Organochlorines, Mercury, and Selenium in Great Blue Heron Eggs from Indiana Dunes National Lakeshore, Indiana

J. Great Lakes Res. 24(1):3-11 Internat. Assoc. Great Lakes Res., 1998

Organochlorines, Mercury, and Selenium in Great Blue Heron Eggs from Indiana Dunes National Lakeshore, Indiana Thomas W. Custer 1*, Randy K. Hines 1, Paul M. Stewart2, Mark J. Melancon 3, Diane S. Henshel4, and Daniel W. Sparkss I United

States Geological Survey Biological Resources Division Upper Mississippi Science Center PO. Box 818 La Crosse, Wisconsin 54602 2United States Geological Survey Biological Resources Division Lake Michigan Ecological Research Station 1100 N. Mineral Springs Road Porter, Indiana 46304 3 United

States Geological Survey Biological Resources Division Patuxent Wildlife Research Center Laurel, Maryland 20708 4 School

of Public and Environmental Affairs Indiana University Bloomington, Indiana 47405-2100

5u.S. Fish and Wildlife Service Bloomington Ecological Field Office Bloomington, Indiana 47403 ABSTRACT. In 1993, 20 great blue heron (Ardea herodias; GBH) eggs (one per nest) were collected from a colony at the Indiana Dunes National Lakeshore, Indiana (INDU). The eggs were artifIcially incubated until pipping and were then analyzed for organochlorines, mercury, and selenium. Livers of embryos were analyzed for hepatic microsomal ethoxyresorufln-O-dealkylase (EROD) activity. Brains were measured for asymmetry. Egg-laying began in early April and the mean clutch size was 4.2 eggs per clutch. Organochlorine concentrations were generally low (geometric mean p,p'-DDE = 1.6 f.lg/g wet weight; polychlorinated biphenyl [PCB] = 4.9 f.lg/g); however, one egg had elevated concentrations of p,p' -DDE (13 f.lg/g) and PCBs (56 f.lg/g). EROD activity in the embryos analyzed from INDU was not elevated. The frequency (II %) ol brain asymmetry was low. Eggshells averaged 3.4% thinner than eggshells collected prior to the use of DDT. Mercury (geometric mean = 0.9 f.lg/g dry weight) concentrations in GBH eggs were within background levels. Selenium (4.0 f.lg/g dry weight) concentrations in eggs were above background levels, but below a concentration threshold associated with reproductive impairment. INDEX WORDS: Organochlorines, great blue herons, PCB congeners, mercury, selenium, Indiana Dunes National Lakeshore.

"Corresponding author. E-mail: tom_w_custer@nbs,gov

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Custer et al.

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INTRODUCTION The northwest corner of Indiana is an area formed by the recession of glacial Lake Michigan and includes the significant ecological resources of the Indiana Dunes National Lakeshore (INDU) and Indiana Dunes State Park (Fig. 1). The INDU includes over 6,800 hectares of dune and swale topography that contain over 1,300 species of vascular plants (PAHLS 1993). INDU provides a stopover for more than 330 species of migratory birds and is summer residence to over 110 species of breeding birds. It also serves as a natural transitory stopover for spring and fall migrants (Brock 1986). The northwest corner of Indiana is also an area of poor air, water, and sediment quality (Fig. I; PAHLS 1993, Indiana Department of Environmental Management 1988). Contaminants have been released from municipal, industrial, and non-point discharges, including persistent organochlorines, petroleum products, heavy metals, and sewage effluents (Hoke et al. 1993, International Joint Commission 1985). Few data exist on environmental contaminant levels in the wildlife that inhabit INDU. Steffeck (1989) found elevated levels of mercury (0.354 f.,lg/g wet weight) in snapping turtles (Chelydra serpentina) and elevated selenium levels (2.0 f.,lg/g wet weight) for earthworms (Lumbricus rubellus and Dendrobaena octaedra) collected within INDU. Elevated contaminant concentrations in fish populations have resulted in fish consumption advisories for polychlorinated biphenyls (PCBs) and mercury for Lake Michigan and inland tributaries (Indiana State Department of Health 1997). No data are available on contaminant levels in birds from the INDU. A breeding colony of approximately 120 pairs of great blue herons (Ardea herodias; GBH) occurs on the INDU (Randy Knutson, Park Resource Management, INDU, personal communication, Fig. 1). The colony has been active since the early 1950s (Taylor et al. 1982). GBHs in other portions of their range accumulate residues (Custer et al. 1997) and are sensitive to environmental contaminants. Concentrations of 2,3,7,8-tetrachlorodibenzo-p-dioxin, the most toxic of the polychlorinated dibenzodioxins, observed in GBH embryos from the Strait of Georgia, British Columbia, were correlated with ethoxyresorufin-O-dealkylase (EROD) activity (Bellward et al. 1990), depressed embryonic growth (Hart et al. 1991), and brain asymmetry (Henshel et al. 1993). Our objective was to evaluate the conta-

