 2001

h Tillitt et al. 1996

i

USEPA 2005a

j

USEPA 2005b

k USEPA 2005c

l

USEPA 2007a

m Giesy et al. 1995

n USEPA 1995

o USEPA 2007b

p USEPA 2005d

486 Environ Monit Assess (2009) 152:469–494
Chromium – Bald eagle Chromium – Mink

Mercury – Bald eagle Mercury – Mink

Selenium – Bald eagle Selenium – Mink

EXPLANATION

< NEHC

1-10X NEHC

> 10X NEHC

Fig. 4 Sites that exceed the no effects hazard concentration (NEHC) for selected elemental contaminants. See Fig. 1 for specific site

locations and Table 8 for NEHC concentrations

Environ Monit Assess (2009) 152:469–494 487
concentrations in piscivorous mammal populations

from this region.

PCBs were used historically as dielectric, hydraulic, and heat transfer fluids; lubricants; and in

carbonless copy paper until the U.S. ban in 1979.

Total PCB concentrations exceeded NEHCs for bald

eagle and mink at multiple sites. Mink are among the

most sensitive organisms to PCBs with their survival

and reproduction affected by small dietary PCB doses

(Leonards et al. 1995). Risk associated with total

PCBs to mink generally was greatest in industrialized

watersheds and near historical production and formulation facilities. NEHCs for both mink and bald eagle

were exceeded at Sites 23, 24, 67, 76, 111, 320, 324,

327, and 330. Elevated PCB concentrations have been

previously documented in water and biota near Site

327 (USFWS 1996; Schmitt et al. 1999; Zappia

2002), and PCBs were identified as a potential cause

of reproductive dysfunction in mink from the coastal

plain of Georgia, South Carolina, and North Carolina

(Osowski et al. 1995). We are not aware of contaminant studies evaluating the risk of total PCBs to

piscivorous wildlife in the Mississippi River Basin

downstream from Memphis (Site 76) or the Ohio

River Valley (Sites 23, 24, and 67). Mink typically do

not occur in the southwestern United States, but other

piscivorous mammals [i.e., raccoon (Procyon lotor)]

may be at risk to total PCBs near Sites 320 and 324.

The Gila River at Site 324 has low or no flow for part

of the year and is recharged by wastewater treatment

plant effluent and urban runoff from the Phoenix area.

The Gila River supports contaminated prey items

including fish with elevated concentrations of organochlorine pesticides and total PCBs (Hinck et al.

2007). Therefore, contaminant exposure to piscivorous wildlife may be increased by limited water and

food resources in this arid region, where riparian

wildlife concentrate. Moreover, bald eagles breeding

in southern Arizona habitually forage in the same

areas of free-flowing water, where fish are the most

commonly consumed prey item (Haywood and

Ohmart 1986; Grubb 1995).

PCBs were formulated as complex mixtures comprising as many as 209 structurally similar compounds (congeners) that vary widely in toxicity and

persistence. The toxicity of PCBs to mink, bald

eagles, and other organisms is largely attributable to

relatively few congeners, specifically the highly toxic

non-ortho- and mono-ortho-chloro substituted compounds that can assume a planar molecular configuration (Aulerich and Ringer 1977; Peterson et al.

1993; Tillitt et al. 1996). These compounds are

structurally similar to 2,3,7,8-tetrachloro-p-dioxin

(TCDD) and are among the compounds collectively

known as planar halogenated hydrocarbons (PHHs).

Concentrations of PHHs in fish can vary over orders

of magnitude depending upon the original composition of the mixture released to the environment,

weathering, and toxicokinetic factors, and they largely

determine the risk assessment of PCBs in the

environment (Safe 1990; van den Berg et al. 2006).

It has been widely accepted that concentrations of

individual congeners, specifically the planar congeners, rather than total PCB concentrations are necessary to assess the risk of PCBs to piscivorous wildlife.

However, recent studies indicate that total PCBs

measurements can be used to assess the toxicological

risks associated with dioxin-like PCB concentrations

in whole-body benthivorous and piscivorous fish

(Bhavsar et al. 2007a, b). High-resolution analyses

required for PHHs were beyond the scope of our

studies; therefore, we determined dioxin-like activity

(i.e., TCDD-EQ) with the H4IIE bioassay. As

measured in our samples, TCDD-EQ includes the

contributions of all the PHHs including the planar

PCB congeners. Dioxin-like activity (as TCDD-EQ)

was elevated at many of the sites where total PCB

concentrations were also comparatively high, indicating that much of the dioxin-like activity was attributable to PCBs. However, the overall risk of TCDD-EQ

to bald eagle and mink was more widespread than risk

of PCBs, indicating some contributions from other

PHHs at sites with high TCDD-EQ concentrations but

low total PCB concentrations. TCDD-EQ-related risk

was greatest in the Columbia, Rio Grande, Mobile, and

Mississippi River Basins, but was minimal in the

Yukon, Colorado, Apalachicola, Savannah, and Pee

Dee River Basins.

