Study Area and Test Species

The purpose of this study was to expand upon that earlier investigation by Beyer and others (2013) by increasing the number of birds examined; and by examining the relations between lead concentrations in soil and in tissues (blood, livers, and kidneys), the associated adverse health effects (for example, δALAD inhibition, oxidative stress, microscopic lesions), and the reproductive outcomes in birds. Beyer and others (2013) determined that ground-foraging birds (northern cardinals and American robins) in the Southeast Missouri Lead Mining District (n=34) had lead concentrations in tissues greater than birds from reference sites (n=39) by factors of 8 in blood, 13 in livers, and 23 in kidneys. In addition, the mean lead concentrations in tissues from the district (Beyer and others, 2013) approached or exceeded adverse effects concentrations reported in the literature (Franson and Pain, 2011), and were greater than concentrations reported in birds from the Coeur d’Alene and the Tri-State Mining Districts (Beyer and others, 2004; Hansen and others, 2011). Birds from the Southeast Missouri Lead Mining District areas experienced a 58- to 82-percent decrease in δALAD activity compared to birds from reference sites, and results for lead in earthworm tissues indicated that birds were exposed to lead via ingestion of soil-dwelling prey species in the district (Beyer and others, 2013).

We examined 8 species of birds (7 songbird species): American robin, eastern bluebird, eastern towhee, Spizella pusilla (A. Wilson, 1810) (field sparrow), Passerina cyanea (Linnaeus, 1766) (indigo bunting), northern cardinal, Molothrus ater (Boddaert, 1783) (brown-headed cowbird) and 1 gamebird species: mourning dove (table 1) at 3 lead-contaminated sites and three references sites in or next to the Southeast Missouri Lead Mining District. These species were selected based on (1) abundance and availability during the study breeding seasons and (2) because they spend considerable time foraging on or near the ground for invertebrates or seeds (Halkin and Linville, 1999; Payne, 2006; Carey and others, 2008; Otis and others, 2008; Greenlaw, 2015) and, as such, have an increased likelihood of ingesting soil-borne lead particles or being otherwise exposed to lead in the environment. We evaluated reproductive metrics for five (eastern bluebird, eastern towhee, field sparrow, indigo bunting, and northern cardinal) of the seven songbird species; the number of nests monitored per year by site and species are shown in table 2.

Three lead-contaminated sites (simply called “contaminated sites”) and three reference sites (in other words, sites having background concentrations of lead and no known source of lead-mining inputs) in the Southeast Missouri Lead Mining District were selected for this study (table 3); only two of the three reference sites were evaluated for reproductive measures. The contaminated sites were (1) within the footprint of the Anschutz Mine, (2) in the Big River floodplain, and (3) near Magmont Vent. In this context, “contaminated” sites refer to areas where mining, milling, and smelting processes released lead into the environment, and where lead concentrations in soil exceeded a background concentration of 40 parts per million (ppm; a concentration of 1 ppm, quantified by portable X-ray fluorescence spectrometry [pXRF], is about equivalent to 1 mg/kg analyzed by inductively coupled plasma-mass spectrometry [ICP–MS]). Previous work by Beyer and others (2013) indicated that the Missouri-specific background for lead concentrations in soil is 20 to 62 mg/kg; the selection of 40 ppm as the background concentration represents the approximate midpoint of this background range. Areas that have no history of lead releases from mining, milling, or smelting processes were selected as reference sites.

The Anschutz Mine site (table 3) is in Madison County, Missouri, southeast of Fredericktown, Mo., in the Old Lead Belt (fig. 1). The site hosted lead mining and smelting operations between the 1840s and the 1960s (USFWS and Missouri Department of Natural Resources, 2014). The Anschutz Mine site is currently under remediation (2008 to present [2022]) and contained areas of exposed chat piles and areas that were partially capped with uncontaminated soil during this study.

The Magmont Vent site is in the Viburnum Trend (table 3; fig. 1) in Mark Twain National Forest (Iron County, Mo.) and was named for a mine air vent shaft near the inactive Magmont Mine (Bixby, Mo.). The site is near active lead mining and smelting operations. One example of a nearby smelting operation is the Buick Resource Recycling Facility (Boss, Mo.). This facility, which is about 1.6 kilometers (km) southwest of the Magmont Mine, is a former primary lead smelter that has been converted to a secondary smelter; the facility has been operating in that capacity over the last two decades (since the early 2000s). Beyer and others (2013) determined the average lead concentration in soil in this area to be about 1,000 mg/kg dw.

The Big River is a tributary of the Meramec River and flows through multiple counties in the Southeast Missouri Lead Mining District. The river’s drainage area includes much of the Old Lead Belt. Decades of erosion, leaching, and dam failures have resulted in releases of lead-contaminated mine tailings into the river, and historical floods have deposited lead-contaminated sediments along the riverbanks. The Big River floodplain site is in Washington and Jefferson Counties, Mo. (table 3; fig. 1); study areas include a public day-use area in Washington State Park (Desoto, Mo.), and a tract of private land, which is on the Big River about 3.5 km downstream (2.5 km north) from the day-use area. Note that our Big River floodplain study site does not encompass the entirety of the river’s contaminated floodplain areas (see Pavlowsky and others, 2017 for more information about the extent of metal contamination in the watershed). Soil and sediment samples previously collected from the floodplain areas at Washington State Park and nearby river segments had a mean lead concentration of 1,700 mg/kg dw (Beyer and others, 2013; Pavlowsky and others, 2017).

The primary reference sites (table 3; fig. 1) for the reproductive part of this study were Johnson’s Shut-ins State Park (Reynolds County, Mo.) and Hawn State Park (Ste. Genevieve County, Mo.). However, the abundances of target species for lethal sampling at the reference sites were low; therefore, in 2018, Cape Girardeau (Cape Girardeau County, Mo.) was added as a third reference site to increase sample sizes for lethally sampled birds only. As previously mentioned, all reference sites were in areas with no history of lead mining or contamination and were expected to have background concentrations of lead (<40 ppm).

Habitat, geology, soil types, and mixtures of vegetation types were similar among the study sites. All sites were in or near the St. Francois Mountains, which is the topographic and structural high point of the southeast Missouri Ozarks. The study sites are largely underlain by Cambrian or Ordovician dolostone, limestone, sandstone, or siltstone (Horton, 2017). Anschutz Mine, Big River floodplain, Cape Girardeau, and Johnson’s Shut-ins State Park also feature shale; chert is present at Anschutz Mine. Johnson’s Shut-ins State Park includes areas with exposed Precambrian igneous rocks, which locally affect soil type and vegetation; however, bird sampling was not conducted in those areas because we aimed to keep the underlying geology relatively consistent among study sites.

Habitat at Anschutz Mine is a mix of Pinus L. spp. (pine) and Quercus L. spp. (oak) woodlands, old fields dominated by grasses, forbs, and Juniperus virginiana L. (eastern redcedar) tree saplings, and capped and revegetated mine/mill wastes, which are covered by grasses and forbs. The Big River floodplain site consists of mowed fields, old fields, riparian forest, and upland Carya Nutt spp. (oak-hickory) forest on the slopes leading out of the river valley. Magmont Vent is dominated by oak-hickory woodlands and forested uplands with scattered pines; the site is also dissected by mowed power line cuts, a rail line, and a private road. The Hawn State Park site is an upland old field surrounded by oak-hickory and pine forest. Johnson’s Shut-ins State Park consists of old fields and mowed fields surrounded by oak-hickory forest with occasional pines. Photographs of nesting sites and habitat types at the study sites are provided in appendix 1.

Previous Research on Lead Uptake by Wild Birds

Although an extensive body of literature addresses effects of lead in common test and game species (for example, chickens and mallards), few studies address lead exposure, uptake, and toxicity in small, ground-feeding wild birds (for example, songbirds); moreover, no toxicity thresholds link contaminant concentrations (in the environment or in tissues) to effects on songbirds. Incidental ingestion of lead-contaminated soil by ground-foraging birds has been documented near shooting ranges and mining operations (Vyas and others, 2000; Hansen and others, 2011). In the Coeur d’Alene mining district in northern Idaho (not shown), which has a lengthy history of lead mining and smelting, Turdus migratorius (Linnaeus, 1766) (American robin) nestlings (n=10) from mining- and smelting-contaminated areas had lead concentrations in blood ranging from 1.24 to 4.00 milligrams per kilogram (mg/kg) on a dry-weight basis (dw, assuming 78.26 percent moisture; Blus and others, 1995); however, blood samples were not collected from reference birds for comparison. Lead concentrations in liver tissues of hatch year American robins from contaminated areas (n=10) were 0.22 to 17.4 mg/kg dw (assuming 67.74 percent moisture) compared to 0.28 to 0.96 mg/kg dw in juveniles from reference areas (n=2, assuming 67.74 percent moisture; Blus and others, 1995). Young Tachycineta bicolor (Vieillot, 1808) (tree swallows) from contaminated areas (n=11) had lead concentrations in blood ranging from not detected (ND) to 3.45 mg/kg dw and lead concentrations in livers ranging from ND to 1.83 mg/kg dw. Tree swallows from reference areas (n=3) had similar lead concentration ranges in tissues (0.92 to 1.89 mg/kg dw in blood, assuming 78.26 percent moisture; and ND to 9.9 mg/kg dw in livers, assuming 67.74 percent moisture; Blus and others, 1995).

Similarly, Johnson and others (1999) observed that American robins and Melospiza melodia (A. Wilson, 1810) (song sparrows) from contaminated areas in the Coeur d’Alene mining district exhibited δALAD inhibition of 51 to 75 percent compared to birds from reference areas. In other words, δALAD activity in song sparrows was reduced from 305 units of δALAD activity at the reference area to 149 activity units at the contaminated area (51 percent); δALAD activity in American robins was reduced from 302 units at the reference site to 76 units at the contaminated site (75 percent). Additionally, the mean lead concentration in livers of song sparrows from contaminated areas (5.98 mg/kg dw, assuming 67.74 percent moisture; n=5) was about twentyfold greater than that in song sparrows from reference areas (0.30 mg/kg dw, assuming 67.74 percent moisture; n=4).

A subsequent study of ground-feeding songbirds (American robins, song sparrows, and Catharus ustulatus [Nuttall, 1840] [Swainson’s thrushes]) in the Coeur d’Alene mining district (Hansen and others, 2011) found that birds from reference sites had lead concentrations in blood less than (<) 0.19 mg/kg dw; comparatively, birds from moderately and highly contaminated sites had elevated lead concentrations in blood (1.09 to 2.06 mg/kg dw, respectively). Mean lead concentrations in bird livers were 0.27 mg/kg dw at the reference sites and 6.56 to 15.8 mg/kg dw at the contaminated sites. Birds at the contaminated sites experienced δALAD inhibition of 31 to 71 percent compared to birds at the reference site. Regression analyses of the soil ingestion rate and the lead concentrations in ingesta and soil indicated that lead concentrations in soil at a study site approximates the lead concentrations ingested by birds (Hansen and others, 2011). As a result, Sample and others (2011) concluded that the lead concentrations in blood and livers of birds in the Coeur d’Alene mining district were directly correlated to the lead concentrations in soil.

In the Tri-State Mining District in southwestern Missouri, northeastern Oklahoma, and southeast Kansas (not shown), Beyer and others (2004) sampled blood, livers, and kidneys of five bird species (American robin; Cardinalis cardinalis [Linnaeus, 1758] [northern cardinal]; Riparia riparia [Linnaeus, 1758] [bank swallow]; Stelgidopteryx ruficollis [Vieillot, 1817] [southern rough-winged swallow]; Toxostoma rufum [Linnaeus, 1758] [brown thrasher]) for chemical, biochemical, and histopathological analyses. Birds from the Tri-State Mining District had greater mean lead concentrations in livers (4.2 to 9.3 mg/kg dw), kidneys (6.2 to 20 mg/kg dw), and blood (1.1 to 1.7 mg/kg dw) compared to birds from the reference sites, which were in Maryland (liver, <0.02 to 5.1 mg/kg dw; kidney, 0.081 to 8.9 mg/kg dw; blood, <0.04 to 0.86 mg/kg dw; Beyer and others, 2004). Mean lead concentrations in tissues of American robins from the Tri-State Mining District were as much as 2.2-fold greater than in robins from the reference sites; lead concentrations in northern cardinals at the Tri-State Mining District were as much as 23-fold greater than in cardinals from reference sites. Tree swallows at the Tri-State Mining District had lead concentrations in tissues 27 to 210 times greater than concentrations in swallows from reference sites. Additionally, all northern cardinals (n=6) and American robins (n=6) sampled in the Tri-State Mining District had greater than (>) 50 percent δALAD inhibition compared to birds from reference sites (n=10: 6 cardinals and 4 robins); 4 individuals (2 robins, 1 cardinal, and 1 thrasher) from the Tri-State Mining District had >90 percent δALAD inhibition. Although microscopic lesions in tissues could not be definitively linked to lead toxicity, several individuals (n=6 of 34) from the Tri-State Mining District had lead concentrations in tissue that exceeded literature-based toxicity thresholds (that is, external signs of lead poisoning occur when lead concentration in blood is 2.6 to 5.2 mg/kg dw or hepatic lead is 20 to 50 mg/kg dw; Pain, 1996).

