Utilizing Soil Characteristics, Tissue Residues, Invertebrate Exposures

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Utilizing soil characteristics, tissue residues, invertebrate exposures and invertebrate community analyses to evaluate a lead-contaminated site: A shooting range case study

Dissertation

Presented in Partial Fulfillment of the Requirements for the Degree Doctor of Philosophy In the Graduate School of The Ohio State University

Sarah R. Bowman, M.S. Graduate Program in Evolution, Ecology, and Organismal Biology The Ohio State University 2015

Dissertation Committee: Roman Lanno, Advisor Nicholas Basta Susan Fisher

Abstract With over 4,000 military shooting ranges, and approximately 9,000 non-military shooting ranges within the United States, the Department of Defense and private shooting range owners are challenged with management of these sites. Ammunition used at shooting ranges is comprised mostly of lead (Pb). Shooting ranges result in high soil metal concentrations in small areas and present unique challenges for ecological risk assessment and management. Mean natural background soil Pb is about 32 mg/kg in the eastern United States, but organisms that live in soil or in close association with soil may be at risk from elevated levels of Pb at shooting ranges. Previous shooting range studies on the ecotoxicological impacts of Pb, with few exceptions, used total soil Pb levels as a measure of exposure. However, total soil Pb levels are often not well correlated with Pb toxicity or bioaccumulation. This is a result of differences in Pb bioavailability, or the amount of Pb taken up by an organism that causes a biological response, depending on soil physical/chemical characteristics and species-specific uptake, metabolism, and elimination mechanisms. Therefore, in order to better characterize exposure and risk at these sites, it is necessary to use site-specific measures of bioaccessibility, bioaccumulation, toxicity tests, and characterizations of biotic communities.

The study sites for this research included a private shooting range (1,457–10,044 mg total Pb/kg soil) and an off-site reference area (Gallant Woods, Delaware County,

OH; 6-11 mg total Pb/kg soil) located in central Ohio, USA. The shooting range

consisted of three areas of interest: shotfall (where most Pb shot falls at the site), extracted (where Pb shot was removed by wet sieving in 2009), and on-site reference area.

Elevated soil Pb levels at shooting ranges present a potential risk for wildlife living in the vicinity of ranges, with the direct ingestion of Pb shot and contaminated soil as the main pathways of exposure. Receptors such as earthworms live in the various soil horizons and are exposed to Pb through the soil pore water (dissolved Pb ions) and through direct ingestion and processing of organic matter. Lead can be incorporated into different tissue fractions within earthworms including the metal rich granule (MRG) and the non-metal rich granule (non-MRG) fractions. These fractions are important because the MRG fraction may be used as a Pb sequestration mechanism in earthworms, reducing the internal fraction of bioavailable Pb, and hence, toxicity. Another key component in ecological risk assessment includes the depuration (clearing of the gut) of organisms used for tissue analysis. Earthworms can be up to 30% soil by weight, therefore, it may be important to depurate organisms during bioaccumulation studies. However, during trophic interaction studies, it may be beneficial to use non-depurated earthworms since predators would consume food as-is.

Standard earthworm reproduction and bioaccumulation assays were conducted with soils from the three areas from the shooting range and the off-site reference area.

Earthworm reproduction (number of cocoons and juveniles) was reduced in the shooting range soils. Bioaccumulation varied for each treatment, but was higher for shooting range soils than off-site reference and negative control soils. Shooting range field-collected earthworm tissues had significantly higher Pb (range: 121–1,574 mg/kg) than worms

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collected at the off-site reference area (range: 2–4 mg/kg). Total Pb in depurated earthworms exposed to on-site reference soil and shotfall soil was lower than total Pb in non-depurated earthworms exposed to the same soils. A weak extraction (ammonium nitrate; ISO19730:2008(E)) of Pb also followed a similar trend. The ammonium nitrate extractable Pb is a measure of easily soluble Pb fractions of soil. This measure may be a better estimate of earthworm Pb exposure than total Pb because earthworms are exposed to Pb through pore water which would include the easily extractable Pb. Total earthworm

Pb was highest when compared to the MRG and non-MRG fractions for all lab-exposed and field-collected earthworms, while MRG fractions contained more Pb (for each soil/treatment type) than the non-MRG tissue fractions. The results also indicate that the percentage of non-MRG Pb is smaller for field-collected earthworms than for lab- exposed earthworms.

Ground beetles also live in close association with soil and feed on plant material or other insects. They can be used as biological indicators of ecosystem change since they are relatively easy to collect and identify with proper training. In order to better understand the community-level effects of Pb along a Pb gradient at the shooting range, the abundance and diversity of ground beetle communities were examined.

Ground beetles (Carabidae) were collected using pitfall traps at the shooting range to determine if community structure (i.e., diversity and abundance) changes along a Pb gradient. Twenty-three genera were collected on the shooting range site from 2010–2012 and overall taxa richness and ground beetle abundance was highest in the extracted area.

An indicator species analysis suggested that four taxa were indicators of the shotfall site, six taxa were indicators of the extracted site, and three taxa were indicators of the on-site

reference area. The results suggest that habitat disturbance of soil wet screening and subsequent rye grass slice-seeding in the extracted area was more important in determining ground beetle abundance and diversity than soil Pb concentrations. Total Pb in depurated Harpalus ground beetles from the shooting range sites (5.8–12.4 mg/kg Pb) was higher than depurated Harpalus ground beetles from the off-site reference area (0–

1.3 mg/kg Pb). Average total Pb in depurated beetles from the extracted area (8.7 ± 1.9 mg/kg Pb) was lower than average total Pb in non-depurated beetles from the extracted area (31.4 ± 7.9 mg/kg Pb).

Small mammals also live in close association with soil and depending on the species, consume plant material, insects, or other small mammals. Many small mammals also exhibit burrowing behavior. Small mammals may be exposed to Pb by direct ingestion of soil or contaminated food items. The meadow vole (Microtus pennsylvanicus), is one of the most common small mammals in the northern United

States and is a receptor of interest for this study. Meadow voles feed mostly on monocot and dicot shoots, and to a lesser extent, seeds, roots, fungi, and insects. In mammals, Pb accumulates in the liver, kidney, and bone tissue.

Meadow voles were collected using snap-traps at the shooting range and off-site reference area for tissue Pb analysis. Lead levels in kidney tissue of voles sampled from the shooting range (105 ± 17 mg/kg) exceeded published critical renal Pb values and were significantly higher than Pb levels in kidneys of voles sampled from the off-site reference area (0.34 ± 0.08 mg Pb/kg). Lead levels in liver tissue from voles at the shooting range (17 ± 2 mg/kg) were also significantly higher than samples from the reference site (0.15 ± 0.03 mg/kg). The percent of soil in the diet of meadow voles (4%)

at the site was estimated using element tracers. The Ohio State University in vitro gastrointestinal (OSU-IVG) method, a liquid extraction technique designed to mimic physiological conditions in the human stomach, was used to estimate the bioaccessible pool of Pb that may be dissolved in the stomach. The OSU-IVG extraction value was divided by the total Pb content of a soil to obtain the percent bioaccessible Pb in the soil.

Values for shooting range soils (69–75%) were not different from positive control soil

(83%), but were different than off-site reference (19%) and negative control (19%) soils.

Literature values and field measurements were used to calculate hazard quotients for meadow voles at the site. Even with adjustments for bioaccessibility of Pb in soil, hazard quotients were greater than one.

Based on laboratory and field-based analysis, it appears that receptors at the site

(earthworms, beetles, and small mammals) may be under stress from elevated Pb concentrations in the soil. Future earthworm studies should focus on multi-generational effects of Pb on earthworm reproduction and bioaccumulation (total Pb, MRG Pb, and non-MRG Pb). Ground beetle laboratory toxicity studies should be conducted in the future to determine concentrations in the soil and tissue residues that cause acute and chronic endpoints. Future ground beetle community field studies should be conducted to monitor trends in taxa abundance, richness, and indicator species at other contaminated sites. Before ground beetles can be used as indicators of contaminated or reference sites, more data collection needs to occur. Future mammal studies should focus on pinpointing tissue concentrations that correlate with histological change endpoints. Population-level effects such as adaptation of shooting range mammals to high soil Pb concentrations should also be investigated. Bioaccessibility assays should be developed and validated for

use with ecological receptors. Future studies at the site should consider higher level carnivores (e.g., raptors) to determine if impacts to lower trophic levels affect carnivores that rely on lower trophic levels for food. It is also recommended that similar studies be conducted at other shooting range sites that underwent similar management techniques

(i.e., Pb shot removal) to see if similar trends exist in invertebrate and mammal data.

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Acknowledgements

I would first like to thank my advisor, Dr. Roman Lanno for help with the planning, support, and implementation of this research. I also thank Roman for being supportive of non-academic endeavors which have allowed for my professional development. Thank you to Dr. Timothy Buckley, Dr. Nick Basta, Dr. Susan Fisher, and

Dr. Dave Stetson for their advice and recommendations and serving on my committee.

I also want to acknowledge the constant love and support of my husband, Jeremy

Bowman. I would never have made it this far without his encouragement. I am grateful for his help visiting my field sites and helping to depurate earthworms.

Special thanks to those that provided lab and field support: K. Albanese, A.

Tyrpak, J. Garrett, V. Strickland, G. Graves, A. Lutton, S. Deardurff, N. Hamilton, E.

Wand, S.N. Bowman, S. Nagel, A. M. Siewe, M. Vasko-Bennett, C. Holland, J. Phillips,

B. Little, and J. Bryant.

I would also like to acknowledge research funding and travel support from the following: Ohio State R.H. Edgerley Environmental Toxicology Fund; Ohio State

Department of Evolution, Ecology, and Organismal Biology; the Society of

Environmental Toxicology and Chemistry; and the Ohio State Council of Graduate

Students. Finally, a very special thanks goes out to an anonymous shooting range owner for allowing me to use the range for my research project.

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Vita

2003...... Highland High School

2007...... B.S. Biology, Baldwin-Wallace University

2010...... M.S. Entomology, University of Georgia

2010 to present ...... Graduate Student, Department of Evolution,

Ecology, and Organismal Biology, The Ohio

State University

Publications

Beganyi S.R., & Batzer D.P. 2011. Wildfire induced changes in aquatic invertebrate communities and mercury bioaccumulation in the Okefenokee Swamp. Hydrobiologia 669:237–247.

Bowman, S.R., & J. Bozich. 2015. Internet-based platforms for science communication. Integrated Environmental Assessment and Management 11:516–518.

Hernout B.V., Bowman S.R., Weaver R.J., Jayasinghe C.J., & Boxall A.B. 2014. Implications of in vitro bioaccessibility differences for the assessment of risks of metals to bats. Environmental Toxicology and Chemistry 34:898–906.

Fields of Study Major Field: Evolution, Ecology, and Organismal Biology

Table of Contents

Abstract ...... ii

Acknowledgements ...... viii

Vita ...... ix

List of Tables ...... xi

List of Figures ...... xiii

Chapter 1: Introduction ...... 1

Chapter 2: Evaluating the ecological risk of lead (Pb) to earthworms at a

shooting range field site: Tissue residues, bioaccumulation, and

reproduction...... 6

Chapter 3: Community structure, distribution and abundance of ground

beetles (Carabidae) along a lead gradient at a shooting range field site ...... 42

Chapter 4: Lead tissue residues, bioaccessibility estimates, and estimation of

incidental soil ingestion of the Meadow Vole (Microtus pennsylvanicus)

collected at a shooting range field site ...... 73

Chapter 5: Conclusion...... 99

References ...... 102

List of Tables

Table 2.1 Soil characteristics for the shooting range and off-site reference

area field soils, and laboratory soils ...... 35

Table 2.2 Earthworm reproduction endpoints for field and laboratory soils ...... 36

Table 2.3 Total lead of earthworm MRG and non-MRG fractions for field

and laboratory earthworms ...... 40

Table 3.1 Mean ground beetle abundance (beetles/trap-day) for each

sampling year for shotfall, extracted, and on-site reference areas ...... 65

Table 3.2 Mean ground beetle taxa richness for each sampling year for

shotfall, extracted, and on-site reference areas ...... 66

Table 3.3 Ground beetle indicator species for shotfall, extracted, and on-

site reference sampling areas for 2010, 2011, and 2012 ...... 69

Table 4.1 Mean (± standard error) for body, liver, and kidney weights for

Microtus pennsylvanicus collected at the off-site reference area

(n = 22 individuals) and the shooting range (n = 18 individuals)

field sites...... 93

Table 4.2 Mean (± standard error) total liver and kidney Pb (d.w.) for

Microtus pennsylvanicus collected at the off-site reference area

(n = 22 individuals) and the shooting range (n = 18 individuals)

field sites...... 94

Table 4.3 Soil characteristics for the shooting range and off-site reference

area field soils, and laboratory soils ...... 95

Table 4.4 Total dry weight (µg) range, mean ± standard error, and n for

stomach contents and fecal pellets for shooting range and off-site

reference Microtus pennsylvanicus ...... 96

Table 4.5 Total Pb (mg/kg) range, mean ± standard error, and n for

stomach contents and fecal pellets for shooting range and off-site

reference Microtus pennsylvanicus ...... 97

Table 4.6 Soil equivalent values (total element feces / total element in soil)

for shooting range fecal samples ...... 98

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List of Figures

Figure 1.1 Conceptual model for the shooting range field site that includes:

soil-dwelling invertebrates, plant material, herbivores, and

carnivores ...... 5

Figure 2.1 Bioaccessibility and bioavailability of cationic metal (Mn+) to

soil organisms ...... 33

Figure 2.2 Aerial photograph of the shooting range (showing shotfall,

extracted, and on-site reference areas) and off-site reference area ...... 34

Figure 2.3 Total lead bioaccumulation (mg/kg, dw) bioaccumulation in

Eisenia fetida exposed to soil treatments and sampled after 0, 7, 14,

28, or 63 days of exposure ...... 37

Figure 2.4 Lab and field earthworm total lead ...... 38

Figure 2.5 Total Pb (mg/kg, dw) in metal-rich granule (MRG) and non-

MRG fraction of laboratory-exposed or field-collected

earthworms ...... 39

Figure 2.6 Total earthworm lead for depurated and non-depurated

laboratory exposed earthworms...... 41

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Figure 3.1: Example pitfall trap with deli cup (not visible) that sits flush

with ground level, four landscape-edging lead-ins, and rain

cover...... 61

Figure 3.2 Ground beetle taxa accumulation curve (cumulative number of

samples vs. cumulative number of taxa) for the duration of the

study ...... 62

Figure 3.3 Mean ground beetle abundance ± SE (beetles/trap-day) for

shotfall (triangle), extracted (circle) and on-site reference

(square) over the duration of the 2010 (a), 2011 (b), and 2012 (c)

sampling seasons ...... 63

Figure 3.4 Mean ground beetle richness ± SE (richness/trap-day) for

shotfall (triangle), extracted (circle) and on-site reference

(square) over the duration of the 2010 (a), 2011 (b), and 2012 (c)

sampling seasons ...... 64

Figure 3.5 Overall non-metric multidimensional scaling (NMDS)

ordination of shotfall (triangle), extracted (circle) and on-site

reference (square) Carabidae beetle data ...... 67

Figure 3.6 NMDS ordination of shotfall (triangle), extracted (circle) and

on-site reference (square) for 2010 Carabidae beetle data ...... 68

Figure 3.7 NMDS ordination of shotfall (triangle), extracted (circle) and

on-site reference (square) for 2011 Carabidae beetle data ...... 70

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Figure 3.8 NMDS ordination of shotfall (triangle), extracted (circle) and

on-site reference (square) for 2012 Carabidae beetle data ...... 71

Figure 3.9 Mean ± SE total Pb for depurated (shaded bars) and non-

depurated (open bars) field-collected Harpalus beetles ...... 72

Chapter 1: Introduction

Lead (Pb) is a naturally occurring element, and is a common soil contaminant due to historical and present anthropogenic sources including paints, gasoline, batteries, and ammunition. Although Pb shot was banned for waterfowl hunting in 1991 (U.S. 50

CFR § 20.108), Pb shot is still in use in many states for upland game species and at shooting ranges. Shooting ranges present unique challenges for ecological risk assessment because Pb shot can be heterogeneously distributed (in high concentrations) over a relatively small area.