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FIG. 1. Map of northwest Indiana showing location of the great blue heron colony in relationship to municipal landfills, incinerators, superfund sites, and Comprehensive Environmental Response, Compensation, and Liability Information System (CERCLIS) sites. Adapted with permissionfrom PAHLS (1993).

minant exposure of GBHs nesting in INDU by determining contaminant concentrations in eggs collected from the GBH colony. METHODS On 28 April 1993, 20 GBH eggs (one egg per nest) were collected from the INDU colony located in Porter County, Indiana (Fig. 1; 41 N 30'43", 86 W 52.5'33") under appropriate state and federal permits. A professional tree climber ascended two 30 m beech trees (Fagus grandifolia) and used a 2 to 8 m extendable pole with a nylon stocking cup attached to the end of the pole to collect eggs that were out of reach (Hines and Custer 1995). After collection, the eggs were placed in foam-lined containers, lowered to the ground, placed in a temperature-controlled (37°C) portable incubator, and then transported to the Upper Mississippi Science Center, LaCrosse, Wisconsin. Eggs were then candled for viability, randomly placed on shelves in an incu-

Contaminants in Great Blue Heron Eggs bator (37.5°C, 60-65% humidity), and checked at least daily for signs of pipping. At pipping, the embryos in all viable eggs were removed from the shells, examined for deformities, and weighed (± 0.1 g). Immediately following death by decapitation, the liver was removed and weighed (± 0.001 g). Portions « 1 g) of the liver of all embryos were placed into two cryotubes, quick frozen in liquid nitrogen, stored in a Revco Ultra Low Freezer at -85°C, and within 6 months analyzed for cytochrome P450-associated monooxygenase activities. Eggs were randomly divided into two groups; 10 eggs for chemical analyses and 10 eggs were used for anatomical measurements of embryos. Contents of eggs that did not pip in the anatomical subsample and all embryos and egg contents in the chemical analysis subsample were placed in chemically clean jars and frozen at -20°e. Samples within the chemical analysis group included the entire contents of eggs, except that livers were removed from embryos at pipping. Within the anatomical subsample, embryos that survived to pipping were decapitated. After decapitation, the skull caps of the heads in the anatomical group were opened and the heads placed in a 10% buffered formalin solution and refrigerated. Date of egg-laying for eggs in 16 nests was estimated based on pipping date. In order to estimate the date when the eggs were laid, day of laying was assigned day 0, day of pipping was day 27, and day of hatching was day 28 (Pratt 1970). Estimates for fledging (first flight) were made by adding 60 days to the hatching date based on a California study showing the youngest age at which GBHs flew (Pratt 1970). Eggshell thickness was measured to the nearest 0.01 mm with a micrometer after the shells had dried at room temperature for at least 1 month. A mean thickness value was derived from three measurements taken at the equator of each egg and included the shell and shell membranes. The following organochlorines were analyzed in GBH eggs by Mississippi State Chemical Laboratory, Mississippi State University, Mississippi State, Mississippi: aldrin; a-, 13-, y- and D-benzene hexachloride (HCH); a- and y-chlordane; oxychlordane; cis-nonachlor; trans-nonachlor; dieldrin; endrin; hexachlorobenzene (HCB); heptachlor epoxide; mirex; toxaphene; o,p' -DDD; o,p'-DDE; o,p'-DDT; p,p'-DDD; p,p'-DDE (DDE); p,p'-DDT; total PCBs; and PCB congeners with Ballschmiter-Zell (BZ) numbers 77, 105, 114, 118, 126, 156, and 169. Total PCBs were estimated based on Aroclor equivalents.

5

Concentrations of PCB 118 may be overestimated because it coeluted with PCB 106. Samples were homogenized, mixed with sodium sulfate, and soxhlet extracted with hexane. After the percent lipid was determined, lipids were removed by Florisil column chromatography and the eluate was subjected to further cleanup on a silicic acid column to separate PCBs from other polar organochlorines. Following silicic acid column chromatography, pesticides and total PCBs were determined by electron capture gas chromatography. Samples for specific PCB congeners were prepared as above, but the PCB fraction from the silicic acid column was further fractionated using an AX-21 carbon on an activated silica gel column to isolate non-ortho and mono-ortho PCBs. PCB congeners were analyzed by electron capture detection. The nominal limits of detection for organochlorines were 0.0001 flg/g wet weight for PCB congeners, 0.05 flg/g wet weight for total PCBs and toxaphene, and 0.01 flg/g wet weight for the remainder of the organochlorines. Duplicate analyses, spikes, blanks, and GC/Mass Spectrometry confirmation analyses were conducted on 10% of the total number of samples analyzed. Egg contents were analyzed for mercury and selenium at Research Triangle Institute, Research Triangle Park, North Carolina. Samples were freeze-dried for determination of moisture and then ground prior to nitric acid digestion in a capped Teflon vessel using a CEM microwave oven. Selenium concentrations were determined using graphite furnace atomic absorption with either a Perkin-Elmer Zeeman 3030 or 4100ZL atomic absorption spectrometer. Mercury concentrations were determined using cold vapor atomic absorption with a Leeman PS200 Hg analyzer. Nominal limits of detection were 0.1 flg/g dry weight mercury and 0.5 flg/g dry weight selenium. Duplicate analyses, spikes, and blanks were conducted on 10% of the total number of samples analyzed. Two standard reference materials, dogfish liver and dogfish muscle, were analyzed concurrently for concentrations of mercury and selenium. All results for mercury and selenium are reported on a dry weight basis. Mean moisture content was 83.7%. Quality assurance and quality control of the organochlorine and trace element analyses were approved by the Patuxent Analytical Control Facility, U.S. Fish and Wildlife Service, Laurel, Maryland. The toxic potency of the aryl hydrocarbon-active (Ah-active) PCB congeners relative to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) was estimated for each sample by summing the products of the mea-