Effects of Lead Exposure on Avian Recruitment

Adult birds that have accumulated lead in their tissues (for example, in bone, blood, and kidneys) can excrete lead from their body via multiple physiological mechanisms. For females, egg laying is a potential route of excretion that can reduce the lead burden of the female but may subsequently expose the offspring to lead. Maternal transfer of lead to eggs has been documented in many avian species, including Anas platyrhynchos (Linnaeus, 1758) (mallards) (Vallverdú-Coll and others, 2016), Branta canadensis (Linnaeus, 1758) (Canada geese) (Tsipoura and others, 2011), Streptopelia roseogrisea (Sundevall, 1857) (ringed turtle doves) (Kendall and Scanlon, 1981), Larus argentatus (Pontoppidan, 1763) (herring gulls) (Burger and Gochfeld, 1993), and four species of terns (Maedgen and others, 1982; Burger and Gochfeld, 1991, 199327). For example, 1 study of Ficedula hypoleuca (Pallas, 1764) (pied flycatchers) nesting near lead smelter operations determined that the median lead concentration in egg contents was 10 times greater at contaminated sites than at reference sites (Nyholm, 1998). Lead concentrations in feathers of adult female Sterna hirundo (Linnaeus, 1758) (common terns) have correlated to concentrations in their eggs (Burger and Gochfeld, 1991), which further indicates maternal transfer of lead to offspring.

Maternal transfer of lead to eggs can reduce egg hatchability. For example, Buerger and others, 1986) observed an about 16-percent reduction in hatchability of eggs laid by lead-exposed (one lead pellet; 72 milligrams [mg] of lead) female Zenaida macroura (Linnaeus, 1758) (mourning doves) compared to unexposed control females; the authors hypothesized that maternal transfer of lead caused increased rates of early embryonic mortality. Additionally, embryo exposure to metals in the egg may inhibit early nestling brain development and motor skills (Grue and others, 1986; Nyholm, 1998; Müller and others, 2008); however, the exact effects of maternally mediated lead concentrations seem to be species dependent. For example, Nyholm (1998) found that maternal exposure to lead did not significantly affect breeding results (that is, number of fledged nestlings per egg) of pied flycatchers, nor was the health of the offspring affected by maternal-mediated lead concentrations.

After hatching, young birds are exposed to lead principally through direct ingestion of lead-contaminated food items and soil. Lead concentrations in the blood of nestlings and adult birds have reflected the lead concentrations in soil of their environment (Nyholm, 1998); lead concentrations in blood increase with increases of lead concentrations in soil (Roux and Marra, 2007). Altricial nestlings, which hatch blind and have little or no down, are unable to leave the nest for 2 to 3 weeks after hatching and rely on prey species provided by parents (Burger and Gochfeld, 1993; Nyholm, 1998; Janssens and others, 2003). Because of their exposure at early ages, altricial young also can have greater sensitivity to lead than adults of the same species (Hoffman and others, 1985b; Grue and others, 1986; Scheuhammer, 1987); nestlings are particularly useful as sentinels of adverse local conditions because their uptake and exposure closely reflects local contaminant bioavailability because of the limited foraging ranges of the parents during the breeding season (Burger and Gochfeld, 1993; Nyholm, 1998; Janssens and others, 2003).

The effects of lead exposure on nestlings are similar to effects on adults and include decreased hemoglobin, inhibited δALAD activity (Grue and others, 1984, 198672; Hoffman and others, 1985b), and decreased cellular immune response (Vallverdú-Coll and others, 2016). For example, Grue and others (1986) determined that adults and nestlings of Sturnus vulgaris (Linnaeus, 1758) (European starlings) exposed to lead in soil (56–410 mg/kg dw) and invertebrates (140–1,200 mg/kg dw) had decreased δALAD activity, but nestlings also experienced decreased brain weights, and decreased hematocrit and hemoglobin concentrations, compared to control birds (soil: 16–34 mg/kg dw lead; invertebrates: 3.2–58 mg/kg dw). Similarly, Falco sparverius (Linnaeus, 1758) (American kestrel) nestlings that were fed a diet containing 125 mg/kg dw lead (as metallic lead or lead acetate) had decreased δALAD activity, hemoglobin concentration, and growth rate, and had increased mortality and delayed development, whereas adult kestrels fed a comparable control diet experienced no change in hematocrit, body weight, or survival (Custer and others, 1984; Hoffman and others, 1985a, 1985b82).

Additionally, altricial young are generally less tolerant of lead than their precocial counterparts (Hoffman and others, 1985b; Grue and others, 1986; Scheuhammer 1987). For example, the altricial kestrel nestlings described in the previous paragraph had multiple exposure effects, including reduced growth, when fed diets containing 125 mg/kg dw lead (Hoffman and others, 1985a), whereas growth impairment in precocial species, such as young chickens and Coturnix japonica (Temminck and Schlegel, 1848) (Japanese quail), was not apparent until the lead concentration in their diets was increased to 500 or 1,000 mg/kg (Stone and others, 1977; Franson and Custer, 1982). Differential growth and feeding rates may affect nestling tolerance to lead. Growth rates of altricial chicks are three to four times greater than those of precocial chicks, so altricial young may be fed lead-contaminated prey at a faster rate than precocial young feed themselves (Ricklefs, 1984). Potential differences in sensitivity to lead between altricial and precocial bird species are particularly of note for this study because the birds in this study are altricial, whereas literature-based toxicity thresholds were derived from studies conducted using precocial species such as chickens and ducks. In this way, the birds in this study may be substantially more sensitive to lead contamination than indicated by the established avian toxicity benchmarks (Franson and Pain, 2011).

Lead exposure has also been correlated with decreased body weight, increased nestling mortality (Hoffman and others, 1985a, 1985b82; Burger and Gochfeld, 1996; Vallverdú-Coll and others, 2016), and decreased nestling weight at fledging, which is a vital part of juvenile survival and fitness (Magrath, 1991; Gebhardt-Henrich and Richner, 1998; Naef-Daenzer and others, 2001). These effects may be caused by the harmful effects of lead exposure on brain development and function during early avian development. The brains of young birds are particularly sensitive to the effects of lead exposure; for example, decreased brain weight in lead-exposed young birds has been documented in at least three species (American kestrels, European starlings, and pied flycatchers) in studies performed in the laboratory and in the wild (Hoffman and others, 1985a; Grue and others, 1986; Nyholm, 1998). Developing male Taeniopygia guttata (Vieillot, 1817) (zebra finches) exposed to lead in drinking water (1,000 micrograms per liter) had impaired song learning, reduced song nuclei, and altered sexual traits, causing reduced attention from females (Goodchild and others, 2021). In addition to physiological effects, lead (100 mg/kg as lead acetate) can adversely affect neurobehavioral and cognitive development and learning in nestlings (Burger, 1998; Burger and Gochfeld, 2000, 200531). Burger and Gochfeld (1996) studied these effects in the field by injecting lead acetate (100 mg/kg in sterile water) into 1 Larus argentatus (Pontoppidan, 1763) (herring gull) nestling in each of 22 broods. They determined that the lead-injected nestlings were smaller, less vigorous, and less able to compete for food with their unexposed siblings; subsequently, they weighed less by the age of 16 days, and increased parental care was required to overcome growth and behavioral deficits in the lead-exposed chicks. Locomotion, thermoregulation, begging, feeding, and response behaviors were also negatively affected in the lead-injected nestlings compared to the control nestlings. Similarly, lead-dosed (0.15 gram [g] lead) wild Sialia mexicana (Swainson, 1832) (western bluebird) nestlings had a decrease in the behavioral response of righting themselves after being placed on their backs relative to within-brood controls (Fair and Myers, 2002).

Lead exposure in nestlings can also delay the date of fledging and reduce overall rates of fledging success. For example, lead-injected herring gull nestlings had significantly lower fledging success than control nestlings (Burger and Gochfeld, 1996). In a study of Parus major (Linnaeus, 1758) great tits nesting along a contamination gradient (emanating from a smelter), fledging occurred substantially later at sites polluted with lead and other heavy metals (Janssens and others, 2003) compared to nestlings at study sites farther away from the smelter. Many studies have documented that fledging date affects juvenile survival and recruitment; individuals that leave the nest earlier generally have higher survival rates (Visser and Verboven, 1999; Monrós and others, 2002). Ultimately, adverse effects on recruitment may lead to population declines and losses of avian ecosystems services (including pest control, seed dispersal, and pollination; Gaston, 2022).

Study Scope

This study aimed to develop multiple lines of evidence of sublethal lead exposure, uptake, effects, and injury related to soil-borne lead for birds nesting at metals-contaminated sites compared to their counterparts at reference sites in the Southeast Missouri Lead Mining District. The specific objectives were the following: (1) determine lead exposure and uptake in birds breeding within the district by quantifying lead concentrations in blood, liver, and kidney tissues, and by evaluating lead concentrations in local soils and terrestrial invertebrates; (2) compare lead concentrations in bird tissues from contaminated sites to bird tissues from reference sites and to literature-based thresholds for avian toxicity; (3) assess sublethal physiological harm using δALAD activity, biochemical indicators of oxidative stress and deoxyribonucleic acid (DNA) damage, and microscopic examinations of kidney and liver tissues for evidence of lesions associated with lead exposure; and (4) determine if reproductive endpoints (clutch size, hatching success, number of young fledged, and nest survival) of five focal bird species were correlated to lead concentrations in blood and in local soils. All data (metadata and digital datasets) from this study are available in a U.S. Geological Survey (USGS) data release (Cleveland and others, 2023), per USGS data management policy, at https://doi.org/10.5066/P9RV1D60.

Bird Tissue Sampling

We collected blood (lethal or nonlethal), liver, and kidney samples from adult birds; blood samples (nonlethal) were also collected from nestlings and composited at the brood level for analyses (table 4). We quantified lead concentrations (blood, livers, and kidneys), δALAD activity (blood), and biochemical indicators of oxidative stress and DNA damage (livers) to assess lead exposure, uptake, and injury of birds. The livers and kidneys were examined for microscopic lesions typically associated with lead toxicoses. Lead concentrations in blood are one of the most-used measurement endpoints for assessing recent exposure to lead (days to weeks; Burger and Gochfeld, 1997; Beyer and others, 2004; Johnson and others, 2007; Buekers and others, 2009; Hansen and others, 2011) because lead concentrations in blood generally represent recent exposure via ingestion and gastrointestinal absorption (Vyas and others, 2001; Franson and Pain, 2011), whereas elevated lead concentrations in other tissues, including kidneys, livers, feathers, and bone, indicate long-term or chronic exposure (months to years). We targeted females because they produce the eggs and generally provide more parental care than the males; however, we sampled males when we were unable to capture the female of the breeding pair. Adult eastern bluebirds were captured using nest box traps (Friedman and others, 2008) or mist nets, whereas the adults of all other species (eastern towhee, field sparrow, indigo bunting, northern cardinal) were captured in their breeding territories using song playback and mist nets or funnel traps.

During lethal sampling (American robin, northern cardinal, mourning dove, brown-headed cowbird), we first collected blood samples for analyses of lead concentration, microhematocrit, and δALAD activity, and then euthanized the animals for collection of liver and kidney tissues (table 4). Individuals were euthanized in the field by inhalant carbon dioxide from a gas cylinder (Fair and others, 2010). Blood for lead analysis was collected from the cutaneous ulnar vein (punctured using a 26-gauge, 0.5-inch-long needle) using nonheparinized capillary tubes; the blood was then immediately transferred into pre-labeled, pre-cleaned glass screw-top test tubes. The test tubes were left capped until the sample was drawn and were re-capped immediately after filling to reduce or eliminate background contamination (app. 2). Samples were stored on ice in a cooler while in the field, and then frozen (−20 °C) until further processing. For δALAD analysis, blood was collected into heparinized capillary tubes and then immediately transferred into cryovials, frozen in liquid nitrogen in the field, and stored at −80 °C until further processing and analysis. Blood samples for microhematocrit were collected in heparinized capillary tubes; tubes were immediately placed on ice in a cooler while in the field and processed by centrifugation within 5 hours of collection.