Natural soil background Pb is around 32 mg/kg in the eastern United States

(USEPA 2005b). At shooting ranges, soil Pb concentrations can reach values greater than 90,000 mg/kg (in backstop berms; Astrup et al. 1999) and greater than 1,000 mg/kg at trap and skeet ranges (Murray et al. 1997). Many shooting ranges exceed Pb ecological soil screening levels (EcoSSLs) for plants (120 mg/kg), soil invertebrates

(1,700 mg/kg), birds (11 mg/kg), and herbivorous mammals (56 mg/kg) (USEPA

2005b). In these cases, it may be necessary to assess additional parameters to determine if receptors are at risk from Pb at the site.

As a result of the deposition of large amounts of Pb in soil and measured high levels of Pb, there is an increased risk associated with Pb exposure for ecological receptors associated with shooting ranges. Figure 1.1 provides a conceptual model of ecological receptors present at shooting ranges and Pb exposure pathways. The toxic risk

posed to birds by the ingestion of Pb shot is well documented (see reviews by Franson &

Pain 2011, and Scheuhammer & Norris 1996) and will not be investigated in this dissertation. Receptors of interest for this study that have received much less ecotoxicological attention include soil-dwelling invertebrates (Lumbricidae earthworms and Carabidae ground beetles) and small mammals (meadow voles, Microtus pennslyvanicus). In describing the conceptual model, the Pb source term (shotgun pellets) has been deposited on the soil surface and through various processes, has been incorporated into the soil compartment. Soil-dwelling invertebrates such as earthworms and ground beetles live in close association with the soil. Earthworms process organic matter from soil while ground beetles feed on plant material or other invertebrates.

Herbivorous small mammals (such as meadow voles) also live in close association with the soil and exhibit burrowing behavior. They mostly feed on plant material (shoots of monocots and dicots) but will occasionally feed on invertebrates. All organisms may incidentally ingest soil during normal daily habits (e.g., grooming and burrowing) and through their diet (as dust on the surface of diet items or soil in the digestive tract of diet items). Lower trophic levels are food sources for higher trophic level carnivores such as the robin, shrew, and hawk.

The source term for Pb at the site provides an interesting conundrum as to what the best measure of exposure might be for the different organisms in the conceptual model. Although total Pb levels have been used previously as measures of exposure, especially in laboratory toxicity studies where Pb salts are added to soil, it is clear that one measure of Pb exposure will not suffice for all organisms. Shot pellets are deposited as elemental Pb, but then transform very slowly over time to various Pb species, each

with different bioavailability. Bioavailability or “the amount of chemical that is actually taken up from the environment and is available to cause a biological response” (NRC

2003) is modified by physical, chemical, and biological interactions among the organism, the chemical, and its environment (NRC 2003). Therefore, instead of measurements of total chemical in soil, it may be beneficial to measure organism tissue residues and compare these to tissue values that are related to adverse effects in that species. However, these types of studies may not always be practical (e.g., endangered species), can be expensive, and require much sampling effort to capture field variability.

In combination with tissue residue approaches, it can be useful to measure the bioaccessibility of a chemical which is defined as the pool of a chemical which is available to organisms for uptake. For small mammals with ingestion as the main pathway of exposure, this is often the amount of chemical that is dissolved in the stomach and available for uptake in the small intestine. The Ohio State University in vitro gastrointestinal (OSU-IVG) method, developed for use in human risk assessment, if validated for ecological receptors, can be used to estimate Pb that is bioaccessible to small mammals. In addition, weak soil extractions (e.g., ammonium nitrate; ISO

19730:2008(E)) can help assess easily soluble trace element fractions in soil and may approximate the bioaccessible fraction of Pb in soil for dermal exposure of earthworms.

There are many ways to incorporate ecological receptor exposure into an ecological risk assessment. This study will examine earthworms, ground beetles, and small mammals in more detail to understand how these receptors may be impacted by soil Pb at the site, especially by examining various measures of Pb exposure. This will be accomplished by conducting laboratory and field-based studies of earthworm

reproduction and bioaccumulation, earthworm tissue residues, ground beetle community composition and tissue residues, and small mammal tissue residues and incidental soil ingestion analysis. The results of this study will hopefully provide more detail about additional management steps at this shooting range as well as inform similar risk assessments in the future.

The current study site is a private shooting range located in central Ohio, USA.

The range is currently used as a trap and skeet range and has been in operation for over

20 years. Areas of interest for this study include a shotfall area, extracted area (where lead shot was removed by wet screening in fall 2009), and an on-site reference area

(more details provided in Chapter 2). Bryant (2010) found that soils at the site exceeded

EcoSSLs for Pb, but not for arsenic or antimony (other common elements in bullets). An off-site reference area (Gallant Woods Preservation Park, Delaware County, Ohio) was also chosen for comparison of Pb tissue values. The following chapters will describe laboratory and field-based assessments of the three receptors at the site, including earthworms (Chapter2), ground beetles (Chapter 3), and meadow voles (Chapter 4).

Chapter 5, the conclusion, will summarize the findings and suggest future steps in shooting range Pb ecotoxicology.

Figure 1.1: Conceptual model for the shooting range field site that includes: soil- dwelling invertebrates, plant material, herbivores, and carnivores.

Chapter 2: Evaluating the ecological risk of lead (Pb) to earthworms at a shooting range

field site: Tissue residues, bioaccumulation, and reproduction

Introduction

Although lead (Pb) is a naturally occurring element, it is a common soil contaminant in many urban areas where Pb is concentrated in the soil due to industrial processes and from the historical use of Pb in gasoline and paints. Lead is also of concern at shooting ranges where Pb shot may be a significant source of Pb to the soil. There are approximately 9,000 non-military shooting ranges and 4,000 Department of Defense

(DoD) military shooting ranges within the United States (USEPA 2005a). Bullets and other ammunition account for approximately 4% of the yearly Pb use in the United States

(USEPA 2005a). It is estimated that more than three million metric tons of Pb was deposited in the United States in the form of bullets and ammunition during the last century (Craig et al. 1999). While some Pb ammunition use is widespread for hunting, most is used for recreational shooting at shooting ranges. In 1991, the United States Fish and Wildlife Service banned the use of Pb shot for hunting waterfowl (U.S. 50 CFR §

20.108) and additional state laws have limited the use of Pb shot for certain types of game. However, Pb shot is still used at most of the 13,000 shooting ranges in the United

States (USEPA 2005a).

Natural background Pb concentrations average 32 mg/kg in the eastern United

States and 38 mg/kg in the western United States (USEPA 2005b). At shooting ranges,

Pb concentrations can vary from background concentrations to values greater than 90,000 mg/kg in backstops and berms (Astrup et al. 1999). Lead concentrations in the soil may vary with the type of shooting range (e.g., trap and skeet or rifle and pistol), background soil characteristics, composition of bullets, and frequency of use (Pb loading). The distribution of Pb at shooting ranges depends largely on the type of facility. Rifle and pistol ranges tend to have small areas with high concentrations (> 90,000 mg/kg Pb) in a backstop or berm (Astrup et al. 1999). At trap and skeet ranges, Pb pellets are spread over a larger area and may be at lower concentrations (~1,000 mg/kg Pb) than at rifle and pistol ranges (Murray et al. 1997). Regardless of the type of range, all of these areas tend to have Pb deposits above natural background levels.

These elevated levels of Pb in the soil at shooting ranges present a potential risk for wildlife living in the vicinity of ranges, with the direct ingestion of Pb shot and contaminated soil as the main pathways of exposure. Earthworms that live in various soil horizons and process soil through their digestive systems (Reynolds 1977) can accumulate metals in tissues, potentially resulting in toxicity to the earthworm or predators that may ingest the worm. Overwhelmingly, almost all toxicity tests that have assessed the toxicity of Pb to soil invertebrates have been conducted in soils spiked with

Pb salts. These types of tests have been useful for determining the relationship between

Pb concentrations in soil, toxicity, and how toxicity is modified by the physical and chemical characteristics of soils such as pH, cation-exchange capacity (CEC), and organic carbon content (Lanno et al., 2004; Bradham et al., 2006). However, Pb at

shooting ranges is not found in the form of Pb salts added to soil, so the actual source term of Pb exposure must be considered as well. Lead pellets in the soil at shooting ranges go through a transformation from elemental Pb to a form that may be described as a white or grey crust material that is composed mostly of hydrocerussite, cerussite, and angelsite, with complete transformation estimated to take 100 to 300 years (Jørgensen &

Willems 1987). Jørgensen & Willems (1987) found that transformation depends on the size of pellets and site-specific management techniques such as tillage which accelerates transformation. Oxidation of Pb to Pb2+, precipitation, and soil components that retain lead can also affect Pb transformation rates. This shows that in certain circumstances, such as the ecological risk assessment of Pb, it may be useful to understand the speciation and transformation of Pb at a contaminated site, or at the very least, conduct toxicity assessments with site soils rather than relying on data derived from toxicity tests with Pb salts.

Soil modifying factors such as pH, cation exchange capacity (CEC), organic carbon, and clay type and content affect the amount of Pb that is in soil solution as the divalent cation (Pb2+), which is thought to be the toxic form of Pb in soil. As a result, while total soil metal concentrations may provide an idea of how much metal has been added to the soil, total metal concentration does not always correlate with toxicity or bioaccumulation. Bioavailability or “the amount of chemical that is actually taken up from the environment and is available to cause a biological response” (NRC 2003) is modified by physical, chemical, and biological interactions between the organism and its environment (NRC 2003). McLaughlin and Lanno (2014) provide a useful example of bioavailability as the interaction of chemicals with a biological membrane (Figure 2.1).

According to McLaughlin and Lanno (2014), the best measures of bioavailability are actual measures of chemicals in organisms or the binding of chemicals to a membrane

(e.g., fish gill epithelium). However, direct chemical residue measurements are not always practical (e.g., in the case of endangered species), can be expensive, and require much sampling effort to capture differences in field variability. Therefore, it is sometimes useful to measure bioaccessibility which is defined as the pool of a chemical which is available to organisms for uptake (Figure 2.1). Uptake or absorption can occur via external dermal surfaces, from ingestion and absorption across the gut epithelium, or absorption through inhalation pathways in the lungs. After absorption, the chemical may be distributed, metabolized, or excreted by the organism. Some of the chemical may bioaccumulate in tissues and some may end up at the site of toxic action, where it may cause adverse effects or death (Figure 2.1). Soil characteristics play a major role in modifying the amount of bioaccessible chemical in the soil solution. The amount of bioaccessible chemical is important for organisms such as earthworms that are exposed to chemicals through pore water in the soil. The ISO 19730:2008(E) ammonium nitrate extraction of trace elements from soil can help assess easily soluble trace element fractions of soil. Therefore, the amount of ammonium nitrate extractable metal, when correlated to adverse effects in earthworms, may prove more useful for assessment metal exposure at field sites than total Pb concentrations.

Further complicating the issue of bioavailabilty are Pb sequestration mechanisms within organisms. Jones et al. (2009) and Vijver et al. (2006) documented metal sequestration within different tissue fractions of earthworms. Within earthworms, metals can be detoxified by two mechanisms: (1) sequestration of metals in calcium phosphate-

based metal-rich granules (MRGs) within chloragosomes and (2) binding of metals to metallothionein proteins. Vijver et al. (2006) demonstrated that in earthworms, most of the Pb was incorporated into MRGs. Jones et al. (2009) referred to the MRG fraction as the detoxified fraction because they hypothesized that Pb stored in this form is not toxicologically bioavailable to the earthworm, while the toxic fraction includes cellular debris, proteins, and organelles (non-MRG fraction).

Another important consideration in assessing metal bioavailability during bioaccumulation is the depuration (allowing gut contents to clear) of organisms before metals analysis. Standard bioaccumulation tests (e.g., Environment Canada 2007) require depuration in order to determine actual tissue metal concentrations (without including soil or metals that are present in the stomach and digestive tract). However, in ecological risk assessment, depuration may not always be the preferred approach. For example, when estimating exposure of a higher trophic level organism to prey items, it would be beneficial to use non-depurated prey items for analysis. Predators, scavengers, and higher trophic level organisms consume food as-is, without depuration and therefore, the metals and soil within the digestive tract of prey items should also be measured. Beyer et al.,

(1994) explained that approximately 20-30% of earthworm weight is soil. Therefore, organisms that feed upon earthworms will ingest a significant amount of soil through their diet. The decision to use depurated or non-depurated prey items will depend on the goals and design of the study, but it is an important consideration.

In a study of a Canadian prairie skeet range, Booth et al. (2003) found that

Eisenia fetida earthworms exposed to field soils (17.5–292 mg/kg total Pb) for 28 days accumulated between 0.04 and 3.5 mg/kg total Pb in their tissues (after depuration). No

mortality occurred in any of the soils and there were no effects on reproductive endpoints

(Booth et al. 2003). Total soil Pb in this study was very low compared to other shooting range soils and therefore it is not surprising that no mortality or effects on reproductive endpoints were observed. Kaufman et al. (2007) found that field-collected Aporrectodea spp. earthworms (after depuration and rinsing) accumulated on average 730 mg/kg tissue

Pb and up to 3,000 mg/kg (maximum) for bullet stop areas at another Canadian shooting range. Total soil Pb averaged 5,000 mg/kg and the bullet stop contained up to 27,000 mg/kg total soil Pb.

The U.S. Environmental Protection Agency (USEPA) issued shooting range best management practices for shooting range owners to manage Pb contamination at their ranges and reduce the potential of lead exposure to humans and wildlife (2005a). One recommendation is for range owners to remove Pb shot from soil for recycling.

Suggested methods for Pb removal include: hand raking and sifting, screening, vacuuming, and soil washing (wet screening, gravity separation, pneumatic separation)

(USEPA 2005a). Although it is suggested that removal techniques reduce exposure for humans and wildlife, to our knowledge no formal study has occurred to document this.

The study site for this research was a private trap and skeet shooting range located in central Ohio, USA (Figure 2.2) and consisted of a shotfall area, extracted area (where lead shot was removed by wet screening in fall 2009), and an on-site reference area. An off-site reference area was also chosen at a nearby park (Gallant Woods Preservation

Park, Delaware County, Ohio). The overall objective of this study was to integrate multiple laboratory and field approaches to examine Pb exposure to earthworms at a shooting range. Additionally, land management practices at the range provided an

opportunity to understand how Pb removal may impact field-collected soil and earthworm Pb concentrations as well as laboratory bioaccumulation and reproductive endpoints. It was hypothesized that lead extraction from soil for the purposes of recycling spent Pb shot should reduce total Pb concentrations in the soil, which would lead to reduced toxicity and bioaccumulation of Pb. Specifically, the following hypotheses will be tested:

1) If Pb removal by wet screening does reduce total Pb in soil, then soils and

earthworms collected from the extracted area should have lower total Pb

concentrations than the shotfall area. As well, reproduction endpoints in

toxicity tests with earthworms (i.e., numbers of cocoons and juveniles), would

increase in worms exposed to extracted soils relative to those exposed to

shotfall area soil. If Pb removal by wet screening reduces total Pb in soil to

reference conditions, then the extracted area should have similar soil and

earthworm total Pb concentrations as the off-site reference area and negative

control soils, and reproduction endpoints would be similar to those observed

in reference soils as well.