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Custer et al.

sured concentrations and TCDD toxic equivalency factors (TEFs) as suggested by Kennedy et al. (1996) and Safe (1990). These two methods were selected in order to make comparisons with earlier studies (Rattner et al. 1994, Custer et al. 1997). The TEFs of Kennedy et al. (1996) are based upon EROD activity measured in cultured chicken embryo hepatocytes. The Safe TEFs are based upon toxic responses in mammals. Estimated toxicity is expressed as toxic equivalents (TEQs, pg/g of sample). Cytochrome P450-associated monooxygenase activities were assayed at Patuxent Wildlife Research Center, Laurel, Maryland. Methods are described in detail in Melancon (1996). Individual liver samples were homogenized and hepatic microsomes were prepared by differential centrifugation. Cytochrome P450-associated monooxygenase assays included the fluorometric determination of EROD. EROD activity in GBH embryos in this study was compared to values from a reference colony located near MacDougall, Minnesota (Custer et al. 1997). Eggs from the MacDougall colony had the lowest PCB concentrations (geometric mean = 1.0 Ilg/g total PCBs) among ten colonies on the Upper Mississippi River. Embryo livers in this study and those from the MacDougall colony were assayed concurrently for EROD activity. Measurements of brain morphometry were made as described in Henshel et al. (1993, 1995). The width, angle, height, and depth of the right and left side of each brain were measured using an engineering ruler with 0.5 mm gradations. If the two sides of the brain had the same measurement, the sample was classified as symmetrical. If the two measurements differed by 0.25mm or more the brain was classified as asymmetrical. Contaminant concentrations were log transformed, using base 10 logarithms, in order to make the data comparable with earlier reported measurements (Custer et al. 1997). Means were calculated only when half or more of the samples had detectable concentrations. For purposes of analysis, samples below the detection value were given one-half the detection limit for use in calculation of the means. RESULTS Oxychlordane, cis-nonachlor, trans-nonachlor, dieldrin, heptachlor epoxide, p,p' -DDE, total PCBs, and PCB congeners 105, 114, 1181106, and 156 were detected in all ten samples from INDU (Table 1). Endrin, hexachlorobenzene, mirex, toxaphene,

p,p'-DDD, p,p'-DDT, and PCB congener 126 were detected in five or more samples, while f)-HCH, a-chlordane, and PCB congeners 77 and 169 were detected in less than half of the samples. The following organic compounds were not detected: a-, y-, and b-HCH; b-chlordane; o,p'-DDD; o,p'DDE; and o,p'-DDT. Mean concentrations of p,p'DDE and PCBs were 1.6 and 4.9 Ilg/g, respectively (Table 1). One egg had elevated p,p'-DDE (13.0 Ilg/g) and PCB (56 Ilg/g) concentrations. The geometric mean toxic equivalents were 465 pg/g (arithmetic mean = 958 pg/g) based on Safe (1990) TEFs and 929 pg/g (arithmetic mean = 2,147 pg/g) based on Kennedy et al. (1996) TEFs (Table 1). The seven PCB congeners measured here accounted for 8.0 percent of the total PCBs. Mercury and selenium were detected in all ten of the eggs analyzed. Concentrations of mercury averaged 0.91 Ilg/g dry weight and one sample reached 4.6 Ilg/g dry weight (Table 1). Selenium concentrations averaged 3.98 Ilg/g dry weight. All selenium samples exceeded 2.0 Ilg/g dry weight, with six out of ten exceeding 4.0 Ilg/g dry weight. Eggshells of INDU herons (mean = 0.380 mm ± 0.007 SE [range 0.303-0.420], Table 2) were 3.4% thinner than eggs collected pre-1947 from southern Canada (mean = 0.393 mm; Anderson and Hickey 1972). EROD activity in livers from embryos in this study was not significantly different (F = 1.08, df = 1, P = 0.31) than a value from a reference location in MacDougall, Minnesota (Table 2). The width, angle, and depth of 1 of 9 (11 %) embryo brains were asymmetrical (0.25 mm); none of the brains were asymmetrical for height. Sixteen of the 20 eggs (80%) survived until pipping and no abnormal embryos were observed. Eggs hatched in the field (n = 1) and in the laboratory incubator (n = 15) between 29 April and 17 May; the median date of pipping was 6 May. We estimated that egg-laying in 16 clutches began between 1 and 19 April; the median date of first egg laid was 8 April. Estimated fledging (first flight) dates of the 16 eggs varied from 31 May through 18 June with a median fledgling date of 7 June. Clutch size varied from 3 to 5 eggs per nesting attempt and averaged 4.2 eggs (four with 3 eggs, eight with 4 eggs, and eight with 5 eggs). DISCUSSION Total PCB concentrations in GBH eggs (geometric mean = 4.9 Ilg/g) collected from INDU were intermediate compared to those found in GBH eggs