The risk of mercury to piscivorous wildlife was

greatest at sites in the southeastern United States

(Mobile, Apalachicola, Savannah, and Pee Dee River

Basins) and Yukon River Basin. Sources of Hg in these

basins include historical gold and cinnabar mines,

releases from chemical manufacturing and coal-fired

power plants, atmospheric deposition, and wetlands

(Rudd 1995; Brumbaugh et al. 2001; Eisler 2004;

Paller et al. 2004; Warner et al. 2005). Orihel et al.

(2007) demonstrated that organic mercury deposited

488 Environ Monit Assess (2009) 152:469–494
directly to aquatic ecosystem was readily converted to

methylmercury, which is available to biota. This

relationship may explain the high fish concentrations

and associated risk of mercury in the southeastern

United States and Yukon River Basin, where mercury

methylation rates are known to be high (Brumbaugh et

al. 2001). Other studies also have reported that

mercury concentrations in these regions represent a

risk to avian and mammalian wildlife (Halbrook et al.

1994; Osowski et al. 1995; Ambrose et al. 2000; Adair

et al. 2003). Mercury has been suspected to cause

reproductive effects, organ toxicity, and mortality in

prothonotary warbler (Protonotaria citrea), mink, and

river otter throughout the southeastern United States

(Halbrook et al. 1994; Osowski et al. 1995; Adair et al.

2003). Ambrose et al. (2000) reported that mercury

concentrations in peregrine falcon (Falco peregrinus

anatum) breeding in the Yukon River Basin were

associated with lower nest success. Conversely, mercury concentrations in bald eagle eggs from the lower

Columbia River Basin were not associated with lower

productivity (Anthony et al. 1993; Buck et al. 2005).

Sources of excess selenium include natural weathering of seleniferous shales, irrigation practices,

uranium ore and coal extraction, and coal-fired power

plants (Sorenson 1991). The risk of selenium to bald

eagle was lower than to mink. Among basins, mink in

the Colorado River Basin were at greatest risk to

selenium exposure. Irrigation and mining throughout

these areas accelerate the rates of the processes

controlling the release and distribution of this naturally occurring element. Selenium exposure to wildlife is most commonly associated with toxicity to

birds (Lemly 1996; Hamilton 2004). Historically,

avian wildlife was impacted by selenium contamination in the Colorado River Basin near Site 312

(Stephens et al. 1992), but we are not aware of

studies evaluating selenium toxicity to piscivorous

mammals in this basin.

The wildlife risk analysis presented here is a

screening level or Tier 1 assessment. As such, it was

based on the most conservative (i.e., lowest) TRVs

available to identify all potential contaminants of

concern (USEPA 1998). The conservative TRVs we

used included values extrapolated from receptors

other than bald eagle and mink. As noted by Sample

and Suter (1999), such an approach indicates only

whether harmful effects are possible, not whether they

are probable. Consequently, the use of the most

conservative TRVs may overestimate the actual risk

of some contaminants to wildlife. For example, we

initially used a TRV for mercury in bald eagle

(0.0064 mg/kg/d) derived from a multi-generational

study of mallards (Anas platyrychos; Heinz 1979;

Sample et al. 1996). Based on this TRV, Yeardley et

al. (1998) estimated that mercury concentrations in

fish from 98% of the lakes in the northeastern United

States represented a risk to piscivorous birds. Our

NEHC calculated from this TRV (0.064 g/g) similarly

indicated that bald eagles were at risk from mercury at

most of our sites (105 of 109 sites), which might be

interpreted as indicative of widespread mercury

poisoning of bald eagles in the conterminous United

States. This conclusions would contradict current data

indicating the recovery of U.S. bald eagle populations

and their recent removal from the Endangered Species

List (U.S. Fish and Wildlife Service 2007), and would

indicate that an alternative TRV should be considered.

Therefore, we selected a NEHC of 0.5 μg/g developed specifically to estimate mercury risk to bald

eagles (Giesy et al. 1995), which resulted in considerably lower estimates of risk to bald eagles at our

sites; only nine sites had mercury concentrations

>0.5 μg/g. However, there is evidence that piscivorous birds are being negatively affected by mercury in

fish elsewhere (e.g., Scheuhammer and Blancher

1994; Scheuhammer et al. 1998; Nocera and Taylor

1998). One might therefore argue that bald eagle

populations would be increasing more rapidly if not

for mercury, which would support the use of the lower

TRVs. This example highlights the limitations of

conservative TRVs and the importance of qualifying

their use in the screening level phase of wildlife risk

analysis. Regardless of the TRV employed, sites at risk

would need to be evaluated further to fully characterize

the exposure and ecological effects of specific contaminants to bald eagles or mink (e.g., Tier 2 risk

assessment).