2) When comparing earthworms exposed to Pb in the laboratory and at

contaminated field sites, certain assumptions must be made. If MRG

formation is similar in all genera of earthworms, then laboratory-exposed

Eisenia fetida and field-collected Lumbricus spp. should express similar

metabolism of Pb and have higher total Pb concentrations in MRG fractions

than in non-MRG fractions.

3) For the purposes of comparing potential Pb exposure via bioaccumulation or

to higher trophic levels, it is necessary to compare depurated and undepurated

earthworms. If depuration voids soil from the intestines of earthworms, then it

would be expected that depurated earthworms should have lower total Pb

concentrations than non-depurated earthworms.

Methods

Study site

The study site was a private trap and skeet shooting range that has been in operation for greater than 20 years. During this study, the range was active six days per week. The three main shooting range areas sampled during this study included: 1)

Shotfall area – an area parallel to the shooting stations, about 40-100 m out into the shooting range where most of the Pb shot is deposited; 2) Extracted area: a band parallel to the shooting stations within the shotfall area, approximately 70-100 m out into the shooting range where Pb was removed by wet screening in 2009; and, 3) On-site reference area that lies outside the main shotfall area (Figure 2.2). Sampling areas were selected to be far enough away from the shooting stations to minimize the presence of fragments of sporting clays present on the surface of the soil. The extracted area was not initially present in the shotfall area but was added to the study after the unannounced extraction occurred. Vegetation at the shooting range consisted of grasses and native prairie plant species and was mowed occasionally to a height of approximately 7–10 cm, with all three areas mowed on the same day. The off-site reference area was chosen at a nearby park in Delaware County, Ohio (Gallant Woods Preservation Park) (Figure 2.2).

The park is approximately 231 acres and includes restored wetlands, forest, and prairies.

Before the land was donated to the park district in 2005, the land was privately owned and used for agriculture. An old field was chosen as the off-site reference area for this study because the vegetation consisted of grasses and native prairie plant species, and was also mowed occasionally, providing habitat and soil type similar to the shooting range.

Soil characteristics and extractions

In August 2010, 12 soil cores were collected using a standard golf hole-cutter

(diameter = 10.8 cm, depth = 15 cm) along representative transects in each of three areas at the shooting range site (on-site reference, shotfall, and extracted areas). Twelve soil cores were also collected from the off-site reference area. Each soil sample was a composite of four cores taken around a central point. In the lab, rocks, visible shot fragments, and debris were removed from the soil. Soils were then dried, ground, sieved to less than 2 mm (stainless steel mesh), and homogenized for each composite sample.

Webster soil previously collected from Iowa, USA (sieved and homogenized as stated above) was used for the laboratory controls. Soil samples were sent to the Soil, Plant and

Water Laboratory at the University of Georgia for the determination of pH (1:1 soil to

0.01 M CaCl2 suspension), organic matter (loss on ignition for 3 hours at 360°C), extractable salts (Mehlich 1 extraction and ICP Spectrograph), and percent sand, silt, and clay (Bouyoucos hydrometer method). Cation exchange capacity (CEC) was determined by adding the milliequivalents of calcum, magnesium, potassium, and sodium present in the Mehlich-1 extract and the milliequivalents of exchangeable hydrogen as determined

by direct titration with 0.023M Ca(OH)2). Total carbon was determined by combustion of the sample in an oxygen atmosphere at 1350°C, then measuring the carbon dioxide gas in an infrared cell. For more detailed soil analysis methods, see Kissel and Sonon (2008).

Amorphous iron and aluminum content in soil was determined by acid ammonium oxalate extraction (McKeague and Day 1966) by the Soil Chemistry Lab at The Ohio

State University.

Total Pb concentrations in soil samples were determined in the Soil Chemistry

Lab at The Ohio State University according to method SW-846 3051a, a microwave assisted aqua regia digestion method (USEPA 2007a). Lead analysis was conducted by inductively coupled plasma atomic emission spectrometry (ICP-AES) using an Agilent

720 ICP-AES (Santa Clara, California) according to USEPA method 6010C (USEPA

2007b). The accuracy of ICP analysis was assessed by measuring Pb concentrations in a certified reference materials (CRM) for soil (Metals in Soil Sigma-Aldrich SQC001-

50G). CRM soil samples were digested and analyzed in the same run with the samples to determine percent recovery of Pb in soil. The Pb content of blanks and check standards was measured every 10 samples. Recovery of Pb in CRM059-050 by USEPA method

3051a was 97% of the certified value.

As a measure of the bioaccessible pool of Pb to which earthworms would be exposed to in the soil pore water, ammonium nitrate (NH4NO3)-extractable Pb was determined using the ISO 19730:2008(E) method. Ammonium nitrate extractable Pb was converted to percent extractable using the following equation:

퐼퐶푃 퐸푥푡푟푎푐푡푎푏푙푒 % 푒푥푡푟푎푐푡푎푏푙푒 푃푏 = ∗ 100 [2.1] 푇표푡푎푙 푃푏

Test organisms

Earthworms (Eisenia fetida) from an in-house culture were used for the reproduction and bioaccumulation tests. The continuous earthworm culture has existed at

Ohio State for greater than five years and was occasionally supplemented with new earthworms (E. fetida) from Carolina Biological Supply Company (Burlington, North

Carolina, USA). Earthworms were maintained in bedding of a mixture of horse manure

(75% dw), shredded newsprint (12.5% dw), and peat moss (12.5% dw) and were fed composted horse manure which was occasionally supplemented with vegetable scraps and rolled oats.

Test chambers

Toxicity and bioaccumulation tests were conducted in 500-ml glass mason jars with screw top lids and a single perforation in the lid for gas exchange. Each test chamber contained 250 g dry weight (dw) of soil that was rehydrated with Milli-Q water until the soil stuck together when squeezed. Soil was determined as too wet if water exuded when squeezed. As a positive control, Webster soil was spiked with 2,000 mg/kg Pb (Fisher

Brand crystalline Pb nitrate). The spiked soil was air dried and rehydrated twice before rehydration for the test. Non-spiked Webster soil served as a negative laboratory control.

Earthworms were introduced to test chambers 24 hours after rehydration of the soil.

Reproduction test

Ten adult (fully-clitellate) earthworms with fresh weights of 260–600 mg (mean =

431 mg ± 94 mg SD) were placed on the soil surface in each test chamber (n=4 for each

treatment). Earthworms were fed composted horse manure (25 g dw) once per week and uneaten food was removed weekly. Total test chamber weights were measured at this point and test chambers were rehydrated weekly by spraying water on the surface of the soil if the total test chamber weight was less than 95% of the original weight.

Earthworms were kept under constant temperature (20±2°C) and constant fluorescent lighting. After 28 days, adult earthworms were removed from the chambers, rinsed, and weighed. Worms were depurated on moist filter paper in petri plates for 24 hours

(hereinafter referred to as depuration). Depurated worms were rinsed with Milli-Q water, and frozen (-70°C) for metals analysis. Soils were hand-sorted under a dissecting microscope and the cocoons were counted. Soils and cocoons were placed back into the test chambers, and after an additional 28 days of incubation, soils were hand-sorted under a dissecting microscope and the juveniles were counted.

Bioaccumulation test

Ten adult (fully-clitellate) earthworms with fresh weights of 200–580 mg (mean =

314 mg ± 95 mg SD) were placed on the soil surface in each test chamber (n=4 for each treatment). Earthworms were fed composted horse manure (25 g dw) once per week.

Total test chamber weights were measured at this point and test chambers were rehydrated weekly by spraying water on the surface of the soil if the total test chamber weight was less than 95% of the original weight. Two earthworms per replicate were removed on days 7, 14, 28, and 63. Earthworms were rinsed, weighed, depurated, rinsed with Milli-Q water, and frozen (-70°C) for metals analysis. Day-0 worms were depurated for 24 hours, rinsed with Milli-Q water and frozen at the start of the test.

Field-collected earthworms and tissue fractions

Earthworms from the field sites were collected as they were available from underneath and next to pieces of plastic landscape edging that were stuck in the ground as part of insect pitfall traps placed in the field for sampling as described in Chapter 3.

Earthworms were only observed and collected at the site after heavy rain when standing water was present and the soil was saturated. After several attempts to extract earthworms from the soil with a mustard solution (40 g mustard to 4 L water), it was determined that collecting them after heavy rain was most successful. Earthworms were collected on 3 October 2013 for the off-site reference and 11 April 2013 for shooting range sites and identified using standard taxonomic keys (Reynolds 1977; Hale 2007).

Earthworms were rinsed, weighed, depurated, rinsed with Milli-Q water, and frozen (-

70°C) for metals analysis. Half of all field-collected earthworms were used for tissue fractionation studies as described below.

In an additional bioaccumulation test, E. fetida were exposed to the suite of field soils collected from the shooting range for 28 days following the experimental design described previously for bioaccumulation tests. Following exposure to site soils, earthworms were depurated for 24 hours, rinsed in Milli-Q water, and frozen until fractionation. These worms were also subjected to the fractionation procedure described below.

In order to compare the partitioning of Pb within the earthworms during bioaccumulation in the lab and field, earthworms were separated into two fractions, one containing metal-rich granules (MRGs) and the other containing all other tissues, according to established fractionation and digestion procedures (Vijver et al. 2006). Since

the major route of Pb detoxification in earthworms is incorporation into MRGs, fractionation of earthworm carcasses was only conducted to the first step described in

Vijver et al. (2006) and the other fractions were not separated and analyzed for this study.

Earthworms were homogenized (OmniTH 30311 tissue homogenizer, 7-mm saw-tooth blade) in Tris HCl buffer (pH 7.0, Fisher Scientific) for three minutes (1 minute full speed, 2 minutes at half speed), and then centrifuged at 10,0000g. The pellet was resuspended in Milli-Q water and heated to 100°C for two minutes, NaOH (1 M) was added and heated for one hour. The sample was then centrifuged at 5,000g for 10 minutes. The pellet of the second centrifugation step contained the MRGs. The pellet was resuspended in Tris HCl buffer and dried overnight at 105°C, weighed, and resolubilized in trace metal grade HNO3 prior to Pb analysis. The supernatants from each centrifugation step were combined to form the non-MRG fraction. This fraction was also dried at 105°C overnight, weighed, and resolubilized in trace metal grade HNO3 for Pb analysis. Lead analysis of the two fractions was conducted by ICP-AES as described below.

Depurated vs non-depurated earthworms

In order to determine the effect of depuration on the Pb content of earthworms, E. fetida were exposed to the suite of soils collected at the field site for 28 days following experimental design described for bioaccumulation tests. After exposure, half of the earthworms were depurated for 24 hours, rinsed in Milli-Q water and frozen until analysis. The remainder of the earthworms were not depurated, but rinsed with Milli-Q water, and frozen until digestion and analysis for Pb.

Earthworm total Pb analysis

Earthworm samples were digested and analyzed by the Soil Lab in the School of

Environment and Natural Resources at Ohio State University. In preparation for analysis of earthworms for Pb, carcasses were digested in 10 ml concentrated nitric acid overnight, then samples were heated to 125°C for 4 hours, cooled to room temperature, diluted to 12.5 ml with concentrated nitric acid, and then diluted to a final volume of 50 ml with Milli-Q water (Havlin and Soltanpour 1980). Lead analysis was conducted by

ICP-AES using an Agilent 720 ICP-AES (Santa Clara, California) according to USEPA method 6010C (USEPA 2007b). The accuracy of ICP analysis was assessed by measuring Pb concentrations in a CRM (TORT-2 Lobster Hepatopancreas, National

Research Council of Canada). CRM samples were digested and analyzed in the same run with the earthworm samples to determine percent recovery of Pb in tissue samples.

Recovery of Pb in Tort-2 was within 10% of the certified value. The Pb content of blanks and check standards was measured every 10 samples. Lead concentrations in the earthworms were calculated assuming 84% water content (USEPA 1993).

Statistical Analyses

R version 3.1.0 (R Foundation for Statistical Computing 2014) was used for statistical analyses. To determine differences in soil characteristics and reproductive endpoints (number of cocoons, number of juveniles, number of juveniles per adult), a one-way ANOVA with treatment as a factor was used. A Tukey’s post-hoc test for multiple comparisons was used to test for significant differences among treatments. Total

Pb values were log10 transformed prior to analysis and percent ammonium nitrate

extractable Pb was arcsine-square-root transformed in order to normalize data to meet the assumptions of the ANOVA. The level of significance for all statistical analyses was α =

0.05.

Results

Soil characteristics & extractions

Total Pb levels were highest in soils from the on-site reference, extracted, shotfall, and positive control soils and these samples were significantly higher compared to the off-site reference soil and negative control soil (p = 1.27 x 10-6; Table 2.1). Coefficients of variation for total soil Pb were highest for the on-site reference, shotfall, and extracted soils (Table 2.1). Percent ammonium nitrate extractable Pb was highest in extracted and positive control soils (p = 2.5 x 10-3; Table 2.1) and significantly different from the off- site reference and negative control. Since ammonium nitrate extractable Pb was below detection limits for the off-site reference and negative control samples, these values are presented as zero after subtracting the blank value.

The pH of negative control soil was not different from the shooting range soils or the positive control soil, while off-site reference soil pH was highest when compared to negative control, extracted, and positive control soils (p = 6.5 x 10-4). The extracted and shotfall soils had significantly lowest CEC than the other soils (p = 1.8 x 10-6; Table 2.1).

Although the overall ANOVA model was significant for Al Oxide and Fe Oxide, no clear trend was observed among the soils (Table 2.1). Soil from the on-site reference area had the highest organic matter content when compared to other, soils except the off-site reference soil (p = 1.7 x 10-3; Table 2.1). All soils were less than 6% total carbon and

although the overall ANOVA was significant, no clear trend was observed among the soils (p = 1.5 x 10-2; Table 2.1). The ANOVA models were not significant for sand (p =

8.0 x 10-2), silt (p = 8.0 x 10-2), or clay (p = 1.0 x 10-1) (Table 2.1).

Reproduction test

Earthworm survival was 100% in all treatments at day 28 except for one replicate of shotfall soil where survival was 90%. Reproduction rates in both the off-site reference soil and the negative control soil were greater than three juveniles per adult earthworm.

Although the results are not shown, the mean weight of individual juveniles in the off-site reference and negative control soils was less than or equal to 2 mg at the end of the test.

Therefore, the study met all three criteria for a valid test (Environment Canada 2007).

The average number of cocoons was greatest for the off-site reference and negative control soils (Table 2.2), while the on-site reference area and extracted area were intermediate, with the lowest number of cocoons produced in shotfall and positive control soils (p = 1.6 x 10-5). The average number of juveniles produced (p = 3.5 x 10-9) and ratio of juveniles/adult (p = 6.9 x 10-9) were highest for the off-site reference soil, followed by negative control soil, then the on-site reference, extracted, shotfall, and positive control soils.