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Contaminants in Great Blue Heron Eggs TABLE 1. Concentrations of organochlorines, mercury, and selenium in 10 great blue heron eggs from Indiana Dunes National Lakeshore, 1993.

TABLE 2. Ethoxyresorufin-O-dealkylase activity in great blue heron embryos from a reference location (MacDougall, Minnesota) and the Indiana Dunes National Lakeshore, 1993. Ethoxyresorufin-O-dealkylase activity (pmol/min/mg protein)

Concentration (flg/g)a Chemical

Geometric mean

Range - b 9NDc-0.08 f3-benzene hexachloride a-chlordane 7ND-0.03 0.02-0.45 Oxychlordane 0.07 Cis-nonachlor 0.07 0.01-0.48 Trans-nonachlor 0.19 0.03-1.10 Dieldrin 0.26 0.05-0.76 Endrin 0.01 5ND-0.01 Hexachlorobenzene 0.01 4ND-0.02 Heptachlor epoxide 0.02-0.46 0.08 Mirex 0.01 2ND-0.17 Toxaphene 0.3 2ND-6.4 p,p'-DDD 2ND-0.40 0.03 p,p'-DDE 0.23-13.00 1.58 p,p'-DDT 3ND-0.12 0.02 Total PCBs 4.9 0.8-56.0 7ND-0.0027 PCB Congener: 77 0.0096-0.830 105 0.0871 114 0.0057 0.0008-0.089 0.0350-1.399 118/1 06 0.2647 2ND-0.012 126 0.0005 0.0349 0.0054-0.460 156 169 8ND-0.00 12 Safe TEQsd 0.000465 e 0.000051-0.004066 Kennedy TEQsd 0.000929 f 0.000088-0.009715 Mercury 0.908 0.33-4.64 Selenium 3.976 2.72-5.17 aOrganochlorines are presented in flg/g wet weight and mercury and selenium in flg/g dry weight. bDashed lines indicate that no mean was calculated, because more than half of the values were below the level of detection. c ND = not detected; number preceding "ND" is the number of samples not detected. d Toxic equivalents of PCB congeners relative to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) were estimated by summing the products of the measured concentrations and TCDD toxic equivalency factors suggested by Safe (1990) and Kennedy et ai. (1996). PCB congeners measured include numbers 77, 105, 114, 118, 126,156,169; (congener 118 coeluted with 106). e Arithmetic mean was 0.000958 flg/g. f Arithmetic mean was 0.002147 flg/g.

from other locations in North America. PCB concentrations found in eggs collected in this study were higher than those found in GBH eggs from the northwest U.S. and British Columbia (Fitzner et al.

Location

n

Mean

± SE

Range

Reference Indiana Dunes

II

12.9 16.6

± 2.3 ± 2.5

1.7-29.3 5.0-41.0

16

1988, Blus et at. 1980, Elliott et at. 1989), Upper Mississippi River (UMR, Custer et at. 1997), and the Gulf coast of Texas (King et at. 1978). However, egg PCB concentrations were lower than those found in eggs collected from the Tennessee Valley (Fleming et al. 1984), Quebec (Laporte 1982), Nueces Bay, Texas (Mitchell et al. 1981), and the Great Lakes and north Atlantic regions (Ohlendorf et al. 1979). These values are only somewhat comparable because organochlorine concentrations in bird eggs have declined over the past 20 years (Custer et at. 1997). Mean PCB concentrations in GBH eggs in this study (geometric mean = 4.9 flg/g) were lower than the levels reported to affect reproduction in other colonial waterbird species. Concentrations of 14 flg/g PCBs in double-crested cormorant (Phatacrocorax auritus) eggs were associated with 15% egg mortality (calculated from Tillitt et at. 1992). However, concentrations of 20 to 40 flg/g PCBs in Caspian tern (Sterna caspia) eggs did not seem to reduce productivity (Struger and Weseloh 1985). Also, GBHs nesting in Nueces Bay, Texas had higher mean concentrations of PCBs in eggs (geometric mean = 6.2 flg/g PCBs) than found in this study and fledged 1.6 young per nest (Mitchell et al. 1981), a value within the range reported for stable populations elsewhere (Pratt 1970, Henny and Bethers 1971, BIus et al. 1980). The calculated toxicity of PCBs in our GBH eggs (arithmetic mean = 958 pg/g TEQs based on Safe [1990] TEFs) was intermediate to other locations where PCBs are dominant relative to TCDD. Toxic equivalents in GBH eggs collected from 10 locations on the Upper Mississippi River were about 30% as high (292 pg/g TEQs based on Safe [1990] TEFs) and were not associated with cytochrome P450 induction or embryo deformities (Custer et at. 1997). In contrast, the toxicity of PCBs in blackcrowned night-heron (Nycticorax nycticorax) eggs