The liver and kidneys were immediately dissected using clean, stainless steel dissecting tools. Tissues were divided into multiple sections, one each for histopathology (kidney and liver), lead analysis (liver and kidney), and oxidative stress and DNA damage (liver). The sections for each analysis were taken from consistent locations within each organ to standardize sampling protocols among individuals to the extent possible. A liver sample for histopathology consisted of an about 5- by 5-millimeter piece of the liver that was placed into a pre-labeled, plastic sample container containing 1:10 neutral buffered formalin (NBF). One lobe of the kidney was also added to this same sample cup. Histopathology samples were kept submerged in NBF at room temperature until shipping. Immediately before shipment to the pathology laboratory, the NBF was drained, and each tissue was wrapped in a NBF-soaked paper towel and sealed back into the labeled container. Tissues for lead analyses consisted of one lobe of kidney and an about 0.25-g ww section of liver, which were placed into two separate, pre-labeled and acid-washed plastic vials and stored on ice in the field. Samples were kept frozen (−20 °C) until further processing. Finally, an about 125-milligram (mg) ww section of liver was placed into a cryovial and immediately frozen in liquid nitrogen for analyses of oxidative stress and DNA damage biomarkers. Samples were kept frozen at −80 °C until further processing. All dissecting tools and surfaces were thoroughly cleaned with isopropyl alcohol between dissections.

Adults captured for nonlethal blood sampling were fitted with a USGS aluminum band, and we recorded standard morphometric measurements (weight, wing chord, tarsus length, bill length, and fat score). As before, we then collected blood from the cutaneous ulnar vein using nonheparinized capillary tubes; blood from the capillary tubes was immediately transferred into pre-labeled, pre-cleaned glass screw-top test tubes. Samples were stored on ice in a cooler while in the field and subsequently frozen (−20 degrees Celsius [°C]) until further processing. The volume of blood collected from individuals was no more than 1 percent of the body weight of the individual, in adherence to State and Federal guidelines and the recommendations of the Ornithological Council (Fair and others, 2010). Adult birds were released near the original site of capture after completing blood sampling.

Nestlings (table 4) were hand captured late in the nestling period, about 1 to 3 days before the expected day of fledging; the date of fledging was based on frequent observation of the nests (three to four times per week; Ralph and others, 1993). Each nestling was fitted with a USGS aluminum band and measured for tarsus length and body weight. The protocol for nestling (nonlethal) blood collection and sampling was the same as for adults; thus, a smaller volume of blood was collected from each individual nestling based on body weight. Therefore, blood samples from all nestlings within a brood were composited into a single, composite sample that represented the whole brood. We only included blood from conspecific nestlings, and not from cowbirds, in brood samples from nests that had been parasitized by cowbirds.

All bird capture and tissue collection occurred in adherence with Federal bird banding regulations (50 CFR parts 10, 13, and 21; R. Brasso 23903, sub-permittees M. Roach and K. Hixson); annual Missouri State collection permit regulations for 2016, 2017, and 2018 (R. Brasso: #16685, 17314, 17753 and M. Roach: #16778, 17268, 17657, respectively); and USFWS Scientific Collecting Permit regulations (D. Mosby, MB210154–0). Protocols were reviewed and approved by the Institutional Animal Care and Use Committees of Southeast Missouri State University and the U.S. Geological Survey’s (USGS) Eastern Ecological Science Center (formerly the Patuxent Wildlife Research Center).

Soil and Terrestrial Invertebrate Sampling

We also collected and analyzed soil and prey species (earthworms and beetle larvae) from the areas directly surrounding nest locations for lead concentrations to document a potential exposure pathway for ground-feeding birds. We hypothesized that exposure to lead-contaminated soils would cause adverse effects at the molecular, tissue, organ, and whole animal level in a dose-dependent manner, and thus potentially affect reproduction. Soil samples were either analyzed onsite or collected for offsite analyses of lead concentration at each site either immediately before (January to early March of 2016, 2017, 2018, or 2019) or after (August to November of 2016, 2017, or 2018) the breeding season. These efforts were offset from the breeding season to avoid disrupting nest building, egg laying, hatching, and other reproductive behaviors. The soil samples were either analyzed for lead onsite using pXRF or collected in plastic zipper bags and returned to the laboratory for offsite lead analyses by pXRF. All pXRF analyses followed the manufacturers’ instructions (similar to U.S. Environmental Protection Agency [EPA] method 6200; EPA, 2007). Quality-control (QC) measures (app. 2) for the pXRF units involved analyses of known standards and silicon dioxide blanks and the use of duplicate samples.

At the reference sites (table 3), we performed an initial round of quantification of lead in soil; those preliminary results indicated that neither reference site warranted intensive soil sampling because the lead concentrations in the reference soils were relatively low and homogeneous across the sites. Therefore, at the reference sites, composite soil samples were collected on a grid independent of exact location of bird nests. All reference soils were analyzed offsite. In contrast, lead concentrations in soil at the contaminated sites (table 3) were expected to be significantly more heterogeneous than the reference sites. Therefore, the lead concentrations of soils at the contaminated sites were characterized on a spatial basis. In this way, most soil sampling locations at the contaminated sites were within 100 m of each nest or nest box at 20-m intervals along two perpendicular transects intersecting nests. If nests were within 50 m of each other, one long transect was designed to intersect multiple nests. Discrete duplicate soil samples were collected from a subset of the onsite sampling points, which were analyzed offsite for lead concentrations to confirm the onsite results; these QC samples were collected using clean plastic or stainless-steel scoops at every third onsite sampling point along the transect (app. 2). Samples were similarly collected for offsite analyses when onsite conditions were too wet; additional discussion of QC measures for analyses of lead in soils, including the use of duplicate sampling and the collection of soils for offsite analyses (because of saturation with water), is in appendix 2. Site-specific information, including global positioning system coordinates, and date and time of sampling, was recorded at each sampling point. Between pXRF readings, the plastic shield that separates the pXRF window from the soil was cleaned with a low-lint cellulose fiber wiper, damp cloth, or disposable paper towel.

Clitellata spp. (earthworms) and beetle larvae (that is, Coleoptera spp. [Linnaeus, 1758] [grubs]) representing potential invertebrate prey species of the target bird species were collected for lead analysis in 2016. A total of 23 composite samples (consisting of about 90 percent earthworms and 10 percent beetle larvae by wet weight [ww]) were collected; sample sizes were n=5 at Hawn State Park, n=3 at Johnson’s Shut-ins State Park, n=3 at Anschutz Mine, n=9 at Big River floodplain, and n=3 at Magmont Vent. The weights of the invertebrate composite samples ranged from about 1 to 20 g ww, depending on invertebrate abundance at each site. Samples were collected from random locations (but locations at which lead concentrations in soil were analyzed using onsite pXRF) throughout the study sites to represent a range of lead concentrations in prey species. Invertebrates were collected from the top 15 cm of soil (in other words, the depth to which they would be most accessible to foraging birds) by digging with clean trowels; any invertebrates found were hand-transferred to plastic zipper bags, composited by site, and frozen until processing and analysis. The invertebrate samples were not depurated or rinsed before analyses to approximate incidental soil ingestion and dietary exposure to birds. We noted that, in general, the Big River floodplain site and areas near human use (that is, State park campgrounds) had considerably more worms than remote, upland sites such as Magmont Vent, Anschutz Mine, and areas within the reference sites that are remote and have less human use.

Analyses of Lead in Ground-Dwelling Invertebrates

Invertebrate samples were freeze-dried, homogenized, microwave-digested in high purity nitric acid and hydrogen peroxide (similar to EPA method 3050B; EPA, 1996), and analyzed for lead on a dw basis using ICP–MS (PerkinElmer Elan DRC–e or PerkinElmer NexION 2000, Waltham, Massachusetts; similar to EPA method 6020B; EPA, 2014) at the USGS Columbia Environmental Research Center (Columbia, Mo.). The QC measures for ICP–MS analyses included the use of a minimum of four calibration standards (traceable to the National Institute of Standards and Technology, Gaithersburg, Maryland) and a calibration blank, second-source continuing calibration standards and blanks, laboratory control solutions, certified reference materials, method and analytical duplicates and spikes, interference check solutions, and analytical dilutions. The QC results are summarized in appendix 2.

Analyses of Lead in Bird Tissues

Blood, liver, and kidney samples (table 4) were lyophilized before analysis; percent moisture in each sample was determined gravimetrically by weight loss on freeze-drying at the USGS Columbia Environmental Research Center. Blood samples were acid-digested in high-purity nitric acid and hydrogen peroxide in the collection tubes on a hot block (similar to the method used by Brumbaugh and others, 2005). Freeze-dried liver and kidney tissues were acid-digested with microwave-assisted heating in high-purity nitric acid and hydrogen peroxide (similar to EPA method 3050B; EPA, 1996). Sample digests were analyzed for lead concentrations using ICP–MS (similar to EPA method 6020B; EPA, 2014). All lead tissue results were reported in milligrams per kilogram on a dw basis. QC measures for ICP–MS analyses of lead in the blood, liver, and kidney samples included the use of at least four National Institute of Standards and Technology-traceable calibration standards plus a calibration blank, second-source continuing calibration standards and blanks, laboratory control solutions, certified reference materials, method and analytical duplicates and spikes, interference check solutions, and analytical dilutions (app. 2).

Determination of δALAD Activity

δALAD activity in blood samples (n=52, table 4) was determined by the method of Burch and Siegel (1971) as modified by Pain (1987) at the USGS Eastern Ecological Science Center (Laurel, Md.). The sample volume requirements for the δALAD assay were scaled down to use a volume of 25 microliters (µL) whole blood for analysis, and all samples were analyzed in duplicate. Briefly, samples were incubated pre-assay for 10 minutes at 38 °C. Freshly prepared 0.01 molar (M) aminolevulinic acid substrate (pH 6.65) was then added to each sample tube, and the mixtures were incubated for an additional 60 minutes. The assay was terminated by the addition of 10-percent trichloroacetic acid. The incubation tubes were then centrifuged at 1,500× g, and the supernatant was reacted with Erlich’s reagent. The absorbance of the reacted supernatant was measured at 555 nanometers, background corrected using the absorbance at 630 nanometers, and pathlength corrected at 90-second intervals for as much as a total of 21 minutes using a FLUOstar Omega microplate reader (BMG LABTECH Inc., Cary, North Carolina). Optimal absorbance occurred at about 15 minutes after adding Erlich’s reagent. Units of δALAD activity correspond to ΔA/RBC/h, where A is the background-corrected absorbance at 555 nanometers; RBC is the hematocrit, in milliliters; and h is the time, in hours (Burch and Siegel, 1971). Hematocrit was measured in the packed red cell fraction on a micro-hematocrit capillary tube reader after centrifugation (IEC International Micro-Capillary Centrifuge, Boston, Mass.) of a plugged capillary tube at 13,460× g for 3 minutes. The packed red cell fraction was measured on a micro-hematocrit capillary tube reader. Preliminary assays to verify the δALAD method, and details of the QC measures, are described in appendix 2. Samples were assayed in 3 batches (consisting of 6 samples, 18 samples, and 28 samples, respectively); chicken blood was assayed multiple times in each batch as a QC sample. No samples had activity or absorbance below the limit of detection (LOD; app. 2, table 2.1).

Analyses of Oxidative Stress Indicators and DNA Damage

Liver samples (n=53) were analyzed at the USGS Eastern Ecological Science Center for biomarkers of oxidative stress, namely total sulfhydryl (TSH), total glutathione (tGSH), reduced glutathione (GSH), oxidized glutathione (GSSG), protein bound sulfhydryl (PBSH), lipid peroxidation (thiobarbituric acid reactive substances [TBARS]), and DNA damage (8-hydroxy-2′-deoxyguanosine [8-OH-dG]). We also measured the ratio of GSSG to GSH (GSSG:GSH). For TSH, tGSH, GSH, GSSG, PBSH and TBARS analyses, an about 100-mg ww piece of frozen liver sample was placed in a 1.5-milliliter tube on wet ice and mixed with a 0.6-milliliter volume of 1× phosphate-buffered saline (PBS, pH 7.4; Fisher BioReagents, Waltham, Mass.). Stainless steel beads (0.9- to 2.0-millimeter blend; Next Advance, Inc., Troy, New York) were added to each tube at 1.3 times the weight of the liver piece; the samples were homogenized using a Bullet Blender (Next Advance, Inc., Troy, N.Y.). The homogenate was centrifuged at 10,000× g for 10 minutes at 4 °C and the supernatant saved. Two aliquots of each supernatant were diluted in 1× PBS at concentrations of 100 and 25 milligrams per milliliter (mg/mL), and separately frozen at −80 °C until analysis. Concentrations of GSH and tGSH were measured using the DetectX Glutathione Fluorescent Detection Kit (Arbor Assays, Ann Arbor, Michigan) following the manufacturer’s protocol. The 25 mg/mL supernatant was thawed on ice and diluted to 12.5 mg/mL in 1× PBS for the analysis of brown-headed cowbird, northern cardinal, and mourning dove samples; the 25 mg/mL supernatant was similarly thawed on ice but diluted to 6.25 mg/mL in 1× PBS for American robin samples.