Bioaccumulation test

Uptake kinetics of Pb from soil were not consistent among treatments. A number of different models (linear, logarithmic, quadratic) were fitted to the data and the model that maximized the fit (greatest R2) was chosen for each soil (Figure 2.3). The linear

model was not the best descriptor of Pb uptake kinetics in any of the treatments.

Logarithmic and quadratic models provided the best fits for the data. Earthworms exposed to the shotfall soil accumulated the most Pb (over 700 mg/kg for one sample), but bioaccumulation factors are not presented due to the differences in the kinetics of Pb uptake among treatments.

Field-collected earthworms and tissue fractions

Total Pb levels in field-collected earthworms (Lumbricus spp. juveniles) were higher for shotfall and extracted soils than for on-site reference and off-site reference soils (p = 2.1 x 10-4; Figure 2.4). Total Pb levels in shotfall-collected earthworms were not significantly different from Pb levels in earthworms collected from the extracted site

(p = 9.0 x 10-4; Figure 2.4).

Total Pb levels in field-collected extracted and shotfall earthworms were significantly higher than in lab-exposed extracted (p = 1.2 x 10-2) and shotfall (p = 1.9 x

10-3) earthworms (Figure 2.4). There was no difference between lab-exposed and field- collected earthworm Pb concentrations for the on-site reference area (T-test, p = 7.4 x 10-1; Figure 2.4).

Total earthworm Pb was higher than the MRG and non-MRG fractions for all lab- exposed and field-collected earthworms, while MRG fractions contained more Pb (for each soil/treatment type) than the non-MRG tissue fractions (Figure 2.5). The field- collected on-site reference MRG percentage was higher than the lab-exposed MRG percentage (Table 2.3). The results also indicate that the percentage of non-MRG Pb is smaller for field-collected earthworms than for lab-exposed earthworms (Table 2.3).

Depurated vs. Non-depurated Earthworms

Total Pb levels (mg/kg, wet weight) were higher in non-depurated earthworms exposed to on-site reference (T-test; p = 2.0 x 10-2) and shotfall soils (T-test; p = 3.0 x 10-

2) (Figure 2.6). There was no statistical difference between depurated and non-depurated worms for the extracted (T-test; p = 4.9 x 10-1) and negative control soils (T-test; p = 5.0 x 10-2).

Discussion

Removal of Pb from the shotfall area soil in situ via a wet extraction procedure did not result in a significant reduction in toxicity or Pb bioaccumulation by earthworms.

While earthworms exposed to the shotfall soil produced the fewest cocoons, juveniles, and juveniles/adult, these were not significantly different from reproductive endpoints of earthworms exposed to extracted, positive control, and on-site reference soils.

Earthworms exposed to the off-site reference and negative control soils produced the highest number of cocoons and juveniles. The original hypothesis that earthworms exposed to soil from the extracted area would have significantly more juveniles and cocoons than earthworms exposed to the shotfall soil was not supported. The extracted, shotfall, and on-site reference soils all had elevated, but not significantly different, total

Pb concentrations, supporting the observation that reproductive output of earthworms in these soils would be similar. This suggests that wet screening extraction of Pb by the shooting range manager did not return the extracted area soil to reference soil conditions.

The concentration of Pb in ammonium nitrate extractions matched closely with the trend seen in the hatchling and juvenile numbers; in general, as Pb extractability increased, the

number of hatchlings and juveniles decreased. All treatments had 100% survival (except for one replicate of shotfall soil) which suggests that bioavailability of Pb is low enough that it is not causing acute lethality in E. fetida in 63 days. Booth et al. (2003) also saw

100% survival in their shooting range soils albeit at lower total Pb concentrations (up to

292 mg/kg total soil Pb).

Two main models (quadratic and logarithmic) described trends in Pb bioaccumulation over 63 days. The quadratic model was the best fit for the off-site reference, negative control, on-site reference, and shotfall soils. The R2 was maximized using the logarithmic model for extracted and positive control soils. Earthworms exposed to shotfall soils reached the highest total Pb concentrations. All of the treatments exhibited a similar trend where total Pb in earthworms reached a peak and then declined.

Laskowski et al. (2010) described this trend as “overshoot”, when higher values are observed in the initial exposure phase of a metal. Laskowski et al. (2010) reported three cases of the overshoot trend for Pb accumulation in the literature for the following organisms: Lithobius forficatus centipede (Descamps et al. 1996), Allolobophora tuberculata earthworm (Neuhauser et al. 1995), and Orchesella cincta collembola

(Posthuma et al. 1992). In this study, overshoot occurred between days 0 and 14 for on- site reference and extracted soils. For shotfall soil, overshoot occurred between 14 and 63 days. Finally, in the spiked, positive control soil, overshoot occurred between 7 and 28 days. Laskowski et al. (2010) suggested that the overshoot points may not be outliers, but may be due to physiological mechanisms involved in the adaptation to and tolerance of metal uptake. In Argasinski et al. (2012), the authors model the same trend with a toxicokinetics cell demography model (TKCD). In this model, metals can damage gut

epithelial cells which can change absorption and excretion. For example, the higher the concentration, the more cell mortality that occurs, and the less metal absorption that occurs. Eventually, invertebrates may shed these epithelial cells and thereby reduce total body concentrations. Shedding mechanisms are common in invertebrates such as

Collembola (Posthuma et al. 1992) and aquatic midges (Groenendijk et al. 1999). Andre et al. (2009), found that most Pb in earthworms is associated with the posterior alimentary canal which includes the intestine and chloragogenous tissue. Therefore, shedding of these cells could cause an overall decline in the total body burden of Pb.

Shedding of gastrointestinal cells may be the reason why there was large variation in total earthworm Pb during the overshoot phase. For each of the soil treatments, during overshoot, the greatest difference between minimum values and maximum values was observed. This large variation may be due to differences in the timing of intestinal epithelial cell shedding by individual worms. In future studies, it would be interesting to monitor overshoot in greater resolution by sampling additional worms at time points between the ones used in this study.

Earthworms collected from the extracted and shotfall areas of the shooting range contained higher total Pb concentrations than earthworms exposed to the same soils for

28 days, while this difference was not evident for earthworms exposed to the on-site reference soil. Overall, total Pb levels in the three soils (extracted, on-site reference, and shotfall) were not statistically different, so it was surprising to see that the earthworms exposed to the on-site reference soil did not fit into the same trend as the earthworms exposed to the extracted and shotfall soils. Kaufman et al. (2007) also examined the metal content of field-collected earthworms at a shooting range site (5,000 mg/kg average total

soil Pb). They collected three species of Aporrectodea earthworms and found that worms contained on average 730 mg/kg Pb which is lower than the average Pb levels for shotfall and extracted area earthworms in the present study. This difference could be due to species-specific uptake and elimination (Lumbricus vs. Aporrectodea), or adaptation or tolerance at either site (affecting the bioaccumulation of Pb). Kaufman et al. (2007) did not conduct laboratory bioaccumulation or reproduction tests and therefore, adverse effects to earthworms (or lack thereof) are unknown. In both studies, field bioaccumulation factors (BAF; total Pb in the organism divided by total Pb in soil) were less than one which indicates low bioaccumulative properties of Pb at all of the sites.

One major difference between lab-exposed and field-collected earthworms in the present study is that they were two different species: E. fetida (lab-exposed) and

Lumbricus spp. juveniles (field-collected). Only juvenile Lumbricus spp. earthworms were collected from the field site which limited the identification to genera. There are two Lumbricus species present in Ohio: Lumbricus terrestris and Lumbricus rubellus

(Hale 2007). Both Lumbricus species have a wide tolerance of habitat factors (fields, stream banks, etc.), are anecic (deep-burrowing species), and ingest much soil during burrowing (Reynolds 1977). E. fetida is an earthworm that is an epigeic (surface dwelling) species and found in habitats that contain large amounts of organic matter, such as manure and compost heaps (Reynolds 1977). In captivity, Lumbricus earthworms can live as long as six years, while Eisenia earthworms can live as long as four years. The main difference between Eisenia and Lumbricus seems to be its behavioral ecology and the preferences for burrowing and feeding. Reynolds (1977) stated that Eisenia does not ingest much soil during feeding. However, Lumbricus can ingest large amounts of soil

during burrowing (Reynolds 1977). This difference alone could be a reason why field- collected Lumbricus earthworms had higher concentrations of Pb than lab-exposed

E.fetida earthworms, but it does not explain why the on-site reference earthworms did not contain different Pb concentrations than the lab-exposed earthworms.

When conducting field studies, it is common to assume that field organisms are at steady state with respect to chemical bioaccumulation from their environment. However, based on the graphs in Figure 2.3, it appears that only on-site reference, extracted, and positive control earthworm exposures resulted in steady state of Pb in earthworm tissues after 63 days. Maturation from a hatchling to a clitellate adult for Lumbricus earthworms can take between 82 and 213 days and is temperature dependent (Daniel 1990). Based on these maturation timeframes, earthworms collected at the field site were exposed to Pb for less than 82 days (because they had not reached sexual maturity). Therefore, it can be assumed that the field-collected earthworms from the on-site reference and extracted areas may not have reached steady state for Pb accumulation. This assumption is made with caution, however, because the laboratory-exposed earthworms were a different species than those collected in the field and Pb bioaccumulation may be species-specific.

Finally, invertebrate adaptation or acclimation to metals is not well studied

(Posthuma 1993), but is a plausible explanation for the results. Standard laboratory exposure tests (like the 28-day exposure discussed here) are only measures of relatively recent individual exposures. However, field organism populations may undergo long- term pollution stress and may be locally adapted to high metal exposure (Morgan et al.

2007). To this end, the percentage of Pb in the detoxified or MRG fraction of earthworms was higher in field-collected worms compared to lab-exposed worms. The non-MRG

fraction (toxic fraction according to Jones et al. (1999)) in the field-collected earthworms was lower than the lab-exposed earthworms. This difference could be due to longer exposure times (and subsequent formation of more MRGs), or phenotypic plasticity of the chloragosome organelles (Andre et al., 2009). Under contaminated conditions, choloragosomes are electron-dense, irregularly-shaped, and have higher metal concentrations (Andre et al., 2009). Future studies should measure MRGs over a geometric series of time points after exposure of naïve earthworms to shooting range contaminated soils. The time series should contain good sampling resolution around the time point for overshoot to determine if MRG levels change with total Pb levels. It would also be worthwhile to collect adult earthworms from a contaminated site, rear them on clean soil in the lab, and then expose future generations to contaminated soils.

It was hypothesized that total Pb concentrations for non-depurated earthworms would be higher than Pb concentrations in depurated earthworms, because soil in the earthworm digestive tract can account for a large proportion of total earthworm body weight. In earthworms exposed to soils with lower amounts of total Pb (negative control soil and extracted soil), there was no significant difference between depurated and non- depurated earthworms (Figure 2.6). The data suggest that depuration matters, especially for soils with high total Pb concentrations. Future studies should address why (or why not) depuration was used to answer the specific questions of the researchers. For example, for toxicokinetic studies, it is important to depurate earthworms so that the contaminant value obtained is reflective of the tissue values. However, in a trophic level feeding study, it may be important to use non-depurated prey items before analysis (also

being mindful of any terms in the risk assessment equations). It is important to keep in mind the end goal and question when deciding whether or not to depurate earthworms.

Total Pb concentrations were lowest in the negative control soil and the off-site reference soil. The on-site reference area was chosen to be intermediate to the off-site reference soil and the other shooting range soils because this area was located outside of the normal shotfall area. However, one sample from the on-site reference soil exceeded the minimum for the extracted and shotfall areas. This could be due to stray pellets in the on-site reference area, or the manual movement of soil (by the shooting range manager) from shotfall areas to other areas of the range. The only movement of soil at the range

(from 2007–2014) was the wet screening removal of Pb from the extracted area in 2009.

However, soil could have been moved prior to 2007. Jørgensen & Willems (1987), estimated that Pb pellets can take more than 100 years to completely transform.

Therefore, it is possible that the on-site reference area could have been contaminated with pellets or soil prior to the present work.

Total Pb levels in soils at the shooting range were highly variable (CV; Table

2.1). Soil samples were composites of four soil cores (to depth of 15 cm) along transects in various areas of the shooting range. Since the distribution of Pb shot is heterogeneous at the shooting range site, it is not surprising that the total Pb levels in the three composite samples in each area were highly variable. The spiked positive control soil, off-site reference, and negative control soil all had lower CV, suggesting they were likely more homogenous with respect to the mixing of Pb in the soil. The methods for sieving

(<2mm) and hand-picking Pb shot out of soil samples before analysis, may not have been completely successful. Hui (2002) reports that radiographing methods are best for

identifying and removing all Pb shot from soil samples. Leftover Pb shot pellets (missed by sieving and hand sorting) could have added to the variability in the soil samples since one fragment could skew total Pb results.

Percent ammonium nitrate extractable Pb was significantly lower for the off-site reference soil and negative control soil compared to the extracted and positive control soils. The extracted, on-site reference, shotfall, and positive control soils were not significantly different. The highest extractable Pb was found in the extracted and positive control soils. Extractable Pb is a measure of the easily soluble trace element concentrations in soil (ISO 19730:2008(E)). It is interesting that the positive control did not have the highest extractable Pb since the soil was spiked with Pb salts which are readily soluble.

CEC was significantly lower for the extracted and shotfall soils compared to the other soils. CEC is the total quantity of exchangeable cation sites per unit mass of dry soil

(Brady and Weil 2010). Clay and silt loam soil CEC range is 15–25 meq/100 g. All of the soils in this study fell within this range (except for the extracted soil which was slightly lower than the other soils (12.9 meq/100 g)). This suggests that although CEC levels were significantly lower for the extracted and shotfall areas in terms of statistical comparison to the other soils, the CEC levels were typical of clay and silt loam soils and may not be functionally lower for the type of soil.

Conclusions

To our knowledge, this is the first study to evaluate (with laboratory and field- based methods) a shooting range field site after Pb removal using wet screening. Even

after Pb shot removal, high levels of Pb still remain in the soil that are capable of bioaccumulation and causing reproductive effects in E. fetida. Earthworms collected in the extracted area contained similar concentrations of lead as the shotfall area. This preliminary study suggests that wet screening Pb removal techniques did not reduce impacts to earthworms. It is recommend that future studies focus on a more comprehensive sampling plan at multiple shooting ranges employing the wet screening technique and other Pb removal techniques.

Figure 2.1: Bioaccessibility and bioavailability of cationic metal (Mn+) to soil organisms. Figure from McLaughlin and Lanno (2014) with permission of the authors.

Figure 2.2: Aerial photograph of the a) shooting range with shotfall, extracted, and on- site reference areas and b)off-site reference area.

Table 2.1: Soil characteristics for the shooting range and off-site reference area field soils, and laboratory soils. Mean ± SE for soil characteristics (n = 3 for each soil). The same letter within rows indicates no significant difference between means at p > 0.05, using a Tukey’s post-hoc test. CV = coefficient of variation.