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Custer et al.

from Green Bay, Wisconsin, was three times greater than our samples (2,997 pg/g TEQs based on Safe [1990] TEFs [Rattner et al. 1994]) and was associated with cytochrome P450 induction (Rattner et al. 1994). Black-crowned night-herons embryos from an earlier collection in Green Bay demonstrated cytochrome P450 induction and increased percentage of abnormal embryos (Hoffman et al. 1993). Total PCBs and toxic equivalents in eggs from Indiana Dunes were apparently not sufficiently high to induce EROD activity. Even though total PCBs were qualitatively higher in this study (geometric mean = 4.9 flg/g) than reference samples from the Upper Mississippi River (geometric mean = 1.0 flg/g, MacDougall, Minnesota, colony, Custer et aI. 1997), EROD activity (mean = 16.6 pmollminute/ mg microsomal protein) was not significantly higher than reference samples (mean = 12.3 pmollminute/mg microsomal protein). The lower frequency of brain asymmetry in GBH embryos at INDU (11 % INDU vs 47% Upper Mississippi River) and the associated higher total PCBs (4.9 vs 3.0 flg/g) and TEQs (929 vs 540 pg/g) support earlier results (Custer et al. 1997) which suggested that PCBs were not correlated with brain asymmetry in GBHs. In that study (Custer et aI. 1997), EROD activity was higher in GBH embryos with asymmetrical brains than symmetrical brains. However, because total PCBs and TEQs were not correlated with EROD activity, no relationship between brain asymmetry and PCBs or TEQs was apparent. It should be noted that direct comparisons between PCBs and brain asymmetry could not be made, because PCBs were not measured in embryos selected for brain measurements. It is possible, as mentioned earlier (Custer et al. 1997), that a chemical or chemicals not measured on the Upper Mississippi River and not present at INDU was responsible for the higher frequency of asymmetry on the Upper Mississippi River. It is also possible that the stress of the 1993 flood event (Custer et al. 1996) on GBHs nesting on the Upper Mississippi River, regardless of contamination, may have influenced EROD activity and brain asymmetryon the Upper Mississippi River. The GBHs nesting in the more flooded areas of the Upper Mississippi River delayed nesting and had smaller clutch sizes compared to areas less affected by the flood (Custer et aI. 1996). On the other hand, the sample size of brains measured for symmetry at INDU (n = 9) may have been too small for an accurate evaluation of the percent that were asymmetrical. Further studies on asymmetry in GBH embryos

are obviously required to test these various hypotheses. Based on the high concentrations of PCBs in their eggs (15 and 56 flg/g) , two female GBHs in this study were probably feeding at PCB contaminated sites. It is uncertain whether these contaminants were obtained in Indiana near the nesting colony or prior to the nesting season at some other location. Local exposure could be determined by measuring accumulation of organochlorines (flg/d) in heron chicks (Custer and Custer 1995) or by analyzing prey items obtained from the chicks or adults. Fish tissue data collected by the IDEM from Lake Michigan, Trail Creek, and Little and Grand Calumet rivers have concentrations of PCBs sufficiently elevated to warrant fish consumption advisories (Indiana State Department of Health 1997). Portions of these waterbodies are within the foraging range of this heron rookery. Although p,p'-DDE concentrations (geometric mean = 1.6 flg/g) in GBH eggs from INDU were generally equal to or lower than at other locations and times (King et aI. 1978, Ohlendorf et al. 1979, Blus et al. 1980, Mitchell et aI. 1981, Laporte 1982, Elliott et aI. 1989, Custer et aI. 1997), one of the ten eggs had an elevated concentration of p,p'-DDE (13 flg/g) that could have caused reproductive impairment. P,p'-DDE concentration> 10 flg/g in eggs were correlated with decreased reproductive success in other heron species (Custer et aI. 1983, Henny et al. 1984, White et al. 1988). Eggshell thickness was comparable to that reported at other locations in recent times (Custer et al. 1997) and does not seem threatening to GBHs nesting success on the UMR. Eggshells from INDU averaged 3.4% thinner than eggshells collected prior to 1947, but not nearly as much as the 15 to 20% thinning associated with population changes (Anderson and Hickey 1972). Mercury concentrations in our study (mean = 0.15 flg/g wet weight, maximum = 0.76 flg/g; converted from dry weight using 83.7% moisture) were similar to or lower than those reported in GBHs at other locations (Custer et aI. 1997) and probably do not constitute a threat to GBHs. The concentration of mercury associated with reproductive failure varies by species (Ohlendorf et al. 1978) and a critical mercury level in eggs has not been determined for any heron species. However, mean mercury concentrations in our study were generally below the concentrations reported to affect reproduction of ring-necked pheasants (Phasianus coIchicus; 0.5 flg/g wet weight; Fimreite 1971), mallards