Concentrations of TSH were analyzed using the Measure-iT Thiol Assay Kit (Invitrogen Molecular Probes, Eugene, Oregon) following the manufacturer’s instructions. For this assay, the 25 mg/mL supernatant from brown-headed cowbird, northern cardinal, and mourning dove samples was diluted to 8.33 mg/mL in 1× PBS; the 25 mg/mL American robin supernatant was diluted to 5.0 mg/mL in 1× PBS. TBARS concentrations were determined using the QuantiChrom TBARS Assay Kit (Bioassay Systems, Hayward, California) following the manufacturer’s instructions. For TBARS, the 100 mg/mL supernatant was used for all four species, although for some American robin samples, the supernatant had to be diluted to 12.5 mg/mL in 1× PBS. Concentrations of PBSH were calculated as the difference between TSH and GSH concentrations. GSSG (that is, [tGSH−GSH]/2), and the ratio of GSSG:GSH were calculated using the measured endpoints. Total protein concentrations were determined with a 5 micrograms per microliter supernatant (in 1× PBS) using the Pierce BCA Protein Assay Kit (ThermoFisher Scientific, Waltham, Mass.). Each sample was assayed in duplicate; any sample for which the coefficient of variation of duplicates was >10 percent was re-analyzed. No duplicate coefficient of variation was >10 percent after re-analysis. Two reference samples were run on each plate to monitor interassay variability; additional QC details are provided in appendix 2.

DNA was extracted from about 5 mg of liver using the Gentra Puregene Tissue Kit (Qiagen, Gaithersburg, Md.). The concentration and purity of DNA were determined using a NanoDrop 8000 microvolume ultraviolet-visible spectrophotometer (Thermo Fisher Scientific, Wilmington, Delaware). The samples were normalized to 15 micrograms of DNA per 100 µL in hydration solution (10 millimolar [mM] tris[hydroxymethyl]aminomethane, 1 mM ethylenediaminetetraacetic acid, pH 7–8) and frozen at −80 °C until analysis. Normalized DNA samples were thawed and denatured by heating for 10 minutes at 100 °C, cooled on ice for 5 minutes, and microfuged for 5 seconds. We then added 8.2 µL of 300 mM sodium acetate buffer (pH 5.2), 1.3 µL of 5 mM zinc chloride, and 7 µL of 25 units per milliliter Nuclease P₁ to each sample. The tubes were inverted and microfuged for 5 seconds, and then incubated at 37 °C for 30 minutes. The samples were pH-adjusted to pH 7.5 to 8.0 by adding 14 µL 1 M tris(hydroxymethyl)aminomethane -hydrochloride (pH 7.5), and then 10.2 µL of 10 units per milliliter alkaline phosphatase was added to each sample. Samples were mixed by inversion, microfuged for 5 seconds and incubated at 37 °C for 30 minutes. The alkaline phosphatase was inactivated by boiling for 10 minutes at 95 °C, and the samples were then placed on ice. Aliquots (30 µL, 0.075 microgram per microliter) were stored at −20 °C until analysis. Concentrations of 8-OH-dG were determined by enzyme-linked immunosorbent assay using the DetectX DNA Damage Immunoassay Kit (Arbor Assays, Ann Arbor, Mich.) using a BMG FLUOstar Omega microplate reader (BMG LABTECH, Inc., Cary, N.C.); additional QC information is provided in appendix 2.

Examinations of Tissues for Microscopic Lesions

Formalin-fixed liver and kidney samples were dehydrated, paraffin-embedded, sectioned at 5 micrometers, and stained with hematoxylin and eosin, Fite’s acid fast, and Ziehl-Neelsen acid fast stains for light microscopic examination (Luna, 1968) at the USGS National Wildlife Health Center (Madison, Wisconsin). The tissues were examined for microscopic abnormalities potentially associated with lead toxicosis, including acid-fast intranuclear inclusions in renal tubular epithelial cells, hepatic or renal tubular degeneration and necrosis, hepatic or renal inflammation, hepatic lipidosis, hepatic hemosiderosis, hepatic bile stasis, hepatic biliary hyperplasia, hepatic or renal cyto/karyomegaly, and renal glomerulopathy (Haschek and others, 2013). Abnormalities were scored as present or absent by a single pathologist.

Monitoring Reproductive Success

Reproductive success was monitored in eastern bluebirds, eastern towhees, field sparrows, indigo buntings, and northern cardinals (table 2). Eastern bluebirds are a secondary cavity nesting species; therefore, wooden nest boxes were erected at each contaminated and reference site to facilitate efficient reproductive monitoring and tissue collection of bluebirds. Nest boxes were constructed using the eastern/western bluebird nest box design from the North American Bluebird Society (2016); we fitted the nest boxes with predator guards to reduce nest predation (for example, by cats, raccoons, and snakes). Each nest box was placed next to open fields with nearby perches and at least 50 m of spacing between boxes in all directions (Willner and others, 1983). The number of boxes at each site was dictated by the amount of suitable habitat (table 2).

Open-cup nests of eastern towhees, field sparrows, indigo buntings, and northern cardinals were located and monitored between early April and late July in 2016, 2017, and 2018 (table 2). We visited sites every 1 to 3 days to search for nests; nests were located using a combination of parental behavior and systematic searching (Ralph and others, 1993). Each nest was designated with a unique identification code and marked with weather-resistant flagging placed ≥5 meters (m) from the nest.

Nest boxes and open-cup nests were monitored using established protocols (Ralph and others, 1993). Nests were monitored every 1 to 7 days during the breeding season (May to July), depending on nest age or stage, until the nest fledged or failed; this interval ensured optimal timing for capture and tissue sampling of adults and their nestlings. The status of the nest and its contents (in other words, the number of eggs or nestlings of conspecific young and cowbirds) were recorded during each visit. Adult birds were targeted for capture during the nestling stage. Nestlings were hand-captured from the nest for sampling about 1 to 3 days before their predicted date of fledging (Ralph and others, 1993). Disturbances to the open-cup nests were limited by quickly completing checks with as little alteration to the surrounding vegetation as possible. Nests were not checked if common nest predators, such as Cyanocitta cristata (Linnaeus, 1758) (blue jays), Corvus brachyrhynchos (C. L. Brehm, 1822) (American crows), or brown-headed cowbirds (a brood parasite), were nearby.

Nest fate was determined using the expected fledge date in conjunction with observations made at subsequent visits or at the final check. Successful fledging was confirmed by observation of at least one nestling leaving the nest, or through a combination of other cues, including adults carrying food repeatedly to the same area (likely indicating they are feeding fledged birds away from the nest), begging calls, and trails of fecal sacs/feces leading away from the nest. Visual confirmations of fledglings were always attempted, but we were also cautious to minimize disturbances to adults and their broods. If no evidence of fledging was found, we monitored the territory for any immediate re-nesting attempts, which would suggest that a predation event had occurred. Nests were considered successful even if they only fledged cowbirds because this still indicated that a nest did not fail due to predation or lack of parental care. Nests with unknown final fates were included in survival analysis, but the final monitoring interval was omitted for nests for which fate was unknown.

At each open-cup nest, several habitat features that reflected the amount of vegetative cover around a nest were measured after the nest fledged or failed (Roach and others, 2018). These variables were assessed for their potential contributions to variation in reproductive success, in addition to local lead concentrations. Nest concealment was determined by averaging the visually estimated concealment of the nest from 1 m north, east, south, west, and above the nest. Percent ground cover was estimated by averaging the visually estimated percent vegetative ground cover in the northeast, southeast, southwest, and northwest quadrants of a circle around the nest defined by a 5-m radius; the mean percent shrub cover was similarly estimated. The number of saplings (2.5- to 12.5-centimeter [cm] diameter at breast height [DBH]), pole timber (13.0- to 27.5-cm DBH), and saw timber (>27.5-cm DBH) sized trees were tallied within an 11.3-m radius of the nest; counts were converted to number of stems per hectare in each diameter class.

Statistical Analyses and Data Treatments

Lead in Soils and Invertebrates

We characterized lead concentrations in soil at over 1,000 separate points among the study sites. Lead concentrations in soil ranged from 24 to 35,967 ppm at Anschutz Mine, and from 79 to 19,100 ppm at Magmont Vent (fig. 2, table 6). The concentrations depended on the exact location of the sample point within the site, but the wide ranges of concentrations of soil-borne lead across these sites indicates heterogeneity and supports our “local lead concentration” approach for evaluation of reproductive metrics. The Big River floodplain site had a narrower range of lead concentrations in soil (21 to 3,070 ppm; fig. 2, table 6) relative to Anschutz Mine and Magmont Vent; this result was not unexpected and is consistent with the inundation of the low-lying areas within the Big River floodplain site by historical flood waters. Flood waters typically deposit fine sediments onto the floodplain surface as the waters recede; the soil sampled in areas that received historical flood waters had the greatest mean lead concentrations. Lead concentrations in soils decreased (to about 50 ppm upslope) as the sampling distance from the outer edge of the historically flooded areas increased (fig. 3).

The greatest mean lead concentration in soil was at Magmont Vent (2,172 ppm; range 79 to 19,100 ppm; table 6). Anschutz Mine had the greatest discrete lead concentration (35,943 ppm) despite having the lowest mean concentration (852 ppm) of the contaminated sites. The mean lead concentration in soil at Big River floodplain was 1,292 ppm. The Hawn State Park and Johnson’s Shut-ins State Park reference sites soils had relatively low lead concentrations (<38 ppm; table 6), and lead concentrations in soil at the Cape Girardeau reference site were 17 to 99 ppm (table 6). Mean reference site lead concentrations were 21, 31, and 36 ppm at Hawn State Park, Johnson’s Shut-ins State Park, and Cape Girardeau, respectively. We did not evaluate the soil results for statistical differences among means at the sites because of heterogeneity. The heterogeneity of lead in the soils at each site is IDW-modeled in figure 3 (Big River floodplain), figure 4 (Magmont Vent), figure 5 (Anschutz Mine), and figure 6 (reference sites); error rates in the interpolation were low to moderate (table 5), but areas of the map far from the point sample locations could have substantial uncertainty. Maps based on the interpolations demonstrated that the reference sites were spatially homogenous and had relatively low lead concentrations (fig. 6), while the contaminated sites had a greater degree of spatial heterogeneity and locally elevated lead concentrations in soil (fig. 3, fig. 4, and fig. 5).

Lead concentrations in invertebrates varied by site but were generally lowest at the reference sites and greatest at the contaminated sites (table 7). Mean concentrations ranged from 9.63 to 15.0 mg/kg dw lead at the reference sites and from 397 to 1,171 mg/kg dw at the contaminated sites. However, invertebrate sample sizes were low (n=23 total samples) and collection was performed in 2016 only; invertebrate availability may vary by season and year, which may limit the representativeness of the 2016 sample results across the 3-year duration of the study.

Lead in Bird Tissues

Blood samples from 328 adult birds and 97 broods (composites, table 4) were analyzed for lead. Adult American robin, brown-headed cowbird, eastern bluebird, eastern towhee, field sparrow, indigo bunting, northern cardinal, and mourning dove were targeted for blood collection; however, not all target species were abundant at all sites. Anschutz Mine was the only site for which blood could be collected from all eight species; and only eastern bluebirds, indigo buntings, and northern cardinals were available at all sites (table 8, fig. 7). Based on the generalized linear model with species, site, and the interaction of Species × Site as fixed factors, lead concentrations in blood in birds from contaminated sites were, on average, 10 times greater than in birds from reference sites (fig. 7). Site (F_5,288=83.15, p<0.001) and Species (F_7,288=6.38, p<0.001) had significant effects on adult lead concentrations in blood. However, the interaction between Site and Species was not significant (F_18,288=0.99, p=0.472), indicating that the effect of site was similar across species.

Inter-site comparisons were difficult to tease apart because lead concentrations in blood demonstrated a fair amount of variation within each site for each species (table 8). For example, the mean lead concentration in blood from adult indigo buntings did not differ among sites (p>0.05) in post-hoc comparisons; this was likely a result of relatively high variation in blood lead concentrations among individuals within sites (that is, the relative standard deviations for lead concentrations in blood samples ranged from 13.5 to 63.4 percent for adults and from 23.8 to 60.0 percent for broods). Eastern bluebirds at the Big River floodplain site had greater lead concentrations in blood than eastern bluebirds at Anschutz Mine (p<0.001) and the Hawn State Park and the Johnson’s Shut-ins State Park reference sites (p<0.001). Northern cardinals showed a similar pattern; the mean lead concentration in blood was greatest at Big River floodplain compared to Anschutz Mine (p<0.001) and the reference sites (p<0.001).