Off-Site Reference Negative Control On-Site Reference Extracted Shotfall Positive Control Overall ANOVA

Total Pb (mg/kg) a23.0 ± 0.3 a 13.4 ± 0.5 b3405.5 ± 2686.9 b 3230.3 ±1113.9 b 7925.7 ± 3655.3 b 1869.9 ± 110.3 p = 1.3 x 10-6

Total Pb CV (%) 2 6 136 60 80 10 -3 Extractable Pb* (%) a0.0 ± 0.0 a0.0 ± 0.0 a,b0.31 ± 0.11 b0.61 ± 0.22 a,b0.16 ± 0.07 b0.67 ± 0.09 p = 2.5 x 10 -4 pH b 7.2 ± 0.03 a5.6 ± 0.01 a,b6.5 ± 0.4 a6.1 ± 0.1 a,b6.4 ± 0.2 a5.6 ± 0.01 p = 6.5 x 10 -6 CEC (meq/100g) b24.2 ± 0.7 b23.2 ± 0.4 b23.5 ± 0.9 a12.9 ± 1.0 a15.0 ± 1.5 b23.2 ± 0.4 p = 1.8 x 10 -2 Al Oxide (mg/kg) b2055 ± 33 a,b1502 ± 34 a,b1699 ± 325 a,b1359 ± 177 a1306 ± 82 a,b1502 ± 34 p = 3.7 x 10 -3 Fe Oxide (mg/kg) b,c7607 ± 61 b2325 ± 52 c8243 ± 2233 c5751 ± 1657 c4454 ± 658 b2325 ± 52 p = 9.3 x 10 -3 Organic Matter (%) a,b5.1 ± 0.08 a4.8 ± 0.03 b5.9 ± 0.4 a4.2 ± 0.2 a4.3 ± 0.3 a4.8 ± 0.03 p = 1.7 x 10 -2 35 Carbon (%) a,b3.0 ± 0.07 a,b3.0 ± 0.008 b5.6 ± 1.0 a2.5 ± 0.4 a,b4.8 ± 1.0 a,b3.0 ± 0.008 p = 1.5 x 10

-2 Sand (%) 22.4 ± 0.7 30.5 ± 0.7 22.3 ± 5.2 23.3 ± 1.8 25.3 ± 1.3 30.5 ± 0.7 p = 8.0 x 10 -2 Silt (%) 38.2 ± 1.2 34.1 ± 1.2 32.6 ± 0.7 41.9 ± 1.2 40.6 ± 1.3 34.1 ± 1.2 p = 8.0 x 10 -1 Clay (%) 39.4 ± 0.7 35.5 ± 0.7 44.8 ± 5.8 34.8 ± 2.7 34.1 ± .03 35.5 ± 0.7 p = 1.0 x 10 * Extractable Pb (ISO 19730:2008(E) Ammonium Nitrate Extractable)

Table 2.2: Earthworm reproduction endpoints for field and laboratory soils. Mean (± SE) for number of cocoons, number of juveniles, and number of juveniles per adult for each soil type (n = 4 for each soil). The same letter within rows indicates no significant difference between means at p > 0.05, using a Tukey’s post-hoc test. Off-Site Negative On-Site Positive Overall Extracted Shotfall Reference Control Reference Control ANOVA Cocoons c22 ± 3 c21 ± 3 b,c14 ± 1 a,b6 ± 2 a2 ± 1 a,b9 ± 2 p = 1.6 x 10-5 Juveniles c66 ± 6 b38 ± 7 a10 ± 2 a6 ± 3 a1 ± 1 a5 ± 3 p = 3.5 x 10-9 Juveniles/Adult c6 ± 1 b4 ± 1 a1 ± 0 a1 ± 0 a0 ± 0 a1 ± 0 p = 6.9 x 10-9

Figure 2.3: Total lead (mg/kg, dw) bioaccumulation in Eisenia fetida exposed to soil treatments and sampled after 0, 7, 14, 28, or 63 days of exposure. Values on graphs are mean ± SE earthworm total Pb for (a) off-site reference, (b) negative control, (c) on-site reference, (d) extracted, (e) shotfall, and (f) positive control soils. Model equations and R2 values are shown on each graph.

10000 * *

1000 Pb(mg/kg)

100 Earthworm

10 Total Total

1 Negative Off-site Extracted On-site Shotfall control reference reference

Figure 2.4: Lab and field earthworm total lead. Mean ± SE total Pb for field-collected Lumbricus juveniles (shaded bars) and 28-day laboratory exposed E. fetida (open bars) earthworms. Asterisks indicate a significant difference between lab and field concentrations for that site (T-test; p < 0.05).

10000

1000

100

10 Total Pb (mg/kg) Pb Total

0.1 Lab Field Lab Field Lab Field Lab Field Negative Control Extracted On-site reference Shotfall Figure 2.5: Total Pb (mg/kg, dry weight) in metal-rich granule (MRG) and non-MRB fraction of laboratory-exposed or field-collected earthworms. Values are total Pb concentrations (mean ± SE) for MRGs (triangles), non-MRG fractions (squares), and whole earthworm (circles). Results are presented for field-collected Lumbricus sp. juvenile worms (open shapes) and 28-day lab exposed E. fetida worms (filled shapes).

Table 2.3: Total lead of earthworm MRG and non-MRG fractions for field and laboratory earthworms. Fraction percentages as compared to total earthworm lead for field-collected Lumbricus juveniles and 28-day lab-exposed E. fetida earthworms. Percentages do not add up to 100% because different earthworms were used for the whole earthworm analysis than the ones that were used for tissue fractionation. Also, the numbers represented here were calculated using means (n = 3) for each treatment type. N/A = no sample collected. Negative On-Site Extracted Shotfall Control Reference Field-collected MRG fraction N/A 76 N/A 51 Non-MRG fraction N/A 5 N/A 6 28-day lab-exposed MRG fraction 32 33 43 58 Non-MRG fraction 10 22 18 35

1600 * 1400

1200

1000 * 800

Earthworm Earthworm 600 Total Pb (mg/kg) Pb Total 400

200

0 Negative control Extracted On-site reference Shotfall

Figure 2.6: Total earthworm lead for depurated and non-depurated laboratory exposed earthworms. Mean ± SE for depurated (open bars) and non-depurated (shaded bars) 28- day laboratory exposed Eisenia fetida earthworms (n = 3 for each treatment). Asterisks indicate significant differences between depurated and non-depurated worms per soil type (T-test; p < 0.05).

Chapter 3: Community structure, distribution, and abundance of ground beetles

(Carabidae) along a lead gradient at a shooting range field site

Introduction

Shooting ranges result in high concentrations of metals, especially lead (Pb), deposited into ecosystems and depending on the concentrations present, these metals can have negative effects on communities of organisms that live at the site. Most laboratory studies focus on toxicity tests that take into account lethal and reproductive endpoints while field studies focus on the bioaccumulation of metals. However, fewer studies focus on the effects of metals on organism community biodiversity. Terrestrial invertebrate communities are often found in close proximity to soil and can be exposed to metals that are concentrated in the soil, providing an opportunity to assess the effects of metals on community-level parameters such as diversity and abundance.

Many invertebrates are being used as biological indicators of anthropogenic or natural change in ecosystems since they are relatively easy to collect and identify with proper taxonomic keys and training. Ranio and Niemelä (2003) suggest that with the proper study design, ground beetles (family Carabidae) can be used as bioindicators because of their known taxonomy, broad geographic ranges, and presence in diverse habitats. Stone et al. (2002) found that Pterostichus oblongopunctatus ground beetles collected at a smelter and mine site with high levels of Pb in the soil exhibited high body

42 residues of Pb. Soil concentrations at smelter and mine sites ranged from 136–2,635 mg/kg resulting in Pb concentrations in beetles ranging from 0.17–6.5 mg/kg

(males) and 0.36–8.7 mg/kg (females). Lagisz and Laskowki (2008) conducted a multigenerational study of Pterostichus oblonbopunctatus ground beetles exposed to a mixture of metals (zinc, cadmium, copper, Pb) at a smelter and mine site. Females collected at contaminated sites and then transferred to an uncontaminated medium were observed to produce more eggs than females collected from reference sites, and eggs from females collected at contaminated sites were also less likely to hatch. This study suggests that ground beetles exposed to high metal concentrations in soil may have altered life history parameters and that not only can beetles be collected to understand bioaccumulation of metals, but communities can be used as bioindicators of higher level ecosystem impacts.

A key component in field studies of bioaccumulation is the depuration of organisms before metals analysis. Standard laboratory bioaccumulation tests (e.g.,

Environment Canada 2007) require depuration in order to determine actual tissue metal concentrations without including soil or metals that are present in the stomach and digestive tract. However, in ecological risk assessment, depuration may not always be the preferred approach. For example, when estimating metal exposure of higher trophic level organisms, it would be beneficial to use non-depurated prey items for estimating dietary exposure. Predators, scavengers, and higher trophic level organisms consume food as-is, without depuration and therefore, metals and soil within the digestive tract of prey items should also be measured as part of the exposure equation. The decision to use depurated

43 or non-depurated prey items will depend on the goals and design of the study, but it is an important consideration.

The study area for this research was a private trap and skeet shooting range located in central Ohio, USA (Figure 2.2) and consisted of a shotfall area, extracted area

(where lead shot was removed by wet screening in fall 2009), and an on-site reference area. An off-site reference area was also chosen at a nearby park (Gallant Woods

Preservation Park, Delaware County, Ohio) for Pb tissue residue comparison only.

Bryant (2010) studied ground beetles at the shooting range site in 2008 and 2009 to determine their usefulness as bioindicators of Pb contamination and found that ground beetle abundance was higher in the reference area than the main shotfall area and that species richness did not differ between the two areas over the two years (2008 and 2009).

Since the initial study, the range owner excavated an area of soil where Pb shot was removed by wet screening in fall 2009, and replanted the field with ryegrass seed in

2010.

The first objective of this study is to understand how lead management, removal, and replanting of ryegrass may impact ground beetle communities. The second objective is to determine whether depuration of ground beetles would have a significant effect on

Pb body residues. It is hypothesized that:

1) If ground beetle communities are influenced by Pb in the soil, then all three

shooting range areas (shotfall, extracted, and on-site reference) will be dissimilar

in community composition. Furthermore, ground beetle abundance and taxa

richness will be reduced in the shotfall area and extracted area (due to higher lead

concentrations) when compared to the on-site reference area.

2) Beetles collected from soils with the highest total Pb concentrations will also have

the highest concentration of Pb in tissues and non-depurated beetles will have

higher total Pb concentrations than depurated beetles.

Methods

Carabidae community sampling

To collect ground beetles, non-baited pitfall traps were set up along a representative transect in the main shotfall, on-site reference, and extracted areas at the shooting range. Transects consisted of two rows of six traps spaced approximately 10 m apart. Soil cores were cut from the ground using a standard golf hole cutter (diameter =

10.8 cm, depth = 15 cm). A deli cup of similar size (946 ml) was fitted in the hole to sit flush with ground level. Landscape-edging lead-ins (approximately 51 cm long) were positioned at 90° angles from the cup at the four coordinates (North, South, East and

West). Finally, a rain cover (0.2 x 0.2 m square plywood) was installed above the deli cup to minimize the amount of rain that enters the cup (Figure 3.1) (see Bryant 2010 for more details). Propylene glycol (marine antifreeze; SuperClean brands Inc., Saint Paul,

Minnesota) was used as a surfactant and preservative in each of the cups (~100 ml).

Traps were serviced approximately every two weeks by pouring contents into bags and storing in a cold room (0–5°C) until sorting and identification. Samples were sorted to

45 separate Carabidae from other taxa and then were stored in 70% ethanol for identification to genus using common taxonomic keys (Arnett & Thomas 2001; Lindroth 1961-1969).

To determine if a sufficient number of samples was collected at the site, a taxa accumulation curve was created. The Chao2 Non-parametric Estimator (bias-corrected) in PC-ORD 6 (Gleneden Beach, Oregon) was used to estimate total species richness at the site. Chao2 uses occurrence data from aggregate samples to estimate the total number of taxa (Chao 1987) and is useful for determining if a sufficient number of samples was collected to capture the majority of species at the site (Gotelli and Colwell 2010).

To standardize samples for further analysis since some sampling periods were slightly longer than others, all abundance data was expressed as beetles/trap-day. To look for trends in abundance and richness over time, the abundance and richness data were plotted by day of the year. To determine differences between years, a one-way ANOVA using sampling area (shotfall, extracted, or on-site reference) as factors (R 3.1.0) was used, with a level of significance of α = 0.05.

The ground beetle community matrix was analyzed using non-metric multidimensional scaling (NMDS; PC-ORD 6). Before analysis, abundance data was log

(x + 1) transformed. The slow and thorough autopilot option with a Sorenson (Bray-

Curtis) distance measure was used. The autopilot method includes 250 iterations with real data followed by 250 iterations with randomized data. Subsequently, a Monte Carlo test of significance is conducted to determine the best ordination and dimensionality. The penalize feature to penalize ties with unequal ordination distances was used. Sampling area (shotfall, extracted, or on-site reference), and year were used as categorical variables

46 in our second matrix. Year and sampling area appeared to group sampling dates, therefore, an Analysis of Similarity (ANOSIM; Primer 6) was used to test for significant differences among communities for sampling area (shotfall, extracted, and on-site reference) and year (2010, 2011, and 2012). The data were separated by year and NMDS ordination analysis was conducted again with sampling area as a factor. NMDS was followed with ANOSIM using sampling area as a factor. To determine taxa that were driving differences in community data, an indicator species analysis (PC-ORD 6) was conducted to analyze groupings of taxa among shotfall, extracted, and on-site reference areas for each individual year. Indicator values are a combination of information on the abundance of a particular taxon within each group, and the faithfulness of occurrence within the group (McCune & Grace 2002).

Total Pb residues in Carabidae

Ground beetles were live-trapped at the shooting range site (shotfall, extracted, and on-site reference) and off-site reference areas for the determination of total body Pb residues. To eliminate possible contamination from sampling fluids, dry, unbaited pitfall traps were used. To reduce escapes (and increase trapping efficiency) traps were fit with a funnel (constructed from the top of a 2-L soda bottle) in the open top. Ground beetles were trapped from 07 October 2013 to 20 October 2013. All beetles analyzed for total Pb concentrations were of the genus Harpalus. Half of all beetles collected were rinsed with

MilliQ water and frozen for analysis (-70°C). The other half were depurated in a petri

47 plate with wet filter paper for 24 hours, rinsed with MilliQ water and then frozen for analysis (-70°C).

Beetles were dried to a constant weight at 60°C in an oven and Pb concentrations were determined in individual beetles. Sample preparation of beetles for Pb analysis was conducted by acid dissolution (Havlin & Soltanpour 1980). Instrumental analysis was conducted using inductively coupled plasma atomic emission spectrometry (ICP-AES) on an Agilent 720 (Sana Clara, California) according to USEPA method 6010C (USEPA

2007b) by the Soil Lab in the School of Environment and Natural Resources at Ohio

State University. Insect digestion included reagent blanks and certified reference materials (CRM) every 20 samples. The blanks were below the instrument detection limit for Pb. CRM Tort-2 lobster hepatopancreas (National Resource Council, Canada) was used for evaluating method recovery, which was within 10% of the certified value.

Instrumental analysis by ICP-AES was calibrated daily by serial dilution from at least two independent stock standards. Stock standards were prepared using ICP grade standards (SPEX CertiPrep Group, Assurance ICP Standards). The ICP linear calibration met the criteria of r2 = 0.995, and calculated standard concentrations within 10% for each standard used in the calibration. Initial calibration verification and continuing calibration verification was achieved using preparations from the certified LPC standard 1 mix QC standard (SPEX CertiPrep Group, Assurance ICP Standards). Continuing calibration verification standards were measured immediately after calibration and after every 10 samples. Standards fell within ± 10% of the certified value.