Contaminants in Great Blue Heron Eggs (Anas ptatyrhynchos; 0.85 f.lg/g; Heinz 1979), common terns (Sterna hirundo; 1.0 f.lg/g; Connors et at. 1975), or herring gulls (Larus argentatus; > 16 f.lg/g; Vermeer et at. 1973). Selenium concentrations in eggs in this study (mean 4.0 f.lg/g dry weight) were above the upper boundary of "normal" selenium concentrations in waterbird eggs (3.0 f.lg/g dry weight) (Skorupa and Ohlendorf 1991). However, selenium concentrations (mean = 0.65 f.lg/g wet weight, maximum = 0.84 f.lg/g wet weight; converted from dry weight using 83.7% moisture) were below the 3.0 f.lg/g wet weight threshold for causing reproductive impairment (Heinz 1996). The timing of nesting was very similar to that reported at the same colony in 1981 (Taylor et at. 1982). The first evidence of incubation (3 April), date of first hatching (3 May), and first fledging (2 July) recorded in 1981 were all within the periods observed in 1993. Nest initiation in this study (median first egg laid = 10 April) was also comparable to that on the UMR in 1995 (median = 5 to 9 April) and in unflooded portions of the UMR in 1993 (median = 13 April, Custer et at. 1997). Nest initiation seemed delayed in flooded reaches of the UMR in 1993 (median = 29 April). Mean clutch size at this colony in 1993 (4.2 eggs per clutch) was somewhat lower than reported in 1981 (5.0 eggs per clutch, Taylor et al. 1982) but comparable to clutch sizes of GBHs from elsewhere in the North America (varied from 3.2 to 5.0 eggs per clutch; Custer et at. 1997). In contrast, abnormally small clutches (mean = 2.2-2.9 eggs per clutch) were observed in flooded portions of the UMR in 1993 (Custer et at. 1997). Our results suggest that the GBH colony at INDU is not threatened by organochlorines, mercury, or selenium. With minor exception, concentrations of these environmental contaminants in GBH eggs were at background levels. Relatively low levels of EROD activity, the low frequency of brain asymmetry, and the information on nest initiation, and clutch size support this contention. On the other hand, few data on contaminant exposure of other avian species are available in this area and the potential exists for elevated levels of PCBs or other contaminants to cause localized problems. The exposure of GBHs at the Indiana Dunes National Lakeshore colony to petroleum contamination, a major contaminant in areas surrounding the heron colony (Hoke et at. 1993), requires further evaluation. We did not attempt to measure poly-

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cyclic aromatic hydrocarbons (PAHs) in the egg contents, because PAHs quickly metabolize in eggs during incubation (Naf et at. 1992) and therefore are unlikely to be detected. Because induction can be correlated with PAH exposure (Lee et at. 1986), the lack of significant induction in livers of our embryos suggests that PAH exposure may not be of concern. Concentrations of PAHs in fresh eggs or chicks from this colony and a reference location would be useful in evaluating PAH exposure at this location.

ACKNOWLEDGMENTS We thank Indiana Dunes National Lakeshore for access to their properties, Jason Butcher for preparing the figures, Randy Knutson, Ralph Grundel, and Al Parker for technical and field assistance, and Christine Custer, J. Christian Franson, John P. Giesy, Fred Meyer, and one anonymous reviewer for comments on the manuscript. This project was funded by the U.S. Fish and Wildlife Service, Division of Environmental Contaminants. REFERENCES Anderson, D. W., and Hickey, J. 1. 1972. Eggshell changes in certain North American birds. Proe. Inter. Ornitho!. Congr. 15:514-540.

Bellward, G. D., Norstrom, R., Whitehead, P. E., Elliot, J. E., Bandiera, S. M., Dworschak, C., Chang, T., Forbes, S., Cadario, B., Hart, L. E., and Cheng, K. M. 1990. Comparison of polychlorinated dibenzodioxin levels with hepatic mixed-function oxidase induction in great blue herons. 1. Toxieol. Environ. Health 30:33-52.

BIus, L. J., Henny, C. J., and Kaiser, T. E. 1980. Pollution ecology of breeding great blue herons in the Columbia Basin, Oregon and Washington. Murrelet 61:63-71.