Based on the GLM model with log-transformed local lead concentration in soil (IDW-modeled) and species as fixed factors and site as a random factor, lead concentrations in blood generally increased with increasing local lead concentrations in soil for all species (adults and broods). Both lead concentrations in soil (SoilPb; F_1,180=29.11, p<0.001) and Species (F_4,18 =3.46, p=0.010) had effects on the lead concentration in blood of adult songbirds. However, the interaction between species and soil lead concentration was not significant (F_4,180=1.23, p=0.298), indicating that the effect of lead concentration in soil on lead concentration in blood did not vary among species (fig. 8).

A similar trend was found with nestlings; there was a significant effect of lead concentration in soil (F_1,78=55.07, p<0.001) and a marginal effect of species (F_4,78=2.45, p=0.053) on lead concentration in nestling blood. The interaction of lead concentration in soil and species was not significant (F_4,78=1.98, p=0.105), indicating that the effect of lead concentration in soil did not vary among species. The sample size of paired adult and nestling lead concentrations in blood for eastern bluebirds (n=44 nests) permitted examination of the relation between exposure in adults and their broods. There was a strong, predictive relation between the lead concentration in blood in adult eastern bluebirds and the mean lead concentration in the blood of their broods (F_1,42=88.76, R²=0.67, p<0.001; fig. 9). In other words, adults with greater lead concentrations in blood had broods with greater lead concentrations in blood.

Overall, lead concentrations in kidneys were generally greater at the contaminated sites compared to the reference sites; however, exact differences were species-specific (table 9; fig. 10). On average, lead concentrations in kidneys of birds from the contaminated sites (n=67) were 8 times greater than the lead concentrations in kidneys of birds from reference sites (n=20). There was an effect of species (F_3,86=12.687, p<0.001), site (F_5,86=30.936, p<0.001), and an interaction between species and site (F_4,86=3.164, p=0.019) on lead concentrations in adult kidneys (table 10). We also found effects of site when analyzing kidneys of American robins (F_2,34=15.051, p<0.001; table 11) and northern cardinals (F_4,32=30.594, p<0.001; table 12); kidneys from individuals at contaminated sites had greater lead concentrations in their kidneys relative to those at the reference sites. The mean lead concentration in kidneys of northern cardinals at Big River floodplain (18.4 mg/kg dw) was greater than the mean lead concentrations in kidneys at all other sites (0.26 to 2.80 mg/kg dw; tables 9 and 13); the northern cardinals sampled at Johnson’s Shut-ins State Park had the lowest mean lead concentration in kidneys (0.26 mg/kg dw).

Similarly, mean lead concentrations in liver tissues were greater at the contaminated sites compared to the reference sites (fig. 11; table 9); there was a significant effect of site (F_5,86=15.65, p<0.001) and an interaction between species and site (F_4,86=2.70, p=0.037) (table 14). However, species alone had no effect (F_3,86=1.796, p=0.155; table 14). There was also an effect of site (F_4,32=13.866, p<0.001) when examining lead concentrations in livers of northern cardinals only (table 15). The mean lead concentration in livers of northern cardinals was greatest at the Big River floodplain (12.1 mg/kg dw), followed by Magmont Vent (2.55 mg/kg dw), the Anschutz Mine (1.97 mg/kg dw), Hawn State Park (0.85 mg/kg dw), and, finally, Johnson’s Shut-ins State Park (0.26 mg/kg dw; table 9) sites. However, there were no significant differences between lead concentrations in livers of northern cardinals between Hawn State Park and Johnson’s Shut-ins State Park, nor were there differences between lead concentrations in liver at Anschutz Mine and Magmont Vent (table 16). We also found site effects for lead concentrations in livers of American robins (F_2,34=20.349, p<0.001; table 17); mean lead concentrations in liver were greatest at Anschutz Mine (7.62 mg/kg dw) and least at the Cape Girardeau reference site (0.96 mg/kg dw; table 9).

δALAD Activity

When averaging values across all species, blood δALAD activity in samples from reference sites (overall mean of 190.3 activity units) was greater (Wilcoxon non-parametric test, Z=4.3738, χ²=19.2158, p<0.001) than activity in birds from contaminated sites (Kaplan-Meier mean of 88.4 to 89.3; table 18). Two of 19 individuals from the reference sites had δALAD activity values <50 percent (that is, <95.2 activity units) of the overall mean (190.3 units) activity for the reference sites, whereas 18 of 34 (53 percent) individuals at contaminated sites had δALAD activity values <50 percent of the overall mean activity (190.3 units) of those at the reference sites. Likewise, δALAD activity of northern cardinals (n=25 total animals) at the reference sites (mean=204.2 activity units) was greater (Wilcoxon non-parametric test, Z=3.2559, χ²=10.7862, p=0.001) than activity of cardinals from the contaminated sites (Kaplan-Meier mean of 88.4 to 90.1; table 18). None of the northern cardinals from the reference sites had activity <50 percent (that is, <102.1 activity units) of the mean activity of cardinals from reference sites, whereas 8 of 16 (50 percent) of the northern cardinals from the contaminated sites had δALAD activity <50 percent of the mean for cardinals from reference sites.

Using samples from 51 individuals (all 4 species combined), regression ANOVA indicated a relation (p<0.001) between blood δALAD activity (NDs assigned LOD) and log-transformed lead concentration in blood (F_1,49=57.8 for NDs assigned 0.001 as a default value; F_1,49=113.8 for NDs assigned LOQ as a default value), kidney (F_1,49=55.7 for NDs assigned 0.001; F_1,49=94.5 for NDs assigned LOQ) and liver (F_1,49=20.0 for NDs assigned 0.001; F_1,49=40.5 for NDs assigned LOQ) tissues. Visual inspection of blood δALAD and tissue lead plots (fig. 12) indicated an inverse relation for activity and lead concentration. Adjusted coefficient of determination (R²) values for these relations ranged from 0.28 to 0.69 (table 19).

When only the δALAD results for northern cardinals were considered, regression ANOVA demonstrated a similar relation (p<0.001) between blood δALAD activity (NDs assigned LOD) and log-transformed lead concentrations in blood (F_1,23=51.1 for NDs assigned 0.001 as a default value; F_1,23=169.1 for NDs assigned LOQ as a default value), kidney (F_1,23=52.4 for NDs assigned 0.001; F_1,23=248.8 for NDs assigned LOQ), and liver (F_1,23=18.2 for NDs assigned 0.001; F_1,23=51.6 for NDs assigned LOQ) tissues. The adjusted R² values for these relations in northern cardinals ranged from 0.417 to 0.912 and were seemingly more robust (that is, greater R² values) than the relations for all four species combined (table 19).

Oxidative Stress Indicators and DNA Damage

Statistical analyses of species and site-type differences in the targeted endpoints (biochemical markers TBARS, GSH, tGSH, GSSG, GSSG:GSH, TSH, PBSH, 8-OH-dG) were limited to northern cardinals and American robins. Analyses could not be performed on data for mourning doves and brown-headed cowbirds due to small sample size (n=1 each). For American robins and cardinals, we analyzed the effects of species, site type (reference versus contaminated sites), and their interaction. We found effects of both factors (species M=195.430, p<0.001; site-type M=51.003, p=0.007) and a marginal interaction (M=25.796, p=0.069). Differences between species (from contaminated versus reference sites) were observed for GSH (p<0.001), tGSH (p=0.007), GSSG (p=0.002), GSSG:GSH ratio (p<0.001), TSH (p<0.001), PBSH (p<0.001), and TBARS (p=0.005) (fig. 13), but not for DNA damage (not shown).

Data for site-type effects were analyzed separately for American robins and northern cardinals (table 20). No differences between site type (contaminated versus reference) were observed for American robins for any endpoints. For northern cardinals, concentrations of GSH (p=0.047), GSSG:GSH ratio (p=0.032), TSH (p=0.032), and PBSH (p=0.047) differed between contaminated and reference samples. Concentrations of GSH, TSH, and PBSH were greater in northern cardinals from contaminated sites compared to reference cardinals, whereas the GSSG:GSH ratio was greater in reference northern cardinals compared to northern cardinals from lead-contaminated sites (table 20). No differences were observed in TBARS or levels of DNA damage (8-OH-dG) between reference and contaminated samples.

A positive relation was observed in American robins between log-transformed lead concentrations in liver and GSH (F_1,15=5.11, p=0.039) and tGSH (F_1,15=8.758, p=0.010) and between log-transformed lead concentrations in kidneys and tGSH (F_1,15=8.182, p=0.012) (fig. 14). In northern cardinals, we observed positive relations between log-transformed lead concentrations in kidneys and GSH (F_1,24=11.45, p=0.002), tGSH (F_1,24=3.653, p=0.068), TSH (F_1,24=6.29, p=0.019), and PBSH (ρ=0.445, p=0.023); there was a negative relation between lead concentrations in kidneys and the log-transformed ratio of GSSG:GSH (F_1,24=7.393, p=0.012) (fig. 15). Similarly, we observed positive relations in northern cardinals between log-transformed lead concentrations in livers and GSH (F_1,24=12.2, p=0.002), TSH (F_1,24=7.634, p=0.011), and PBSH (F_1,24=5.279, p=0.030); there was a negative relation between lead concentrations in livers and the log-transformed GSSG:GSH ratio (F_1,24=8.52, p=0.008) (fig. 16). In brown-headed cowbirds, no relations were observed between lead concentrations in livers or kidneys and any of the monitored endpoints. No relations between 8-OH-dG or TBARS with lead concentrations in either liver or kidney tissues were found in any species.

Microscopic Lesions in Liver and Kidney Tissues

The prevalence of microscopic changes within the livers and kidneys of birds from contaminated and reference sites was similar (table 21), and it was not possible to definitively link lesion occurrence with lead concentrations in blood or tissues in individual birds (individual bird data not shown; see Cleveland and others, 2023). Microscopic lesions in birds from contaminated and reference sites were often tied to parasitic infections, including inflammation and degeneration or necrosis in the kidneys and livers with intralesional protozoa in northern cardinals, and hepatic necrosis with nematode parasites in American robins. Other microscopic changes were consistent with common background lesions in wild birds. These included inflammation in the livers and kidneys of mourning doves and inflammation and necrosis in the livers of brown-headed cowbirds, particularly at Magmont Vent. The single brown-headed cowbird available from a reference site also had inflammation in its liver. We did not observe any acid-fast intranuclear inclusions in renal tubular epithelial cells in any bird.

Effects of Lead on Reproductive Success

The numbers of nests monitored for each species are reported in table 2; we monitored 95 eastern bluebird nests and 484 nests of the 4 open-cup species (eastern towhee, field sparrow, indigo bunting, northern cardinal). Twenty-two of 95 eastern bluebird nests failed (23 percent); 15 failures were due to abandonment, 5 were due to predation, and 2 failed due to weather-related events. In total, 320 open-cup nests failed (65 percent): 34 to abandonment, 283 to predation, and 3 to weather-related events. Also, 19 of the 484 open-cup nests were parasitized and had 1 brown-headed cowbird egg or nestling in them.

Descriptive statistics indicated that, based on successful nests, means for clutch size, proportion hatched, number of young fledged based on all nests, and number of young fledged were greater at reference sites than at contaminated sites in 17 of 20 comparisons (table 22). Nonetheless, we used more robust statistical inferences, based on GLM model results, to examine the effects of lead concentrations in soil and blood on reproductive success because the GLM models directly relate reproductive success to lead concentrations while controlling for site and year effects. Mean local lead concentrations in soil around nests ranged from 20 to 4,468 ppm for eastern bluebird; 19 to 4,165 ppm for eastern towhee; 19 to 4,295 ppm for field sparrow; and 19 to 4,280 ppm for indigo bunting. Mean local lead concentrations for northern cardinal nests ranged from 19 to 4,297 ppm, except for a single extreme value of 12,488 ppm. We removed this single extreme lead concentration from further analysis of bird reproductive success because it only occurred for one individual at a single point location, was extreme compared to all other values, and likely resulted from a unique, highly localized condition.

Nest Success

We monitored the success of 95 eastern bluebird nests over a total of 574 nest-check intervals. The relation between eastern bluebird nest success and lead concentration in blood could not be estimated because the model would not converge on a solution. This was likely because of a relatively small sample size (number of nests with BloodPb values=66) and a high nest success rate (only 22 of 95 nests failed). We successfully fit a model for the effect of SoilPb on nest success of eastern bluebirds, and there was no support for the effect of local lead concentration. The model with no SoilPb effect had a lower DIC than the model with SoilPb (ΔDIC=1.22). The credible interval for the mean effect of SoilPb overlapped zero, and the posterior distributions indicated a 0.56 probability of a negative effect of SoilPb (table 24). Mean daily nest survival of eastern bluebirds was 0.998, and the period nest survival was 0.928 (table 24).