Results

Carabidae community structure

The experimental design and sampling regime accounted for 760 samples over 20 sampling dates between 2010 and 2012. Within these samples, 4,766 total ground beetles representing 23 different genera were collected. The samples are representative of the taxa at the site as suggested by the taxa accumulation curve approaching an asymptotic value (Figure 3.3). The Chao2 estimator of taxa diversity estimated a value of 24 taxa with a variance of 3 taxa. The taxa richness (23 genera) of the site falls within the range of the Chao2 estimator and therefore, the sampling was likely sufficient to capture the majority of genera present at the site. Based on the Chao2 estimator, with additional sampling, it is possible that up to 27 genera could be collected at the site. The samples were comprised mostly of five genera (Poecilus, Harpalus, Pterostichus, Agonum,

Scarites) that were found in greater than ten percent of samples. Two genera were represented by only one individual: Diplocheila, and Acupalpus.

Ten percent of the samples (76 individual samples) were found flooded or disturbed (cup removed from ground). Of the 644 collected samples, 81 samples did not contain ground beetle taxa and were therefore removed from the analysis. Seven sample dates were excluded from the analysis because they were collected in the spring through early summer of 2012 and the other years did not encompass those months. Therefore, the final analysis included 458 samples. There were no clear trends in total abundance per sampling date (Figure 3.3) or genera richness (Figure 3.4) for shotfall, extracted, and on- site reference areas over the three sampling years. However, when data was combined for

49 each year, there were differences among years (Tables 3.1 & 3.2). For all three years, the extracted sampling area had the highest average abundance of ground beetles (Table 3.1).

The extracted sampling area also had the highest average taxa richness for all three years

(Table 3.2).

The overall (including all three years) NMDS autopilot ordination (PC-ORD 6) resulted in a 3 dimensional solution that explained 76% of the variability. Axis one explained 27% of the variability and axis two explained 30% of the variability in the community dataset (Figure 3.5). Axis three (not shown in Figure 3.5), explained 19% of the variability. The final stress for the 3-dimensional solution was 18. A one-way

ANOSIM with sampling area (shotfall, extracted, on-site reference) as a factor was significant (p = 0.001). Year was also a significant factor (p = 0.001). Therefore, the data were separated by year and the NMDS ordinations and ANOSIM analysis using sampling area (shotfall, extracted, and on-site reference) as a factor were conducted.

NMDS ordination for 2010 resulted in a 3-dimensional solution that explained

86% of the variability, with a final stress of 15 (Figure 3.6). ANOSIM with sampling area as a factor suggested that the groupings were significantly different among sampling areas (overall p = 0.001, Global R = 0.219). The ANOSIM pair-wise results also suggest that each sampling area is significantly different than the others (p = 0.001 for each pair- wise comparison). The NMDS plot shows distinct shotfall and extracted communities

(that slightly overlap each other). The on-site reference community encompasses both the extracted and shotfall communities (Figure 3.6). Indicator Species analysis suggests that

Amphasia, Chlaenius, Cincindela, Harpalus, Ophonus, Patrobus, and Pterostichus are

50 driving differences in communities (Table 3.3). Amphasia was only found in shotfall samples, while Cincindela was only found in extracted samples. All five other taxa were found in at least two of the sampling areas.

NMDS ordination for 2011 resulted in a 3-dimensional solution that explained

65% of the variability with a final stress of 19 (Figure 3.7). ANOSIM with sampling area as a factor suggested that the groupings were significantly different among sampling areas (overall p = 0.001, Global R = 0.294). ANOSIM pair-wise results were also significant for each combination (p = 0.001). The NMDS plot shows the on-site reference area encompassing most of the shotfall and extracted communities. All three communities overlap for the 2011 sampling year. Indicator Species Analysis suggests that

Agonum, Amara, Chlaenius, Colliurus, Harpalus, Patrobus, Poecilus, Pterostichus, and

Scarites are driving differences in these communities (Table 3.3). Colliurus had high site fidelity for the extracted sampling area (only found in that area in 2011). All other eight taxa were found in at least two of the sampling areas.

NMDS ordination for 2012 resulted in a 3-dimensional solution that explained

84% of the variability with a final stress of 14 (Figure 3.8). ANOSIM with sampling area as a factor suggested that the groupings were significantly different (overall p = 0.001,

Global R = 0.216). Pair-wise ANOSIM tests were also significant for each pair-wise combination (p = 0.001). The NMDS plot shows that the extracted and on-site reference communities occupy similar space in the ordination (similar communities). The shotfall community overlaps both the extracted and on-site reference area, but occupies less space. The shotfall community also has unique community features that do not overlap

51 the extracted and on-site reference communities. Indicator Species Analysis suggests that

Agonum, Harpalus, Poecilus, Pterostichus, and Scarites are driving differences among the communities for 2012 (Table 3.3). All five taxa were found in at least two of the sampling areas. Of the 12 indicator taxa presented in Table 3.3, three taxa had high site fidelity and were only found in one sampling area: Amphasia (shotfall), Cincindela

(extracted) and Colliurus (extracted). Chlaenius was never found in the shotfall sampling area.

Total Pb residues in Carabidae

Total Pb concentrations in Harpalus non-depurated ground beetles were significantly higher than depurated beetles from the same area (Figure 3.9). T-tests were not possible to compare the shotfall or on-site reference depurated versus non-depurated because too few samples were collected.

Discussion

Over the three-year sampling period, a total of 23 different ground beetle taxa were collected. Lövei and Sunderland (1996) suggest that is common to find between ten and forty species of ground beetles in the same habitat during one season. The extracted sampling area had the greatest ground beetle abundance over the three sampling years compared to the shotfall and on-site reference areas which both had lower abundance

(Table 3.1). It was hypothesized that ground beetles would be most numerous in the on- site reference area because this area had the lowest levels of Pb in the soil and due to its

52 close proximity to a tree line which would provide a more heterogeneous habitat.

Therefore the data do not support the hypothesis. Lock et al. (2001) found no correlation of ground beetle abundance along a pollution (Pb and zinc) gradient at a historical mining area which supports the findings of the current study.

The extracted area was recently disturbed by the removal of Pb by wet screening in fall of 2009 and again in spring of 2010 when the extracted area was slice-seeded with rye grass. The first sampling date for the present study was about one year after the initial disturbance of wet screening. The increased abundance could be due to disturbance, as disturbance can increase the abundance of ground beetles (da Silva et al., 2008). In a study of ground beetles along a land-use gradient, the greatest species abundance and richness was found in the highly disturbed homogenous agricultural open fields (da Silva et al, 2008). Vanbergen et al. (2005) also found a similar pattern with the highest abundance and richness located in areas with frequent agricultural disturbances. All three treatment areas at the shooting range were mowed occasionally throughout the spring through fall and all were mowed at the same time. This suggests that the differences in the extracted area are not due to mowing alone but from other impact(s). Although the traps were set along the same transect each year, there were differences in ground beetle communities among years. Carabids are known to have year to year differences in population abundance and therefore Niemalä et al. (2000) suggested that the whole activity season should be used to analyze communities. Other invertebrate studies also analyzed years separately because of community differences in beetles (Vanbergen et al.,

2005) and microarthropods and nematodes among years (Selonen et al., 2014).

It was hypothesized that the ground beetle communities would be dissimilar in each of the sampling areas. While significant differences among communities were found, there was also considerable overlap (Figures 3.6–3.8). The final stress for the overall NMDS ordination (for the 3-dimensional solution) was 18. According to McCune and Grace (2002), stress scores between ten and twenty can result in drawing false inferences and the user should not place too much reliance on the details of the plot.

However, the plot is representative of the community because ANOSIM suggested significant differences among shotfall, extracted, and on-site reference communities.

Stress for the individual years ranged from 14 to 19 which falls within the same range

(10-20) as the overall ordination.

In a study of forest-dwelling ground beetles, Shibuya et al. (2011) found that ground beetle distributions were related to the vegetation of the forest floor, soil surface temperature, and the abundance and quality of leaf litter on the forest floor. Since transects in the present study were all located in areas with similar vegetation (although not directly quantified), the impact from vegetation and leaf litter should be minimal. Soil surface temperatures can vary on small scales to create microclimates which can impact the ground beetle community in the immediate vicinity, mostly in terms of availability of egg-laying sites (Tréfás and van Lenteren 2008). Kromp (1999) also points out that soil moisture and temperature may be more important for forest-dwelling species than field- dwelling species. Temperature and soil moisture were not measured and therefore no conclusions can be drawn based on these parameters. Future studies should take into

54 account soil temperature and moisture to determine how these parameters affect ground beetle communities.

The indicator species analysis suggested that some species are more common and have higher site fidelity for certain sampling areas (Table 3.3). Amara, Ophonus, and

Pterostichus genera were indicators of the shotfall sampling area. These three genera are mostly herbivorous and/or seed eaters (Lindroth 1961-1969), although Pterostichus has also been found to feed on aphids (Kromp 1999). Most Pterostichus species in North

America are flightless (Lindroth 1961-1969) which suggests that these beetles have a limited home range based on non-flying movement. Beetles without functional wings would not be able to move to more favorable conditions and therefore may be tolerant or resistant to the high Pb concentrations of the shotfall area because the genus cannot fly to a new area. Agonum, Cincindela, Colliurus, Harpalus, Poecilus, and Scarites were indicators of the extracted sampling area. All six of these genera (with the exception of

Harpalus and Poecilus which are herbivorous) are carnivorous (Lindroth 1961-1969). All six of these genera are also able flyers as adults (Lindroth 1961-1969) and could move to more suitable conditions. It is interesting that these genera were found to be indicators of the recently disturbed extracted sampling area. This recently disturbed area may have a greater diversity of prey items and/or vegetation to feed on, resulting in a greater diversity of beetles occupying that sampling area. Chlaenius, Patrobus, and Poecilus were indicators of the on-site reference area. Chlaenius is mostly a carnivorous genus and most are excellent flyers as adults (Lindroth 1961-1969). These beetles are mostly hygrophilous and so it is interesting that they were found in a field away from permanent

55 standing water. The on-site reference area was prone to temporary standing water after heavy rains (personal observation); however, standing water and rainfall were not directly measured and therefore cannot be correlated with the presence of this genus. Poecilus is a mostly herbivorous genus and is found in fields and open areas (Lindroth 1961-1969).

Patrobus is mostly carnivorous and found in open country (Lindroth 1961-1969). The on- site reference area is located in close proximity to a tree line and this ecotone could be the reason for a different community structure.

The ground beetle community was not analyzed according to feeding guild or trophic level since most ground beetles are polyphagous and opportunistic feeders that eat both plant and animal material (Kromp 1999). Furthermore, many ground beetle species use a combination of extraoral and internal digestion which complicates estimates of dietary composition based on gut-content analysis. Most studies of ground beetle dietary composition have occurred in the lab by offering various diet items to the beetles and observing which items were eaten. While these studies show that ground beetles will eat a particular item, they do not show the true field condition and dietary preferences.

Without confidence in feeding guild or trophic level, that analysis would not be very meaningful.

The ground beetle community in the present study was not analyzed based on the presence or absence of functional wings because many species exhibit both flying and non-flying forms. In many species, only dispersing individuals have functional wings.

However, what is known about the function of wings in the genera collected was discussed because it could be an important next step for a future study. Lagisz et al.

(2008) found that body length of beetles collected at mining sites (Pb and zinc contamination), was shorter than beetles from reference sites. Measures of wing functionality should be incorporated into future analysis of ground beetles at contaminated sites. Whether or not an individual inhabits a polluted site could provide insight into which taxa may have means to tolerate high concentrations of contaminants or perhaps which (if any) populations at a site may have adapted to the contaminant concentrations and inherited resistance at the site.

In a study of microarthropod communities at a shooting range, Migliorini et al

(2004) found that communities were not largely affected by contaminants. However, certain groups (Collembola, Protura, and Diplura) had higher abundances in contaminated areas than non-contaminated areas. Symphyla was the only taxa that was sensitive to high concentrations of metals and was not found in heavily contaminated sites. This study points out the importance of looking at community structure (and not just total invertebrate abundance) because when certain taxa are absent, others may be more abundant. The indicator species analysis in the present study identified genera that may be more tolerant of Pb than others.

Total Pb concentrations in depurated Harpalus ground beetles were highest in beetles collected at the shooting range (shotfall, extracted, and on-site reference) when compared to the off-site reference area (overall p = 2 x 10-3; ANOVA with pair-wise comparisons). Total Pb concentrations in non-depurated Harpalus ground beetles were higher in beetles collected at the extracted and shotfall areas when compared to the off- site reference and on-site reference areas (overall p = 6 x 10-4; ANOVA with pair-wise

57 comparisons). There was no difference between non-depurated beetle total Pb between the on-site reference and off-site reference (p = 0.07). Higher total Pb concentrations were found in non-depurated ground beetles collected from the extracted area than depurated beetles (p = 1.8 x 102; T-test; Figure 3.9). This supports the hypothesis and is likely due to soil remaining in the digestive tract when beetles were not permitted to depurate prior to metals analysis. However, no difference was found between the depurated and non-depurated beetles collected from the off-site reference area (p = 0.7;

T-test). Since total Pb concentrations in soil were lowest in the off-site reference area

(Table 2.1), soil and diet items at the off-site reference area would not have been expected to contain high concentrations of Pb. Total Pb levels in non-depurated beetles from the extracted area were almost two-fold higher than in depurated beetles from the same area. This suggests that the diet of Harpalus contains Pb, but that not all ingested

Pb is incorporated into tissues. Harpalus beetles are largely herbivorous seed eaters and may be ingesting Pb in the form of soil dust on the outside of seeds or plant material.

Ground beetles are able to efficiently regulate a number of metals. At smelter and mine sites, Stone et al. (2002) found that Pterostichus oblongopunctatus ground beetles collected at the highest soil Pb concentrations also exhibited the highest tissue residue concentrations of Pb. Soil concentrations at the smelter and mine sites ranged from 136 mg/kg to 2,635 mg/kg. Beetle tissue Pb concentrations ranged from 0.17–6.5 mg/kg

(males) and 0.36–8.7 mg/kg (females). In the present study, depurated beetles from shooting range sites ranged from 5.85 mg/kg total Pb to 12.37 mg/kg total Pb. The average beetle total Pb from the present study equals the highest value (8.7 mg/kg) from

58 the mining study by Stone et al. (2002). This could be due to differences between

Harpalus and Pterostichus in species specific uptake and elimination, diet, or soil ingestion and Pb bioavailability.

In a feeding study, Maryański et al. (2002) found that Poecilus cupreus larvae could regulate zinc (Zn) very efficiently and could somewhat regulate cadmium (Cd), although beetles still accumulated more Cd than control beetles. For Cd, the beetle larvae never reached concentrations higher than diet items (house fly larvae). In the present study, beetles did not accumulate Pb to concentrations higher than found in the soil, but they still contained higher Pb concentrations than beetles collected at an off-site reference area.

In a study of forest ground beetles in Michigan, U.S.A., Gandhi and Herms (2008) found that a large proportion of the collected ground beetle community had morphological anomalies. Examples of anomalies include: malformed elytral striae, depressions on the head, perforations of the elytra, tumors, and lateral lines on the head and pronotum. Gandhi and Herms (2008) suggested a few reasons why the anomalies exist including: pollution, acid rain, nitrogen deposition, genetic inbreeding, and parasites. While identifying individuals for the present study, no obvious anomalies were recorded as in Gandhi and Herms (2008).