There are other assumptions and uncertainties

associated with the approach we used to evaluate risk

to piscivorous wildlife. Some of the NOAELs used

here were derived from studies conducted several

decades ago, and other NOAELs have been proposed

and used elsewhere. Consensus NOAELs based on

whole-body fish concentrations, such as those adopted for arsenic, cadmium, chromium, copper, nickel,

and lead, are not available for many contaminants.

The use of other NOAELs would result in different

Environ Monit Assess (2009) 152:469–494 489
NEHCs and could change the perceived risk to

piscivorous wildlife. Other wildlife food chain analyses incorporate water and sediment/soil concentrations to estimate contaminant concentrations in fish

and evaluate the risk of contaminants to piscivorous

wildlife because fish concentrations are not available.

While more recent toxicity thresholds may be

available for water, soil, and sediment, there are

additional uncertainties associated with estimating

exposure to piscivorous wildlife from these endpoints.

Factors such as organic carbon content, sulfides,

hardness, and pH can influence the bioavailability

and toxicity of contaminants (e.g., Björnberg et al.

1988; Chapman et al. 1998; Heekyoung Choi et al.

1998). These points emphasize that updated consensus

toxicity thresholds based on whole-body fish concentrations associated with reproductive performance,

growth, and survival are needed, especially for contaminants such as pentachloroanisole, Dacthal, and mirex

for which toxicity thresholds are not currently available.

However, studies determining toxicity thresholds in

whole-body fish are not likely to be conducted for some

of these contaminants because of time and expense

requirements for multi-generational studies with birds

and mammals and the current focus of research on new

or emerging contaminants.

The selection of appropriate receptors in the

analysis is important. Bald eagles and mink consume

large, adult fish such as those collected and analyzed

in this study and are therefore reasonable receptors.

The risk of contaminants to other piscivorous wildlife

such as belted kingfisher (Ceryle alcyon), Caspian

tern (Sterna caspia), or common loons (Gavia immer)

may be overestimated if these species are used in the

model, because they consume smaller fish that are

presumed to have lower contaminant concentrations

than larger fish because of their shorter lifespans, higher

growth rates, and higher metabolism (Scheuhammer

et al. 1998; Rose et al. 1999). It is also unlikely that the

diet of adult bald eagles or mink would be entirely

composed of fish from one site, which is tacitly

assumed in this approach. Nevertheless, this assumption may be valid for nestling birds or cubs fed locally

procured prey items. Moreover, the risk of accumulative contaminants to nestlings/cubs and juveniles may

be greater because of their small body size and

correspondingly higher food ingestion rates compared

to adult wildlife. Inexperienced juveniles also may

have greater exposure to chemical contaminants

because of their initial small home range or when they

scavenge on dead or dying fish associated with natural

or anthropogenic events.

As a final consideration, this assessment assumes

that the contaminants act independently. Many of the

contaminants co-occur at the sites, and their effects

are not necessarily independent or additive (e.g.,

Wren et al. 1987). Effects also may vary depending

on other factors such as the disease, nutritional, and

reproductive status of the organism. Nevertheless, the

NOAELs and subsequent NEHCs represent conservative, screening-level criteria for the assessment of

potential adverse effects in wildlife from chemical

contaminants in the fish that they consume.

Overall, concentrations of many persistent, bioaccumulative organochlorine chemical residues have declined over the past two decades. Our results indicate

that organochlorine residues, in particular p,p′-DDE and

total PCBs, remain a risk to piscivorous wildlife at

some sites. Concentrations of mercury and selenium,

which have not declined, also represent a risk at a

number of sites. In addition, mercury and selenium

concentrations in aquatic systems may rise because of

increasing global coal combustion and subsequent

atmospheric deposition (USEPA 1997; Mason et al.

1999). The full geographic extent of the risk associated

with contaminants in fish cannot be determined from

our data because some of the sites were not selected in a

truly unbiased manner and only screening level analyses

were conducted. Regardless, our findings indicate that

environmental contaminants are a risk to piscivorous

wildlife at some sites. Characterization of exposure and

ecological effects for contaminants at specific sites

would be the next step in refining the risks determined

in our screening level wildlife risk analysis.

Acknowledgements This study was conducted as a component

of the Large River Monitoring Network of the Biomonitoring

Environmental Status and Trends (BEST) Project. This project is

managed by the Status and Trends Program of the U.S. Geological

Survey. A. Bernhardt, N. Beyer, and G. Linder reviewed earlier

versions of this paper and provided additional information.

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