We found support for an effect of local lead concentration in soil on nest survival of some open-cup nesting birds; in other words, the nest survival of some species decreased as the local lead concentration in soil increased. We monitored success of 484 open-cup nests over a total of 2,084 nest-check intervals; the number of nests ranged from 52 to 184 per species. Mean daily nest survival of open-cup nesting species was 0.913 to 0.934 (table 25), and the period nest survival was 0.095 to 0.202. The model with SoilPb effects had more support than a model without SoilPb effects (ΔDIC=2.04). The credible intervals for the mean effect of SoilPb on each species overlapped zero for all species except indigo bunting. However, posterior distributions indicated a 0.97, 0.99, and 0.67 probability of a negative effect of SoilPb on eastern towhees, indigo buntings, and northern cardinals, respectively (table 25). As a result of the predicted effect of SoilPb, period nest survival for eastern towhees and indigo buntings decreased from >20 to <5 percent as local lead concentration in soil increased from 18 to 4,400 ppm (fig. 17). Period nest survival for northern cardinals decreased from 23 to 14 percent as the local lead concentration in soil increased from 18 to 4,400 ppm (fig. 17). Conversely, nest survival of field sparrows increased from 20 to 27 percent as lead concentrations in soil increased from 18 to 4,400 ppm (fig. 17).

Habitat Effects on Reproductive Success

Nest survival was the only reproductive success measure for which there was statistically significant support for an effect of local lead concentration in soil; therefore, we considered whether any of the habitat variables explained additional variation in nest survival. However, the addition of nest concealment, percent ground cover, percent shrub cover, number of sapling-sized trees, number of pole timber-sized trees, or number of saw timber-sized trees to the nest survival model for open-cup nesting birds did not improve the model. The ΔDIC value for the addition of each variable ranged from 0.01 to 1.93, indicating that the model with a habitat effect added was no better or worse than the model without the habitat effect; moreover, addition of habitat effect did not help explain variation in nest survival. Furthermore, no habitat effects had credible intervals that did not overlap zero. Although habitat did not explain additional variation in nest survival, there were some patterns in habitat variables around nests based on examination of descriptive statistics (table 26). Percent ground cover was consistently lower for all species on contaminated sites (31 to 53 percent) than on reference sites (43 to 84 percent). In addition, numbers of sapling-sized trees were consistently lower for all species on reference sites (120 to 842) than on contaminated sites (423 to 1,187). The remaining habitat variables did not vary consistently among species between reference and contaminated sites, and the 95-percent confidence intervals for the means of all habitat variables overlapped between reference and contaminated sites (table 26).

Lead in Soil

Beyer and others (2013) concluded that lead concentrations ≥1,000 mg/kg dw (assumed to be equivalent to 1,000 ppm measured by pXRF) in soil in southeast Missouri were sufficient to expose ground-foraging birds to toxic lead concentrations, as evidenced by δALAD inhibition, elevated lead concentrations in tissues, and the presence of renal acid-fast inclusions in birds from contaminated sites compared to reference sites. None of our reference soils had lead concentrations that exceeded the 1,000 mg/kg dw threshold, whereas a total of 17 percent (n=66 of 394) of the soil samples at Anschutz Mine, 74 percent (n=258 of 347) at Magmont Vent, and 70 percent (n=258 of 370) at Big River floodplain had lead concentrations that exceeded the 1,000 ppm threshold by factors of as much as 35, 19, and 3, respectively (table 6).

However, the mean lead concentration in soil only exceeded 1,000 ppm at Big River floodplain and Magmont Vent (table 6). This discrepancy between mean concentrations, concentration ranges, and threshold exceedances further demonstrates the heterogeneity of lead in soils across the contaminated sites and supports the use of local lead concentrations in soil for assessment of reproductive effects. In other words, a site-wide mean soil lead concentration calculated for a heterogenous site may overestimate or underestimate effects depending on the exact location of a nest within a site, whereas raster grids estimate the lead concentrations in soil within the nesting territories, such that effects can be more closely linked to exposure.

We found support for some species effects in the relation of lead concentration in blood to lead concentration in soil and weak to moderate support for interactions between species and lead concentration in soil. Therefore, although lead concentration in blood differed among species for a given lead concentration in soil, the rate at which lead concentrations in blood increased in relation to lead concentrations in soil was similar across species and sites. Mean lead concentrations in blood ranged from <0.1 mg/kg dw in multiple species to 4.76 mg/kg dw in eastern bluebird broods (table 8). Differences in the lead concentrations in blood among species likely resulted from variability in interspecific foraging behaviors, diet, rate of soil ingestion, and local lead concentrations in soil (Beyer and others, 1994; Hansen and others, 2011). For example, at the Anschutz Mine site, American robins had a mean lead concentration in blood of 2.04 mg/kg dw, and northern cardinals had a mean concentration of lead in blood of 0.61 mg/kg dw (table 8). Both species are ground and arboreal foragers (table 1), but cardinals are generally opportunistic and will consume buds and insects on trees as the canopy develops (Halkin and Linville, 1999). In contrast, American robins feed heavily on earthworms during the breeding season (Beyer and Sample, 2017). In addition, high-protein diets may be protective against lead absorption (Franson and Pain, 2011); for example, Kendall and others (1982) observed that lead-dosed ringed turtle doves fed a mash diet with 27 percent protein had substantially lower lead concentrations in their tissues compared to birds fed corn (low protein). The authors hypothesized that protein-induced secretion of pepsin in the gastrointestinal tract and the thiol groups on the pepsin molecules chelated the lead as it dissolved.

Previous studies have collected and analyzed ingesta in euthanized adult birds to estimate the rate of incidental soil ingestion (and subsequently ingestion of lead in soil) by individuals (Grue and others, 1984, 198672; Blus and others, 1995). Hansen and others (2011) found that soil ingestion rate differs greatly among species; American robins had an estimated soil ingestion rate of 20 percent, whereas Swainson’s thrush had a soil ingestion rate of 0.7 percent. Both American robins and Swainson’s thrush forage on the ground; however, Swainson’s thrush gleans food items from litter and foliage, whereas the foraging strategy of robins includes probing their bills into the soil. Further, soil ingestion rates likely fluctuate daily and seasonally as individuals alter their diet in response to food availability. For example, the proportion of fruit in the diet of the American robin is much greater in the fall and winter compared to the spring and summer (Wheelwright, 1986). Precipitation events may also affect soil ingestion rates; earthworms emerge during rain, becoming more readily available to birds. Rainwater may wash the exterior of the worms, reducing the amount of soil ingested by the bird; in contrast, immediately after rains, the soils may more readily stick to the wet earthworm bodies, increasing the amount of soil ingested by the bird. However, we did not evaluate the overall dietary composition of our study birds, nor did we evaluate the amount of external or internal sediment on the invertebrate bodies.

Variability in lead concentrations in blood among species and individuals can also depend on the bioavailability of the lead, particularly because of lead speciation and solubility (Hemphill and others, 1991; Kastury and others, 2019). Factors that affect lead bioavailability in soils include the lead concentration, the organic carbon concentration in the soil, soil pH, clay content, cation exchange capacity, particle size, and other predominant cations and anions in the soil (Hettiarachchi and Pierzynski, 2004; Ming and others, 2012; Yan and others, 2017). We believe that spatial differences in lead bioavailability within and among sites likely affected the uptake of soil-borne lead in our study animals.

Relative absorption rates of dietary lead are also affected by the lead compound (for example, metallic lead versus lead salts); for example, kidney absorption of lead from lead carbonate is twelvefold that of lead from metallic lead and 1.6 times that of lead from lead acetate in rats (Barltrop and Meek, 1975). Lead carbonate and lead acetate are readily soluble in the digestive system and, therefore, generally have a high degree of bioavailability via gastrointestinal absorption (Casteel and others, 2006; Scheckel and others, 2009); indeed, lead acetate is considered to have 100 percent relative bioavailability (RBA). Lead speciation and bioavailability studies have been previously conducted on soils in mining districts in Missouri, including the Southeast Missouri Lead Mining District (Yang and others, 2002; Casteel and others, 2006; Weber and others, 2015; Beyer and others, 2016). Lead-contaminated soils in Missouri generally have elevated concentrations of lead carbonate because the soils are weathered from limestone or dolomite, or from limestone or dolomite forming the host rock for the ore; these elevated concentrations resulted in the surficial presence of weatherable, soluble lead-bearing minerals in the mine/mill wastes (Weber and others, 2015).

Beyer and others (2016) dosed Japanese quail with soil from several lead-contaminated sites from around the United States, including from the Big River floodplain (Washington State Park, Desoto, Mo.) and Magmont Vent sites. Lead concentrations in blood of quail that had been dosed with lead acetate (100 percent RBA), were compared with lead concentrations in blood of quail dosed with field-contaminated soils. The mean lead concentration in blood samples from quail fed lead acetate (41 mg/kg dw lead in feed; 100 percent RBA) was 1.25 mg/kg dw. When soil from the Big River floodplain and Magmont Vent study sites were added to control diets, the mean lead concentrations in blood in quail increased from 0.019 mg/kg dw (with the control diet) to 1.65 mg/kg dw (86 mg/kg dw lead in feed) and to 2.26 mg/kg dw (137 mg/kg dw lead in feed), respectively (Beyer and others, 2016). The lead in soil at Big River floodplain had 63 percent RBA, and the Magmont Vent soil had 54 percent RBA; soils from Anschutz Mine were not evaluated as part of that study. The authors noted differences in minerology and co-occurring contaminants between the Big River floodplain and Magmont Vent soils, which may have affected the RBAs (Beyer and others, 2016). The mean lead concentrations in the blood of birds at the Big River floodplain and Magmont Vent sites measured in the present study (0.27 to 6.34 mg/kg dw; table 8) were generally consistent with those quantified by Beyer and others (2016), which supports the hypothesis that our target birds were exposed to bioavailable lead in soils in the Southeast Missouri Lead Mining District. Although we did not evaluate the mineralogy of our soil samples, differences among study sites likely affected our results in a similar manner to Beyer and others (2016). A discussion of the other potentially toxic metals (cadmium, zinc, copper, nickel, and cobalt) in soil, invertebrates, and blood at the study sites and their potential effects on the toxicity and bioavailability of lead is presented in appendix 3.

We used a simplified bioavailability metric that calculated the ratios of mean lead concentrations of blood to mean lead concentrations in soil (BloodPb:SoilPb, termed “bioavailability ratio”) for species common to all sites (eastern bluebird, indigo bunting, and northern cardinal) to compare relative uptake of lead. Birds exposed to Big River floodplain soils had the greatest mean bioavailability ratios (0.001 to 0.004). The bioavailability ratios at Anschutz Mine and Magmont Vent were less than or equal to 0.001. These results suggest that lead in Big River floodplain soils had greater bioavailability to birds than soils from the other sites, and this finding is consistent with the laboratory studies performed by Beyer and others (2016).

The EPA performed a site-specific baseline ecological risk assessment (BERA) for the Big River Mine Tailings (Superfund) Site (EPA, 2006; https://cumulis.epa.gov/supercpad/cursites/csitinfo.cfm?id=0701639) and an ecological risk assessment (ERA) for Southwest Jefferson County Mining (Superfund) Site (EPA, 2020; https://cumulis.epa.gov/supercpad/cursites/csitinfo.cfm?id=0705443) to evaluate potential sources of contamination to the Big River watershed and floodplain. Both sites are in Missouri’s Old Lead Belt. Ecological risk assessments use site-specific media (in other words, earthworms, soil, water, vegetation, and so on) and literature-based ingestion rates for specific receptors to determine how the average daily dose of lead (or other target metals) at the site compares to literature-based values for the no-observed-adverse-effect level and the lowest-observed-adverse-effect level (LOAEL) toxicity reference values (TRVs) (EPA, 1997).

For the Big River Mine Tailings Site BERA (EPA, 2006), a LOAEL TRV dose of 11.3 milligrams of lead per kilogram of body weight per day (mg/kg BW/d) was used, which Edens and others (1976) found resulted in a decrease in egg hatching success in Japanese quail. The BERA also found that site-specific lead concentrations in soil (353 mg/kg dw) and depurated earthworms (126 mg/kg dw) for the Big River Mine Tailings Site (St. Francois County, Mo.) exceeded the LOAEL TRV for Scolopax minor (J. F. Gmelin, 1789) (American woodcock) by a factor of nine. In other words, the site presents a risk to avian vermivores because of ingestion of lead in earthworms and soil when lead concentrations in soil and earthworms exceed 353 mg/kg dw and 126 mg/kg dw, respectively. We found that >70 percent of soil samples (n=787 of 1,110) collected from our contaminated study sites exceeded this 353 mg/kg lead threshold; 100 percent (n=15 of 15) of invertebrates from the contaminated sites exceeded the 126 mg/kg dw threshold. Therefore, the birds at our contaminated study sites were likely adversely affected by soil- and invertebrate-borne lead, assuming soil ingestion rates similar to the defaults used in the ERA models.