Conclusions

This is the first study of ground beetle communities at a shooting range site after

Pb removal by wet screening. Communities were significantly different from each other

59 in each of the sampling areas. Differences in the extracted (wet screened area) may be due to multiple disturbances (wet screening and subsequent slice-seeding) at the site.

Based on field-collected tissue concentrations, it does not appear that wet screening reduced total tissue Pb to the reference condition. Future studies at other shooting ranges should include locations with similar management techniques used for the removal of Pb to determine whether these extraction techniques could actually reduce the amount of bioavailable Pb. Future studies should also incorporate other taxa (e.g., ants, Formicidae) into the community analysis to obtain a more complete picture of community interactions

(Lövei and Sunderland 1996).

Figure 3.1: Example pitfall trap with deli cup (not visible) that sits flush with ground level, four landscape-edging lead-ins, and rain cover.

Cumulative Number Taxa Number Cumulative 5

0 0 100 200 300 400 500 600 Cumulative Number Samples

Figure 3.2: Ground beetle taxa accumulation curve (cumulative number of samples vs. cumulative number of taxa) for the duration of the study. Each sample is represented as a dot for the curve and the curve was not smoothed.

2.5 a

day - 1.5

Beetles/Trap 0.5

0 215 225 235 245 255 265 275 285 295 305 2010 Sample Date (day of year)

2.5 b

day - 1.5

Beetles/Trap 0.5

0 215 225 235 245 255 265 275 285 295 305 2011 Sample Date (day of year)

6 c 5

day) - 3

1 Beetles/Trap

0 215 225 235 245 255 265 275 285 295 305 2012 Sample Date (day of year) Figure 3.3: Mean ground beetle abundance ± SE (beetles/trap-day) for shotfall (triangle), extracted (circle) and on-site reference (square) over the duration of the 2010 (a), 2011 (b), and 2012 (c) sampling seasons. Means are plotted against the day of the year that each sample was collected.

4.5 4

3.5 day - 3 2.5 2 1.5

1 Richness/Trap 0.5 0 215 225 235 245 255 265 275 285 295 305 2010 Sample Date (day of year)

4.5 b 4

3.5 day - 3 2.5 2 1.5

1 Richness/Trap 0.5 0 215 225 235 245 255 265 275 285 295 305 2011 Sample Date (day of year)

4.5 c 4

3.5 day)

- 3 2.5 2 1.5 1

Richness/Trap 0.5 0 215 225 235 245 255 265 275 285 295 305 2012 Sample Date (day of year) Figure 3.4: Mean ground beetle richness ± SE (richness/trap-day) for shotfall (triangle), extracted (circle) and on-site reference (square) over the duration of the 2010 (a), 2011 (b), and 2012 (c) sampling seasons. Means are plotted against the day of the year that each sample was collected.

Table 3.1: Mean ground beetle abundance (beetles/trap-day) for each sampling year for shotfall, extracted, and on-site reference areas. P values are given for significant ANOVAs (R 3.1.0) with sampling area as a factor. Different letters within rows indicates a significant difference in a Tukey’s post-hoc test (R 3.1.0). Shotfall Extracted On-site reference Overall p 2010 0.60a,b 0.83b 0.54a 0.031 2011 0.31a 1.41b 0.64a 1x 10-4 2012 0.22a 0.56b 0.24a 1x 10-4

Table 3.2: Mean ground beetle taxa richness for each sampling year for shotfall, extracted, and on-site reference areas. P values are given for significant ANOVAs (R 3.1.0) with sampling area as a factor. NS given for non-significant overall ANOVAs. Different letters within rows indicates a significant difference in a Tukey’s post-hoc test (R 3.1.0). Shotfall Extracted On-site reference Overall p 2010 2.3 2.5 2.2 NS 2011 1.7b 3.0a 2.3b 1 x 10-4 2012 1.2a,c 1.9b 1.4b,c 3 x 10-3

Axis 2

Axis 1

Figure 3.5: Overall non-metric multidimensional scaling (NMDS) ordination of shotfall (triangle), extracted (circle) and on-site reference (square) Carabidae beetle data. The plot is comprised of axis 1 (r2 = 0.27), axis 2 (r2 = 0.30), and axis 3 (not shown, r2 = 0.19) that explained 76% of the variability; overall stress for NMDS analysis = 18. A one-way analysis of similarity (ANOSIM) with sampling area as a factor suggested that the groupings were significantly different among sampling areas (overall p = 0.001).

Axis 2 (24%)

Axis 1 (42%) Figure 3.6: NMDS ordination of shotfall (triangle), extracted (circle) and on-site reference (square) for 2010 Carabidae beetle data. The plot is comprised of axis 1 (r2 = 0.42), axis 2 (r2 = 0.24), and axis 3 (not shown, r2 = 0.20) that explained 86% of the variability; overall stress for NMDS analysis = 15. A one-way analysis of similarity (ANOSIM) with sampling area as a factor suggested that the groupings were significantly different among sampling areas (overall p = 0.001; Global R = 0.219). The ANOSIM pair-wise results also suggest that each sampling area is significantly different from the others (each p = 0.001).

Table 3.3: Ground beetle indicator species for shotfall, extracted, and on-site reference sampling areas for 2010, 2011, and 2012. The maximum group and p values are given based on Indicator Species analysis in PC-ORD 6. 2010 2011 2012 Genus Group(s) p Group(s) p Group(s) p Acupalpus Agonum extracted 0.0020 extracted 0.0160 Amara shotfall 0.0032 Amphasia shotfall 0.0344 Anisodactylus Badister Bradycellus Calleida Chlaenius on -site reference 0.0006 on -site reference 0.0168 Cincindela extracted 0.0028

Colliurus extracted 0.0050 Cyclotrachelus Dicaelus Diplocheila Harpalus extracted 0.0020 extracted 0.0002 e xtracted 0.0002 Notiophilus Ophonus shotfall 0.0020 Patrobus on-site reference 0.0020 on -site reference 0.0002 Poecilus on-site reference 0.0006 extracted 0.0198 Pterostichus shotfall 0.0008 shotfall 0.0002 shotfall 0.0002 Scarites extracted 0.0002 extracted 0.0006 Selenophorus Stenolophus

Axis 2 (24%)

Axis 1 (24%)

Figure 3.7: NMDS ordination of shotfall (triangle), extracted (circle) and on-site reference (square) for 2011 Carabidae beetle data. The plot is comprised of axis 1 (r2 = 0.24), axis 2 (r2 = 0.24), and axis 3 (not shown, r2 = 0.18) that explained 65% of the variability; overall stress for NMDS analysis = 19. A one-way analysis of similarity (ANOSIM) with sampling area as a factor suggested that the groupings were significantly different among sampling areas (overall p = 0.001; Global R = 0.294). The ANOSIM pair-wise results also suggest that each sampling area is significantly different from the others (each p = 0.001).

Axis 2 (30%)

Axis 1 (36%)

Figure 3.8: NMDS ordination of shotfall (triangle), extracted (circle) and on-site reference (square) for 2012 Carabidae beetle data. The plot is comprised of axis 1 (r2 = 0.36), axis 2 (r2 = 0.30), and axis 3 (not shown, r2 = 0.18) that explained 84% of the variability; overall stress for NMDS analysis = 14. A one-way analysis of similarity (ANOSIM) with sampling area as a factor suggested that the groupings were significantly different among sampling areas (overall p = 0.001; Global R = 0.216). The ANOSIM pair-wise results also suggest that each sampling area is significantly different from the others (each p = 0.001).

* 40

15 Total Beetle (mg/kg) Pb Beetle Total 10

0 shotfall extracted on-site reference off-site reference

Figure 3.9: Mean ± SE total Pb for non-depurated (shaded bars) and depurated (open bars) field-collected Harpalus beetles. Asterisk indicates a significant difference (p < 0.05; T-test) between depurated and non-depurated beetles for that sample area. Analysis was not possible for shotfall or on-site reference soils because too few samples were collected.

Chapter 4: Lead tissue residues, bioaccessibility estimates, and estimation of incidental soil ingestion of the Meadow Vole (Microtus pennsylvanicus) collected at a shooting range field site

Introduction

Lead (Pb) is a naturally occurring element that was historically mined and used for many applications such as batteries, lead-based paint, leaded gasoline, fishing lures, and ammunition. While Pb shot pellets were phased out by the U.S. Fish and Wildlife

Service (USFWS) for waterfowl hunting in wetlands in 1991 (U.S. 50 CFR § 20.108), Pb shot is still commonly used for upland game as well as at shooting ranges across the U.S

(Avery & Watson 1991). Approximately 4% of annual Pb use in the U.S. is for for bullets and ammunition (USEPA 2005a). There are over 10,000 military and private shooting ranges across the U.S. (USEPA 2005a) which results in the deposition of Pb shot on the soil surface.

A result of this surficial deposition of Pb is the exposure of terrestrial organisms, such as small mammals, to elevated concentrations of Pb. Three routes of exposure to Pb for small mammals include ingestion, inhalation, and dermal absorption and for small mammals living at shooting ranges, the primary route of exposure is ingestion of soil and contaminated diet items (Ma 1994). Inorganic forms of Pb do not easily cross the skin barrier, and inhalation of dust is considered negligible for ecological receptors at shooting ranges. Once Pb enters the bloodstream, it may be deposited into soft tissues and bone

(longer term). Pb is excreted via the kidneys in renal clearance, or in the feces via bile 73 release from the liver. Long-term Pb storage may occur in bone, where calcium is replaced by Pb, and around 90% of the total body burden of Pb may be deposited in bone

(Ma 2011). A range of toxic effects in mammals may be attributed to Pb exposure including neurological impairment, immunosuppression, hypertension, impaired vision and hearing, reduced fecundity, tissue and organ damage, paralysis, and death (Ma 2011).

Ma (2011) suggests that Pb accumulation in kidney tissue is an important biomarker for Pb exposure at contaminated sites because Pb levels in kidney tissue generally reaches steady state before an organism reaches adulthood. Background Pb levels in kidneys for shrews range from 3 to 11 ug/g dw (clay and peat soils) and 13 to 19 ug/g dw on sandy soil (Ma et al. 1992). Soils with lower pH tend to have higher Pb bioavailability as shown by Ma (1989) where earthworms and moles (Talpa europea) collected from sites with lower pH (but similar total soil Pb) had higher total Pb tissue levels.

The Somatic Kidney Index (SKI) is the ratio of total fresh weight of the kidneys to the total fresh body weight. Previous studies have found that SKI was elevated in mammals exposed to high Pb concentrations (Myodes glareolus and Sorex araneus) due to the presence of renal edema (Ma 1989). In addition to measurable differences in SKI and Somatic Liver Index (SLI), histopathology of kidney tissue in Pb exposed individuals may show acid-fast intranuclear inclusion bodies (Goyer 1970; Stansley & Rosco 1996).

Intranuclear inclusion bodies are formed in the kidney when Pb binds to protein complexes in the proximal tubule and this is a sign of acute Pb nephropathy. Stansley &

Roscoe (1996) found intranuclear inclusion bodies in 81 out of 100 renal tissue segments

74 and total Pb residues of 1,506 mg/kg in one individual Blarina brevicauda (Northern short-tailed shrew) collected at a shooting range site. However, out of 23 Peromyscus leucopus (White-footed mouse) collected at the same site, only two exhibited signs of intranuclear inclusion bodies in the kidneys and kidney Pb residues averaged 34.9 mg/kg.

Goyer (1970) reported critical renal Pb ranges of 25-40 mg/kg associated with the formation of intranuclear inclusion bodies and reductions in body weight and SKI for rats orally dosed with Pb salts in drinking water.

Lead can also affect the behavior of organisms. Punzo and Farmer (2003) exposed Blarina brevicauda (Northern short-tailed shrew) to Pb salts via drinking water

(25 mg/kg/day for 90 days) and found that exposed shrews spent more time using exploratory behavior and exhibited slower movements than control shrews. This change in behavior, compared to control shrews, could make Pb-exposed shrews more vulnerable to predation or other ecological stressors.

Microtus pennsylvanicus (meadow vole; Rodentia: Muridae) is a common small mammal in the United States and is found wherever there is abundant grass. Meadow voles are mostly herbivorous and are a common prey item in the diet of hawks, owls, and some mammalian predators (USEPA 1993). The home range of the species depends on season and sex of the individual, but previous studies suggest the area of the home range falls between 0.0002 ha and 0.083 ha (USEPA 1993; Madison 1980; Douglas 1976;

Ostfeld et al 1988). The meadow vole diet depends on season and food availability but includes mostly dicot and monocot shoots, and to a lesser extent, seeds, roots, fungi, and insects (USEPA 1993; Lindroth & Batzli 1984). Herbivores grazing on monocot and

75 dicot shoots, and to a lesser extent roots, at contaminated sites, could be exposed to Pb through their diet.

While the accumulation of Pb by plants at shooting ranges can occur either via aerosol or soil sources, the most likely source of Pb to plants is from the soil, with accumulation dependent upon the plant species and the bioavailability of Pb in soil. Soil factors such as pH can alter the amount of dissolved Pb available for plant uptake and

Blaylock et al (1997) found very little Pb accumulation for plants associated with higher pH soils (between 5.5 and 7.5). Most of the Pb that enters plant tissues stays localized in the roots and is not translocated to the shoots (Lane & Martin 1977; Kumar et al 1995).

According to a review by Sharma & Dubey (2005), Pb levels in various plant parts decreases in the following order: roots > leaves > stem > inflorescence > seeds.

Furthermore, dicots tend to accumulate higher amounts of Pb in the roots than monocots

(Huang & Cunningham 1996). In a study of plants grown on acidic and calcareous shooting range soils, plants grown on acidic soil tended to accumulate more Pb than plants grown on calcareous soils (Evangelou et al. 2012). Pb concentrations in a variety of plants grown on the soils (soil total Pb 466 mg/kg and 644 mg/kg) ranged from less than 10 mg/kg to 62 mg/kg dw (Evangelou et al. 2012). In a study of plants collected from a mining site (total soil Pb 65.7 mg/kg), the above-ground portions of forbs and grasses accumulated 2.3 and 4.5 mg/kg Pb, respectively, while the below-ground portions accumulated 13.9 and 36.3 mg/kg Pb, respectively. Even though these levels of Pb approach background soil levels and are much lower than those typically found in soils at

76 shooting ranges, Pb from plants could be a source of exposure for herbivores at a contaminated site.

Another source of exposure to Pb at a shooting range site includes the direct ingestion of soil during burrowing, grooming, or while feeding. Direct ingestion of soil can be an important source of Pb, especially for mammals that consume ground-dwelling prey items or plants and seeds that are covered in soil or dust from the site. There are two main methods for estimating incidental soil ingestion by organisms. Beyer et al. (1994) used acid-insoluble ash of scat and digestibility of food to estimate soil within the diet of

28 species. They found that the diet of Microtus pennsylvanicus consists of approximately 2.4% soil. Using rare earth element tracers is also an option for estimating the percent of soil in the diet of an organism (Calbrese & Stanek 1995). These methods take into account the amount of tracer in the soil and the feces in order to calculate the percent soil consumed.

For organisms that may be exposed to soil or chemicals through ingestion, the bioaccessible fraction could be considered the amount of chemical that is dissolved in the stomach and available for absorption in the small intestine. Sometimes confusion arises when distinguishing between the terms extractable and bioaccessible. Soil extractions are different types of laboratory treatments designed to extract a specific fraction of chemical from the soil. For metals, there are many different types of extractions (e.g., ammonium nitrate and Mehlich). Extractions should only be considered to contain a portion of the bioaccessible pool of chemical and should not be used to predict bioavailability unless they have been validated using biological models (McLaughlin & Lanno 2014).