Brock, K. J. 1986. Birds of the Indiana Dunes. Bloomington Indiana: Indiana Univ. Press. Connors, P. G., Anderlini, V. C., Risebrough, R. W., Gilbertson, M., and Hays, H. 1975. Investigations of heavy metals in common tern populations. Can. Field-Nat. 89: 157-162.

Custer, T. W., and Custer, C. M. 1995. Transfer and accumulation of organochlorines from black-crowned night-heron eggs to chicks. Environ. Toxieo!. Chern. 14:533-536.

_ _, Hensler, G. L., and Kaiser, T. E. 1983. Clutch size, reproductive success, and organochlorine contaminants in Atlantic coast black-crowned night-herons. Auk 100:699-710.

_ _, Hines, R. K., and Custer, C. M. 1996. Nest initiation and clutch size of great blue herons on the Mis-

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Custer et al.

sissippi River in relation to the 1993 flood. Condor 98:181-188. _ _, Hines, R. H., Melancon, M. J., Hoffman, D. J.. Bickham, J. W., Martin, J. W. and Henshel, D. 1997. Contaminant concentrations and biomarker response in great blue heron eggs from 10 colonies on the upper Mississippi River, USA. Environ. Toxicol. Chem. 16: 260-271. Elliott, J. E., Butler, R. W., Norstrom, R., and Whitehead, P. E. 1989. Environmental contaminants and reproductive success of great blue herons Ardea herodias in British Columbia, 1986-87. Environ. Pollut. 59:91-114. Fimreite, N. 1971. Effects of dietary methylmercury on ring-necked pheasants. Can. Wildl. Ser., Occ. Papers 9. Fitzner, R. F., Blus, L. J., Henny, C. J., and Carlile, D. W. 1988. Organochlorine residues in great blue herons from the northwestern United States. Colon. Waterbirds 11 :293-300. Fleming, W. J., Pullin, B. P., and Swineford, D. M. 1984. Population trends and environmental contaminants in herons in the Tennessee Valley, 1980-81. Colon. Waterbirds 7:63-73. Hart, L. E., Cheng, K. M., Whitehead, P. E., Shah, R. M., Lewis, R. J., Ruschkowski, S. R., Blair, R. W., Bennett, D. C., Bandiera, S. M., Norstrom, R., and Bellward, G. D. 1991. Dioxin contamination and growth and development in great blue heron embryos. J. Toxicol. Environ. Health 32:331-344. Heinz, G. H. 1979. Methylmercury: Reproductive and behavioral effects on three generations of mallard ducks. J. Wildl. Manage. 43:394-401. ___. 1996. Selenium in birds. In Environmental Contaminants in Wildlife, Interpreting Tissue Concentrations, eds. W.N. Beyer, G.H. Heinz, and A. W. Redmon-Norwood, pp. 447-458. Boca Raton, FL: Lewis Publishers. Henny, C. J., and Bethers, M. R. 1971. Population ecology of the great blue heron with special reference to western Oregon. Can. Field-Nat. 85:205-209. _ _, Blus, L. J., Krynitsky, A. J., and Bunck, C. M. 1984. Current impact of DDE on black-crowned night-herons in the Intermountain West. 1. Wildl. Manage. 48:1-13. Henshel, D. S., Cheng, K. M., Norstrom, R., Whitehead, P., and Steeves, J. D. 1993. Morphometric and histological changes in brains of great blue heron hatchlings exposed to PCDDs: Preliminary analysis. In ASTM STP#1179: First Symposium on Environmental Toxicology and Risk Assessment: Aquatic, Plant, Terrestrial, eds. W. Landis, J. S. Hughes and M. A. Lewis, pp. 262-277. American Society for Testing and Materials. ___ , Martin, J. W., Norstrom, R., Whitehead, P., Steeves, J. D., and Cheng, K. M. 1995. Morphometric abnormalities in brains of great blue heron hatchlings