The Southwest Jefferson County Mining (Superfund) Site ERA (EPA, 2020) differed from the Big River Mine Tailings Site BERA (EPA, 2006) in that for the former, the EPA (1) used nondepurated, unwashed worms to determine the ADD, (2) included American robins in their avian risk analysis, and (3) used a greater LOAEL TRV (27.4 mg/kg BW/d lead) than was used for the Big River Mine Tailings Site BERA (11.3 mg/kg BW/d lead). The increased LOAEL TRV value (27.4 mg/kg BW/d) was based on egg production in chickens (Sample and others, 2019). The earthworm concentration used for the average daily dose in the Southwest Jefferson County Mining (Superfund) Site ERA was 499.2 mg/kg dw lead. The resulting modeled dose to the American robin (using 499.2 mg/kg dw lead in undepurated earthworms) exceeded their adverse reproductive effects threshold for lead (27.4 mg/kg BW/d) by a factor of four. We found that >70 percent of the invertebrate samples collected from our contaminated study sites (n=12 of 17) exceeded the worm concentration (499.2 mg/kg dw) used in the Southwest Jefferson County Mining (Superfund) Site ERA, whereas no invertebrates from the reference site invertebrates exceeded the threshold (table 7).

Lead in Bird Tissues

Few published studies have reported lead concentrations in blood of songbird species. Indeed, this is the first known analysis of lead in brood blood for eastern towhees, field sparrows, indigo buntings, and northern cardinals. Broods generally had greater mean lead concentrations in blood at the contaminated sites; however, these values were not statistically tested because of small sample sizes for broods (table 8). Broods in 8 of 12 comparisons at the contaminated sites had greater mean lead concentrations in blood than their parents; however, the concentration ranges were similar between the age classes (table 8).

Adults have a restricted foraging range while tending to young in the nest. In this way, uptake of lead during the breeding season, as reflected by increased lead concentrations in blood, are likely caused by local exposure to lead contamination, rather than offsite exposures. Elevated lead concentrations in broods at contaminated study sites compared to the reference sites further demonstrate that exposure to and uptake of lead occurred locally, within the breeding territory. In other words, it is expected that broods would primarily be exposed to lead via ingestion of lead-contaminated food items collected by the adults within the breeding territory and around the nest location. Therefore, adults and their broods would be expected to have similar levels of exposure to local lead, resulting in similar lead concentrations in the blood.

Seasonal variations in diet and subsequent incidental soil ingestion may also result in differing exposures to lead, and temporal changes in lead concentrations in blood, among individual adults and their broods. For example, a study in Michigan found that earthworms constituted 10 percent of bluebird nestling diet in May but were completely absent from the diet in July (Pinkowski, 1978). Additionally, Pinkowski (1978) found that nestlings most-commonly consumed ground-dwelling spiders in the spring and plant-dwelling spiders in the summer. As previously mentioned, the diet of nestlings is restricted spatially and temporally; in this way, broods hatching earlier in the breeding season, when adults might provide nestlings with more ground-dwelling invertebrates, may have had more exposure to soil-borne lead that, consequently, caused greater lead concentrations in brood blood. However, we did not evaluate lead concentrations in blood temporally according to hatching date.

Currently (2022), no literature-based thresholds exist for harm to Passeriformes (songbirds) based on lead concentrations in their blood. Therefore, in accordance with previous songbird studies (Hansen and others, 2011; Sample and others, 2011), we used Franson and Pain’s (2011) suggested interpretations of lead concentrations in tissue (blood, livers, and kidneys) for the orders Anseriformes (waterfowl), Falconiformes (raptors), and Columbiformes (pigeons and doves) (table 27) to assess the potential for adverse effects. Note that the exact concentrations in each of the threshold ranges (no evidence of lead poisoning, subclinical poisoning, clinical poisoning, or severe clinical poisoning) depends on the avian order (Anseriformes, Falconiformes, or Columbiformes); the most sensitive order is Falconiformes. The clinical poisoning category represents the lead concentration in a tissue at which clinical manifestations of lead poisoning (for example, anemia, weight loss, and muscle incoordination) would be expected in adult birds. However, avian development stage can affect the extent of harm. For example, nestlings and fledglings are hypothesized to be more sensitive to lead exposure than adults, so adverse effects after exposure to lead may be more profound in broods compared to adults. Therefore, interpreting brood blood concentrations according to Franson and Pain (2011) may not be entirely suitable and may underestimate adverse effects in broods (and indeed not only for songbird broods but for Anseriformes, Falconiformes, and Columbiformes broods as well), but we hypothesize that it is a useful way to examine and compare nestling and adult exposure and is deserving of further investigation. In addition, we note that diet, size, metabolism, behavior, and habitat differ between songbirds and Anseriformes, Falconiformes, and Columbiformes; these differences, particularly increased exposures to soil-borne lead, may result in underestimation of risk to songbirds relative to other orders when using the interpretative ranges. Wild birds may also be more susceptible to lead toxicoses than the test birds and controlled exposure scenarios used to create these thresholds because of inadequate diets, stressors, and other factors (Franson and Pain, 2011).

Blood concentrations <0.9 mg/kg dw lead are generally considered to be background (assuming 78.26 percent moisture; table 27). About 97 percent (n=102 of 105) of adult blood samples and 98 percent (n=48 of 49) of brood samples at the reference sites had lead concentrations consistent with background (table 27). The exceptions were 2 indigo bunting adults (1.80 and 2.06 mg/kg dw) at Johnson’s Shut-ins State Park, 1 American robin adult (0.91 mg/kg dw) at Cape Girardeau, and 1 eastern bluebird brood (1.25 mg/kg dw) at Hawn State Park. At the contaminated sites, lead concentrations in blood in 39.7 percent (n=85 of 214) of adult birds and 45.8 percent (n=22 of 48) of broods were consistent with background; 60.3 percent (n=129 of 214) of adult birds and 54.2 percent (n=26 of 48) of broods had lead concentrations in blood that were consistent with some level of lead poisoning (that is, subclinical, clinical, or severe clinical). Eight adults and two brood blood samples at the contaminated sites (table 27) had lead concentrations consistent with severe clinical poisoning.

Lead concentrations in 40.3 percent (n=27 of 67) of livers of adult birds at the contaminated sites were consistent with some degree of lead poisoning (subclinical, clinical, or severe clinical; table 27); 1 individual had a lead concentration in liver associated with severe clinical effects. A total of 59.7 percent (n=40 of 67) of adults had lead concentrations in liver consistent with background (<6.2 mg/kg dw, assuming 67.74 percent moisture); in contrast, 100 percent of all reference livers had lead concentrations consistent with background.

Lead concentrations in 25.4 percent (n=17 of 67) of adult kidneys from the contaminated sites were consistent with background concentrations (<6.2 mg/kg dw; table 27), whereas 95 percent (n=19 of 20) of reference kidneys had background lead concentrations. About 75 percent (n=50 of 67) of kidneys from contaminated sites had lead concentrations associated with adverse effects, and as much as 38.8 percent of adult kidneys had lead concentrations associated with severe clinical poisoning. We note that although we did not observe signs of clinical or severe clinical lead poisoning (for example, wing droop, lethargy, or emaciation) in our study birds, the biological significance of tissue-based lead concentrations in birds may be difficult to ascertain for an individual bird with unknown history and chronicity of lead exposure (Franson and Pain, 2011). Many factors affect the absorption of lead in tissues. However, the results for lead in blood, paired with results for δALAD activity and other markers of biochemical damage (described in subsequent sections), suggests injury to birds at our contaminated study sites.

We also examined the interpretative benchmarks by species and site (table 28). Using the blood benchmarks, we found that several adults or broods were categorized as having been severely clinically poisoned; these individuals were sampled at Anschutz Mine (1 adult American robin and 1 adult mourning dove), Big River floodplain (1 adult eastern bluebird, 2 eastern bluebird broods, and 4 adult northern cardinals), and Magmont Vent (1 adult mourning dove). Adult birds having lead concentrations in blood suggestive of clinical lead poisoning were found at all three contaminated sites (table 28). Further, each contaminated site had at least one brood having a lead concentration in blood suggestive of some degree of poisoning (that is, subclinical, clinical, or severe clinical). Clinically poisoned broods were found at Anschutz Mine (1 eastern bluebird brood and 2 field sparrow broods), Big River floodplain (1 indigo bunting brood), and Magmont Vent (1 eastern bluebird brood and 1 eastern towhee brood); subclinically poisoned broods were found at Anschutz Mine (5 eastern bluebird broods, 1 eastern towhee brood, 2 field sparrow broods, and 1 northern cardinal brood), Big River floodplain (1 indigo bunting brood), and Magmont Vent (4 eastern bluebird broods, 1 field sparrow brood, and 2 northern cardinal broods).

We found that the use of the liver and kidney poisoning benchmarks (Franson and Pain, 2011) changed not only the numbers of individuals considered to have been adversely affected compared to the blood benchmarks (table 28), but also the number of exceedances varied by choice of liver or kidney benchmark. For example, 1 American robin adult had lead concentration in its blood that was suggestive of severe clinical poisoning, no American robin livers had lead concentrations indicative of severe poisoning, and 12 American robin kidneys had lead concentrations suggestive of severe poisoning. It is likely that lead in the blood was rapidly distributed to various other storage sites (for example, liver, bone, and kidneys; Eisler, 1988; De Francisco and others, 2003; Berglund and others, 2011). In addition, lead ingestion and uptake rates likely varied daily. Nevertheless, the measured lead concentrations in the tissues (blood, livers, and kidneys) of our study birds indicate that adults and their broods are exposed to and ingest lead at the contaminated sites, and these lead concentrations are potentially harmful.

δALAD Activity

The first measurable effects of lead absorption into the avian bloodstream is inhibition of several enzymes involved in heme synthesis, including δALAD, and enzyme inhibition can occur at <0.2 mg/kg dw lead in blood (assuming 78.26 percent moisture; Franson and Pain, 2011). δALAD activity is also inhibited in other tissues, including the liver, kidneys, and brain; moreover, the assay is not affected by age, sex, or breeding condition of birds (Finley and Dieter, 1978), which is why δALAD activity inhibition is often used as a sensitive biomarker of acute and chronic lead exposures. In total, 18 of 34 samples (53 percent) for all species combined at contaminated sites, and 8 of 16 samples (50 percent) for northern cardinals at contaminated sites, had δALAD activity <50 percent of that observed at the reference sites (table 18). Dieter and Finley (1979) observed a substantial increase of butyrylcholinesterase, a marker suggestive of brain damage, at 75 percent δALAD activity inhibition in blood; 13 of 34 study birds (38 percent) at the contaminated sites had >75 percent δALAD inhibition relative to the mean δALAD activity at the reference sites.

These findings also indicate the characteristic inverse relation between δALAD activity and lead concentrations in tissues (blood, livers, and kidneys) of the study species (fig. 12; tables 18 and 19); this is a well-known relation described in many studies (for example, Franson and Pain, 2011) that results from displacement of zinc, and replacement by lead, on the sulfhydryl group at the active site of the δALAD enzyme. Inhibition of δALAD activity can persist for weeks to months after exposure (Franson and Pain, 2011); further, although birds can tolerate some reductions in blood and tissue δALAD activity without effects on hematocrit and hemoglobin concentration or the presence of overt signs of lead intoxication, anemia can result from chronic δALAD inhibition. Lead inhibition of δALAD activity also results in the buildup of excess aminolevulinic acid substrate, which is associated with oxidative damage and cell death (Scinicariello and others, 2007). The results for δALAD, coupled with those for lead concentrations in the blood, liver, and kidney tissues of the study birds indicate that birds nesting at the contaminated sites are likely to experience adverse outcomes.

Oxidative Stress Indicators and DNA Damage

Environmental contaminants, like lead, can overwhelm the protective antioxidant systems that protect birds against normal oxidative processes (for example, those required for metabolism, immune defense), leading to a buildup of reactive oxygen species (ROS) that damage biomolecules (for example, cell injury or death, and DNA damage) and can affect immune function, growth, survival, and reproductive success (Koivula and Eeva, 2010). We observed similar patterns of expression of oxidative markers relative to reference samples between northern cardinals and American robins (table 20), although concentrations of the oxidative markers differed between the 2 species. Despite statistical differences only being clear for northern cardinals, concentrations of GSH, TSH, and PBSH in both species were greater in livers from contaminated sites compared to those from the reference sites. Lead binds the sulfhydryl groups of GSH and other thiols, depleting their stores in the liver and thereby interfering with their antioxidant activities (Ahamed and Siddiqui, 2007). Therefore, the increases that we observed may be indicative of compensatory measures taken to counter increased levels of ROS. Increased synthesis of GSH and other thiols may allow these birds to survive despite chronic exposure to elevated lead concentrations. Monaghan and others (2009) reported that prolonged elevation of ROS can cause a permanent increase in baseline antioxidant levels. This is supported by the positive relations that we observed between these indicators of thiol status (GSH, tGSH, TSH, PBSH) and lead concentrations in liver and kidney tissues. Furthermore, the GSSG:GSH ratio was lower in northern cardinals from contaminated locations compared to those from reference sites, largely because of greater GSH concentrations.

Basal (reference site) levels of GSH, tGSH, and TBARS were greater in American robins than in northern cardinals, whereas GSSG levels were lower. Differences in oxidative stress-related endpoints between bird species have been reported previously. For example, Guan (2010) found that GSH and TBARS differed among four avian species and postulated that longevity of the species may play a role in these differences. Isaksson and others (2017) reported differences in lipid peroxidation between four species of songbirds and attributed these in part to dietary variations that affect intake of antioxidants. Similarly, Rainio and others (2013) reported differences among songbird species for some indicators of oxidative stress but not for tGSH or the ratio of GSH:GSSG. This suggests that species may use varying strategies to regulate their oxidative state. Intra- and inter-specific differences in susceptibility to lead have also been noted in birds (Mateo and Hoffman, 2001).

Previous studies have shown substantially greater levels of DNA damage and lipid peroxidation in animals exposed to lead compared to reference animals (Mateo and others, 2003; Matović and others, 2015; Vallverdú-Coll and others, 2016). The DNA lesion 8-OH-dG is formed by the oxidation of the C-8 position of 2′-deoxyguanosine and is induced in the liver after exposure to lead. Formation of the 8-OH-dG lesion has been correlated with elevated levels of ROS in lead-treated rats (Liu and others, 2012). Lipid peroxidation is the process by which ROS attack unsaturated lipids producing lipid radicals (Ayala and others, 2014). Both DNA damage and lipid peroxidation occur under baseline conditions but can be repaired and maintained at low levels until repair capacity or compensatory mechanisms are overwhelmed. Neither lipid peroxidation nor DNA damage were increased in the lead-exposed birds of the present study, suggesting that the increased concentrations of GSH (and possibly other antioxidant species) were able to mitigate any oxidative stress generated by exposure to lead and reduced the amount of detectable tissue damage or reproductive impairment.

Microscopic Lesions in Liver and Kidney Tissues

Our soil and tissue results suggest that the lack of significant microscopic lesions that are generally associated with lead exposure (for example, acid-fast intranuclear inclusions or hepatocellular degeneration lacking an apparent etiology such as parasites) does not rule out the possibility of lead poisoning of birds in the Southeast Missouri Lead Mining District. We did not observe acid-fast intranuclear inclusions, a lesion previously associated with lead poisoning in mourning doves and other species (Kendall and others, 1983), in any of our study birds, despite the elevated lead concentrations in the tissues of the study birds at the contaminated sites (tables 8, 9). However, Beyer and others (2004) were similarly unable to link lesions to lead exposure in the Tri-State Mining District; and in the Southeast Missouri Lead Mining District, acid-fast inclusions were only observed in the two American robins with the greatest concentrations of renal lead (952 and 1,030 mg/kg dw; n=6 American robins of 34 total birds; Beyer and others, 2013). In the current study, the greatest renal lead concentration was 134 mg/kg dw in one mourning dove (table 9; Magmont Vent); the relatively low concentrations of renal lead measured in the current study compared to those measured by Beyer and others (2013) may explain the lack of intranuclear inclusions. It has also been documented that lesions may be less severe in birds with chronic lead poisoning compared to acute lead poisoning (Bono and Braca, 1973).

Moreover, little is known about the microscopic changes associated with lead poisoning in the species assessed in this study; lesion susceptibility may be species or tissue specific. For example, microscopic lesions were uncommon in the kidneys of lead-poisoned eagles; renal tubular necrosis is reported in only 5.4 percent of lead-poisoned Haliaeetus leucocephalus (Linnaeus, 1766) (bald eagles) and 17.6 percent of Aquila chrysaetos (Linnaeus, 1758) (golden eagles; Franson and Russell, 2014). In that study, microscopic lesions from lead poisoning were most common in the heart, a tissue that was not examined as part of this study.

Although inflammation and other lesions were common in the livers and kidneys of all four species examined (table 21), the prevalence of these changes was similar for birds at contaminated and reference sites. Further, the occurrence of lesions in individual birds did not seem to correspond to elevated lead concentrations in tissues; this suggests that the lesions are not likely caused by lead exposure. In northern cardinals and American robins, protozoan or metazoan parasites were present in many areas of inflammation and necrosis; these parasites were the most likely cause of these lesions. Some evidence indicates that birds exposed to lead experience immunosuppression and are more susceptible to parasitic infections (Vallverdú-Coll and others, 2019); moreover, birds experiencing oxidative stress may mount less effective immune responses to parasitic infection (Cram and others, 2015).

Effects of Lead on Reproductive Success

Simple comparisons of descriptive statistics indicated that species-specific means for clutch size, proportion hatched, number of young fledged based on all nests, and number of young fledged by successful nests of the five focal species (eastern bluebird, eastern towhee, field sparrow, indigo bunting, northern cardinal; table 22) were greater for reference sites than contaminated sites in 17 of 20 possible comparisons. However, when reproductive success was directly related to local lead concentration in soil while controlling for site and year in a GLM model, we found mixed support for effects on reproductive success. We found the strongest support for a negative effect of local lead concentration in soil on nest survival of open-cup nesting species. As previously described, our models indicated that period nest survival of eastern towhees and indigo buntings decreased from >20 to <5 percent each as local lead concentration in soil increased from 18 to 4,400 ppm. Period nest survival for northern cardinals was decreased from 19 to 13 percent based on a model with more support for SoilPb than a model without lead effects. However, these estimates had some uncertainty. Additionally, while there was a probability of a negative effect of SoilPb on the period nest survival for eastern towhees (0.97 probability), indigo buntings (0.99 probability), and northern cardinals (0.67 probability), the 95-percent credible interval for the mean effect overlapped zero for eastern towhees and northern cardinals. For field sparrows, there was a 70-percent probability of a positive relation between period nest survival and SoilPb concentrations. Nest success of eastern towhees, indigo buntings, northern cardinals, and field sparrows generally range from 18 to 45 percent in woodlands and field edges in other areas of Missouri with no known lead contamination (Woodward and others, 2001; Burhans and others, 2002; Peak and others, 2004; Roach and others, 2018). Nest success of the four open-cup nesters in this study was within the 18 to 45 percent range at low lead concentration in soil (<100 ppm), but was reduced to 2.0, 4.8, and 14.3 percent for eastern towhees, indigo buntings, and northern cardinals, respectively, when local lead concentrations in soil were ≥4,400 ppm.

The results for other reproductive measures were more equivocal. Overall, the models without lead effects for clutch size, hatching success, and number fledged had more support than models with lead effects. However, the differences in support were small (ΔDIC=1.2–1.6) for eastern bluebird, suggesting that models with lead effects should receive some consideration. Probabilities of a negative effect of local lead concentrations in soil on eastern bluebird reproductive measures were as great as 0.64 on clutch size and 0.78 for hatching success.

Pattee (1984) also did not find a difference in clutch size, clutch initiation date, or egg fertility between lead-dosed (50 ppm metallic lead) and control American kestrels. Similarly, in European starlings, there were no differences in clutch size, hatching success, or number and weight of 18-day-old nestlings between lead-contaminated (56 to 410 mg/kg dw lead in soil; 170 to 1,200 mg/kg dw lead in invertebrates) and control (16 to 34 mg/kg dw lead in soil; 1.7 to 58 mg/kg dw lead in invertebrates) sites (Grue and others, 1986). However, the mean brain weight of starling nestlings from the contaminated site (1.1 g) was significantly lower than the mean brain weight for control nestlings (1.3 g). Decreased brain weight associated with lead exposure has also been documented in the altricial nestlings of American kestrels and pied flycatchers (Hoffman and others, 1985a; Nyholm, 1998). This brain weight decrease suggests that brain development could be a more sensitive indicator of lead exposure than the reproductive parameters measured in this study.

Differences in the nesting ecology between the cavity-nesting eastern bluebird and the four open-cup nesting species (eastern towhee, field sparrow, indigo bunting, northern cardinals) likely affected the observed differences in relations of reproductive success with local lead concentrations. Nest predation is the primary cause of nest failure for songbirds (Ricklefs, 1969; Thompson, 2007; Remeš and others, 2012) and was the cause of most nest failures in this study. Indeed, nest survival was, on average, much lower for open-cup nesters (10 to 20 percent) than for eastern bluebirds (93 percent) even at reference sites. Nest box nests are considerably more protected from predators than open-cup nests, and the predator guards on our nest boxes provided nests with additional protection from predation. Open-cup nesting species were exposed to a similar range of local soil lead concentrations as eastern bluebirds. However, even at the minimum local lead concentration in soil (18 ppm), nest survival success of open-cup nesting species (17 to 23 percent) was much lower than the average nest survival of eastern bluebirds (93 percent), despite normalization with baseline predation levels, which supports the hypothesis that open-cup nesters were subjected to greater overall nest predation.

We believe that the observed disparity between effects on eastern bluebirds and open-cup nesting birds, and the observation that nearly all nest failures in open-cup nesting birds was from predation, is important. We hypothesize that lead exposure may affect adult and nestling behavior in such a way that makes nests more readily detectable by predators or that results in poorer nest defense behaviors. Lead affects the avian nervous system and can cause behavioral changes, including decreased food acquisition and poor coordination (Haig and others, 2014). Although birds have evolved approaches to reduce activity at the nest, which include altering incubation and feeding behavior and synchronization of parental behaviors (Conway and Martin, 2000; Leniowski and Węgrzyn, 2018), greater nest predation could result if lead poisoning impaired the ability of adult birds to perform or coordinate these behaviors. For example, increased parental activity at the nest can result in higher nest predation (Martin and others, 2000). This hypothesis could be investigated by video monitoring at bird nests to document adult and nestling failure and causes of failure (Ellison and Ribic, 2012).

At elevated local lead concentrations in soil (4,400 ppm), period nest survival was only 2 to 14 percent for three of the four open-cup nesting bird species (eastern towhee, indigo bunting, northern cardinal; fig. 17). Although field sparrows showed an inverse relation in which nest success increased slightly as lead concentrations in soil increased, support for this was minimal because the credible interval for this effect overlapped zero, and the effect had only a 0.69 probability of being positive. Donovan and Thompson (2001) used a simple population model to demonstrate that a typical migrant songbird with 60 to 70 percent adult survival would likely exhibit population decline if period nest survival was below 20 to 30 percent. Although other demographic parameters, such as number of nesting attempts and the propensity to double brood, need to be factored into a population model, it is possible that birds occupying areas with elevated local lead concentrations are not producing enough young to balance adult mortality and sustain a local population. Nevertheless, birds may persist at a site despite low reproduction if the local population is supported by immigrants from other areas, creating a population sink. Although sink populations can persist under certain scenarios, they can have consequences for the larger metapopulation and even create regional declines (Pulliam, 1988; Donovan and others, 1995). Eastern bluebird, eastern towhee, field sparrow, indigo bunting, and northern cardinal are all exhibiting population declines of 0.19 to 2.28 percent per year in the Central Hardwood Region (not shown), which encompasses our study sites (Sauer and others, 2017).

Nestling growth and fledgling weight, which affect survival after fledging, may also be sensitive indicators of environmental conditions and are therefore potential alternative measures of reproductive performance (Furness and Greenwood, 1993). Although nestlings in the present study were weighed and measured before fledging, their ages varied or were unknown. However, future analyses could assess body condition as the residuals from a regression of weight on tarsus or wing chord and relate this to lead concentrations. Moreover, reproductive success was based on a sample of birds that started nesting. It is possible that the individual birds that were the most severely affected by lead exposure and uptake died before having the opportunity to breed; alternately, the affected birds did not breed as a direct result of their exposure or left the area without breeding. This may have reduced the frequency with which lead could be observed to affect the reproductive outcomes that we measured. In addition, this study focused on breeding adults and did not account for lead effects on nonbreeding individuals inhabiting the study sites. Additional analyses of cumulative effects, such as annual reproductive success using species-specific information (in contrast to the daily nest survival metric used here), may be warranted.