exposed in the wild to PCDDs. Environ. Health Perspeet. 103:61-66. Hines, R. K., and Custer, T. W. 1995. Evaluation of an extendable pole-net to collect heron eggs in the canopy of tall trees. Colon. Waterbirds 18:120-122. Hoffman, D.J., Smith, G. J., and Rattner, B. A. 1993. Biomarkers of contaminant exposure in common terns and black-crowned night herons in the Great Lakes. Environ. Toxicol. Chem. 12: 1095-1103. Hoke, R. A., Giesy, J. P., Zabik, M., and Unger M. 1993. Toxicity of sediments and sediment pore waters from the Grand Calumet River-Indiana Harbor, Indiana Area of Concern. Ecotoxicol. Environ. Safety 26:86-112. Indiana Department of Environmental Management. 1988. Northwest Indiana Environmental Action Plan. Draft Area of Concern Remedial Action Plan. Indiana Department of Environmental Management, Indianapolis, IN. Indiana State Department of Health. 1997. 1997 Indiana Fish Consumption Advisory. Indianapolis, Indiana. International Joint Commission. 1985. Report of Great Lakes Water Quality. International Joint Commission, Great Lakes Water Quality Board, Kingston, Ontario, Canada. Kennedy, S. W., Lorenzen, A., and Norstrom, R. 1996. Chicken embryo hepatocyte bioassay for measuring cytochrome P4501A-based 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalent concentrations in environmental samples. Environ. Sci. Technol. 30:706-715. King, K. A., Flickinger, E. L., and Hildebrand, H. H. 1978. Shell thinning and pesticide residues in Texas aquatic bird eggs, 1970. Pest. Monit. Journal 12: 16-21. Laporte, P. 1982. Organochlorine residues and eggshell measurements of great blue heron eggs from Quebec. Colon. Waterbirds 5:95-103. Lee, Y. Z., O'Brien, P. J., Payne, J. F., and Rahimtula, A. D. 1986. Toxicity of petroleum crude oils and their effects on xenobiotic metabolizing enzyme activities in the chicken embryo in Ovo. Environ. Res. 39:153-163. Melancon, M. J. 1996. Development of Cytochromes P450 in Avian Species as a Biomarker for Environmental Contaminant Exposure and Effect: Procedures and Baseline Values. In ASTM STP#1306: Environmental Toxicology and Risk Assessment: Biomarkers and Risk Assessment, eds. D. A. Bengston, and D. S. Henshel, pp. 95-108. American Society for Testing and Materials. Mitchell, C. A., White, D. H., and Kaiser, T. E. 1981. Reproductive success of great blue herons at Nueces Bay, Corpus Christi, Texas. Bull. Texas Ornithol. Soc. 14:18-21. Naf, C., Broman, D., and Brunstrom, B. 1992. Distribution and metabolism of aromatic hydrocarbons

Contaminants in Great Blue Heron Eggs (PAHs) injected into eggs of chicken (Gallus domesticus) and common eider duck (Somateria mollissima). Environ. Toxico!. Chem. II: 1653-1660. Ohlendorf, H. M., Risebrough, R. W., and Vermeer, K. 1978. Exposure of marine birds to environmental pollutants. U. S. Fish Wild\. Ser., Wild\. Res. Rep. No.9. _ _, Klaas, E. E., and Kaiser, T. E. 1979. Environmental pollutants and eggshell thickness: Anhingas and wading birds in the eastern United States. U. S. Fish and Wild\. Ser., Spec. Sci. Rep.-Wild\. No. 216. PAHLS, Inc. 1993. The Environment of' Northwest Indiana. 102 N. Morgan Blvd., Valparaiso, IN 46383. Pratt, H. M. 1970. Breeding biology of great blue herons and common egrets in central California. Condor 72:407-416. Rattner, B. A., Hatfield, J. S., Melancon, M. J., Custer, T. W., and Tillit, D. E. 1994. Relation among cytochrome P450, AH-active PCB congeners and dioxin equivalents in pipping black-crowned nightheron embryos. Environ. Toxico!. Chem. 13: 1805-1812. Safe, S. 1990. Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and related compounds: Environmental and mechanistic considerations which support the development of toxic equivalency factors (TEFs). Crit. Rev. Toxico!. 21:51-88. Skorupa, J. P., and Ohlendorf, H. M. 1991. Contaminants in drainage water and avian risk thresholds. In The Economics and Management of Water and

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Drainage in Agriculture, eds. A. Dinar and D. Zilberman, pp. 345-368. Boston, MA: Kluwer Academic. Steffeck, D. W. 1989. A survey for contaminants in biota near the Midco I, Midco fl, and Ninth Avenue dump hazardous waste sites in Gary, Lake County, Indiana. U.S. Fish and Wildlife Service, Bloomington, IN. Struger, 1., and Weseloh, D. V. 1985. Great Lakes Caspian terns: egg contaminants and biological implications. Colon. Waterbirds 8:142-149. Taylor, T. M., Reshkin, M., and Brock, K. 1. 1982. Recreational land use adjacent to an active heron rookery: A management study. Ind. Acad. Sci. 91: 226-236. Tillitt, D. E., Ankley, G. T., Giesy, J. P., Ludwig, J. P., Kurita-Matsuba, H., Weseloh, D. V., Ross, P. S., Bishop, C. A., Sileo, L., Stromborg, K. L., Larson, J., and Kubiak, T. J. 1992. Polychlorinated biphenyl residues and egg mortality in double-crested cormorants from the Great Lakes. 1992. Environ. Toxicol. Chem. 11:1281-1288. Vermeer, K., Armstrong, F. J., and Hatch, D. M. 1973. Mercury in aquatic birds at Clay Lake, Western Ontario. J. Wild!. Manage. 37:58-61. White, D. H., Fleming, W. J., and Ensor, K. L. 1988. Pesticide contamination and hatching success of waterbirds in Mississippi. 1. Wildl. Manage. 52: 724-729. Submitted: 21 June 1997 Accepted: 26 October 1997 Editorial handling: Peter F. Landrum