Techniques to estimate the bioaccessible fraction of chemicals in soil after incidental soil ingestion have been developed and validated for use in human risk assessment (e.g., Drexler and Brattin 2007; Schroder et al. 2004; Ruby et al. 1996; Ruby et al. 1993). These methods are physiologically-based extraction techniques that mimic the conditions in the stomach and/or the small intestine of mammal models such as swine and mice. The extraction results in a measure of bioaccessible metal that is correlated with relative bioavailabilty from in vivo laboratory studies. Validated methods can be used in site-specific risk assessments as estimates of metal bioavailability from soil that is incidentally ingested.

While bioaccessibility methods for some metals (e.g., Pb, arsenic (As)) are well developed for human risk assessment, little model validation exists for bioaccessibility models for ecological risk assessment. Early bioaccessibility methods for ecological receptors were developed for waterfowl (e.g., Turner and Hambling 2012; Furman et al.

2006; Martinez-Haro et al. 2004) to assess risks associated with Pb shot ingestion. For small mammals, As bioaccessibility has been examined in Sorex cinereus (masked shrews) (Moriarty et al 2010) and Peromyscus maniculatus (deer mice) (Ollson et al.

2009), while Kaufman et al. (2010) focused on developing a technique for use with multiple species exposed to Pb. While these authors used estimates of bioaccessible metals instead of total metal concentrations to estimate daily metal intake, no attempts were made to correlate bioaccessibility estimates with tissue concentrations. In order for bioaccessibility estimates to be useful for risk assessment, field and/or laboratory validation with actual bioavailability or bioaccumulation endpoints is necessary.

During a screening level assessment of a contaminated site, it is useful to first compare concentrations of chemicals in soil to ecological soil screening levels

(EcoSSLs). EcoSSLs are screening level chemical values of soils that were derived to avoid underestimating risk at a site. They are conservative numbers used for screening purposes only (never as a goal for clean-up). If a soil exceeds an ecoSSL for a chemical, then that chemical may be investigated in more detail during subsequent phases of the risk assessment. In the second phase of an ecological risk assessment, the chemicals considered in the screening level assessment may be examined in more detail by calculating a hazard quotient (HQ) using the highest environmental concentration of chemical at the contaminated site divided by a no observed adverse effect level

(NOAEL). The NOAEL is the highest exposure level at which there is no biologically significant increase in the frequency or severity of adverse effect between the exposed population and its appropriate control (USEPA 1997). The HQ can also be calculated by dividing an estimated dose of chemical by the NOAEL. The dose is the estimated amount of chemical that is taken in by an animal, in terms of body weight (e.g., mg contaminant/kg body weight per day) (USEPA 1993). The exposure dose is calculated using conservative assumptions regarding a number of mammal attributes such as area use factor, bioavailability, life-stage, body weight, food ingestion rates, bioaccumulation, and dietary composition (USEPA 1993). If the HQ exceeds one, then harmful effects cannot be ruled out for that receptor. If deemed necessary, a site-specific assessment may be designed to determine which receptors are at risk from contaminants at the site.

The study site for this research was a private trap and skeet shooting range located in central Ohio, USA (as described in Chapter 2). However, for the purposes of this chapter, the shooting range as a whole was compared to the off-site reference area.

Bryant (2010) studied the same shooting range site and found that of the three contaminants of potential concern at the site (antimony, As, and Pb), only Pb exceeded

EcoSSL values for herbivorous mammals (USEPA 2005). Therefore, this study is a continuation of assessing the risks posed by Pb at the shooting range through screening level risk assessment and site-specific assessment phases. The objectives of this study are: 1) To calculate HQs based on site-specific measures of incidental soil ingestion and bioaccessibility, and 2) To collect small mammal tissue samples from the shooting range and off-site reference area to determine if tissue Pb levels exceed described toxicity threshold values. It is hypothesized that:

1) When site-specific measures of incidental soil ingestion and bioaccessibility are

incorporated into dose calculations for a HQ, the HQ will be reduced.

2) Small mammal liver and kidney tissue residues will be higher at the shooting

range site than at the off-site reference area. Elevated Pb concentrations in soils

will result in higher SKI and SLI values in voles capture at the shooting range site

compared to voles captured at the reference area.

Methods

Small mammal tissue and stomach contents collection

The shooting range and off-site reference areas are describe in detail in Chapter 2 and small mammals were collected from these areas from July 2012 to October 2012 and

April 2013 to September 2013. Museum Special and Victor snap-traps baited with rolled oats and peanut butter were set in areas with obvious small mammal activity (e.g., in runways through the vegetation). Traps were set in the evening and checked every 12 hours. Animals caught in the traps were sacrificed via cervical dislocation from the snap- trap mechanism. This method was preferred for contaminant analysis because it does not contaminate tissues with additional chemicals (AVMA 2007). Animals that were trapped, but still alive and injured, were euthanized by cervical dislocation by trained individuals.

This research protocol was approved by the Ohio State University Institutional Animal

Care and Use Committee (IACUC # IS00000728). Necessary permits to conducted small mammal trapping were obtained from the state of Ohio Department of Natural Resources,

Division of Wildlife (ODNR DOW Permit #13-31 and #14-281) and from the Delaware

County, Ohio Preservation Parks (Permit # GWP-3-1-2012 and GWP-2013-3-20b).

All mammals were identified to species using a standard taxonomic key for mammals of the region (Kurta 2005) and only Microtus pennsylvanicus (meadow vole) were used for further analysis. The sex of each mammal was identified and the following measurements were recorded: total length, tail length, hind foot length, and weight.

Animals were dissected to obtain liver, kidney, and stomach contents for analysis. During

81 dissection, fecal pellets were collected when they were already formed and readily available in the rectum of organisms.

Total Pb analysis of tissues

Liver and kidney samples were dried to a constant weight at 60°C and then digested with concentrated nitric acid in Teflon crucibles on a hot plate. Indium was added as an internal standard, and the samples were brought up to a final volume of 100 mL with MilliQ water (by sample weight). Samples were analyzed by Inductively

Coupled Plasma Mass Spectrometry (ICP-MS) on a Perkin Elmer Sciex Elan 6000 ICP-

MS (Ontario, Canada). The accuracy of ICP analysis was assessed by measuring Pb concentrations in TORT-2 (lobster hepatopancreas, National Research Council of

Canada) a certified reference material (CRM). CRM samples were digested and analyzed in the same run with the mammal tissue samples to determine percent recovery of Pb in tissue samples. Percent recovery was between 100-110% (n = 4). The Pb content of blanks and check standards was measured every 10 samples.

Bioaccessibiliy of Pb in soil

Bioaccessible Pb was determined by using the Ohio State University in vitro gastrointestinal method (OSU-IVG) method (Schroder et al. 2004) which is a sequential extraction that simulates the physiological conditions of the human gastrointestinal system. This method has been developed for Pb and validated in swine dosing trials by

82 correlating IVG extractable Pb with blood Pb levels in swine fed the same soil (Schroder et al. 2004). The percent bioaccessible Pb was determined by equation 4.1.

퐼퐶푃 퐼푉퐺 % 푏푖표푎푐푐푒푠푠푖푏푙푒 = ∗ 100 [4.1] 푇표푡푎푙 푃푏

Element analysis of stomach contents and fecal samples

Stomach content, fecal, and soil samples were analyzed by the Ohio State

University Trace Element Research Lab (TERL). All samples were dried to a constant weight at 60°C and digested in 1:1 mixture of concentrated (68%) hydrofluoric (HF) and concentrated (48%) nitric acid (HNO3) on a hotplate, dried down, and redissolved in

HNO3 only. The final solutions were 3% HF and HNO3 by volume. Indium was used as an internal standard for Pb measurements. Samples were analyzed for the elements aluminum (Al), yttrium (Y), zirconium (Zr), cerium (Ce), praseodymium (Pr), neodymium (Nd), samarium (Sm), europium (Eu), gadolinium (Gd), terbium (Tb), dysprosium (Dy), holmium (Ho), erbium (Er), thulium (Tm), ytterbium (Yb), lutetium

(Lu), and Pb by Inductively Coupled Plasma Mass Spectrometry (ICP-MS) on a

PerkinElmer Elan 6000 (Ontario, Canada).

Data Analysis

Somatic kidney index (SKI; ratio of kidney fresh weight to whole body fresh weight) and somatic liver index (SLI; ratio of liver fresh weight to whole body fresh weight) were calculated for all voles collected at both sites. A t-test was used to test for differences between the shooting range and the off-site reference area for body weight,

83 liver weight, kidney weight, SKI, and SLI. Since no easy way to age voles exists, data were not age-normalized for statistical analysis. Total Pb concentrations in tissues and organ weight data were log transformed before analyzing to meet the assumptions of parametric statistics. Interaction plots and a one-way ANOVA with sex and collection site as factors were used to determine significant factors for total Pb in kidney and liver tissues. A t-test (R 3.0.1) was used to compare total Pb concentrations in kidney and liver between the shooting range site and off-site reference area.

For stomach and fecal samples, soil equivalent percentages were calculated by dividing the total concentration of each rare earth element in the stomach or feces by the concentration of that element in the soil (Calabrese and Stanek 1995). It was assumed that meadow voles in this study have a daily feeding rate of 0.35 g/g-day (USEPA 1993).

Average body weights of 32 g (shooting range) and 34 g (off-site reference area) were also used, and this equates to individuals eating approximately 11.2 g/day (shooting range) and 11.9 g/day (off-site reference area). The food digestibility estimate for

Microtus pennsylvanicus (0.56) from Karasov et al (1986) was also used. The approach of Mayland (1975) was used to calculate the organic fecal mass of plant origin (Equation

2, below) and then the mass of soil in feces (Equation 3). Once the mass of soil in feces was calculated, this was used to calculate the estimated percent of soil in ingested material. Mean soil equivalent percentages for five fecal samples from the shooting range

(8%) were used to estimate that 92% of the fecal matter was organic matter. Finally, the amount of Pb ingested per day was calculated by multiplying the amount of soil ingested per day by the total soil Pb concentration.

푑푟푦 푚푎푡푡푒푟 푖푛푡푎푘푒 (푔) − (푑푖푔푒푠푡푖푏푖푙푖푡푦 %)(푑푟푦 푚푎푡푡푒푟 푖푛푡푎푘푒 (푔)) [4.2] 푑푎푦

표푟푔푎푛푖푐 푓푒푐푎푙 푚푎푠푠 (푔) [4.3] 표푟푔푎푛푖푐 푚푎푡푡푒푟 푓푟푎푐푡푖표푛 표푓 푓푒푐푒푠

Hazard quotient calculation

Hazard quotients (HQ) for M. pennsylvanicus at the shooting range site were calculated following the methods of Kaufman et al. (2007). Equation 4.4 was used to calculate the estimated daily intake (EDI) of soil and food, where Ci is the concentration of Pb (mg/kg dw) in the soil or food, IRi is the ingestion rate (kg/d) of soil or food, RAFi is the relative absorption factor of soil or food, and BW is body weight (kg). Hazard quotients were calculated using equation 4.5 with a tolerable daily intake (TDI) of 4.70 mg Pb/kg-bw-day (NOAEL from feeding studies; USEPA 2005b). Daily soil ingestion rates estimated from this study were used in HQ estimates and it was assumed that plant food sources did not contain Pb (i.e., all ingested Pb came from soil). For our initial calculations, we assumed that RAF = 1 for soil. Subsequent HQ calculations were made by adjusting the RAF by using IVG bioaccessibility values for soil at the site. An area use factor was not included because all voles were assumed to spend 100% of their time foraging at the site.

퐶 푥 퐼푅 푥 푅퐴퐹 퐸퐷퐼 = ∑푚 푖 푖 푖 [4.4] 푠,푓 푖=1 퐵푊

퐸퐷퐼 + 퐸퐷퐼 퐻푄 = 푓 푠 [4.5] 푇퐷퐼

Results

Body weight, kidney weight, liver weight, SKI, and SLI were not significantly different between sites (Table 4.1). There was no significant difference between total Pb concentrations for liver or kidney based on sex or the sex*site interaction. Therefore, males and females were grouped together for each site and a t-test was used to compare sites. Liver samples from mammals trapped at the shooting range site had significantly higher concentrations of Pb than mammals trapped at the off-site reference site (p = 1.1 x

10-7; Table 4.2). Total Pb concentrations in kidney tissue were also higher at the shooting range site (p = 2.2 x 10-6; Table 4.2). Tukey’s post-hoc comparisons suggested that soil bioaccessible Pb (IVG Pb; Table 4.3) was lowest for the off-site reference and negative control soils; there was no difference among on-site reference, extracted, shotfall, and positive control (overall p = 7.9 x 10-4).

No significant difference in total dry weight for stomach content samples between shooting range and off-site reference samples were evident (Table 4.4; T-test p = 2.4 x

10-1), but total Pb concentrations (mg/kg) were significantly higher in the stomach content samples from voles sampled at the shooting range (Table 4.5; T-test p = 0.026).

Soil equivalent values from five fecal samples collected at the shooting range site (Table

4.6) were used to estimate that 4% of the diet of Microtus pennsylvanicus at the site was soil. When a HQ was calculated assuming an RAF of 1 for soil, the result was 12. When

IVG bioaccessibilty was incorporated into the RAF term for soil, the HQ decreased to 8.

Discussion

Lead concentrations in the liver and kidney of M. pennsylvanicus trapped at the shooting range site were significantly higher compared to Pb concentrations in the same tissues of voles trapped at the off-site reference area. Since the home range for M. pennsylvanicus is documented as 0.0002–0.083ha (USEPA 1993), which is smaller than the study site, it can be assumed that the collected voles spent 100% of their time on the site. This would suggest that voles were burrowing into contaminated soil, and consuming food that was contaminated with soil. Critical renal Pb ranges of 25–40 m/kg

(Goyer 1970) were clearly exceeded at the site as well as field thresholds for renal edema of 47 mg/kg (Ma 1989) and 25 mg/kg (Roberts et al. 1978). Eight-five percent of individuals collected at the shooting range site (20 out of 23) exceeded the 25 mg/kg threshold for renal edema. However, since samples were collected with snap-traps, renal edema could not be confirmed in the samples because histological analysis requires freshly sacrificed animals. No individuals from the off-site reference area exceeded any cited thresholds.

Lead concentrations in kidney (104.59 mg/kg) and liver (16.51 mg/kg) tissue from voles trapped at the shooting range were almost seven and three times higher, respectively, than Pb concentrations in kidney (15.8 mg/kg) and liver (5.1 mg/kg) samples from bank voles (Myodes glareolus) at a shooting range in the Netherlands (Ma

1989). The soil at the shooting range studied by Ma (1989) had a lower pH (4) compared to the soil from the shooting range in the present study (pH of 6) and lower pH generally results in more mobile forms of Pb and higher bioavailability. However, voles from the

87 present study (despite higher soil pH) had higher tissue Pb concentrations. Therefore, in future studies, incorporating a measure of bioaccessibility of Pb in soil could be useful for comparison. Meadow vole kidney Pb (104.59 mg/kg) from the present study was lower than kidney Pb (269 mg/kg) of Sorex araneus (Common Shrew) from a previous study in the Netherlands (Ma 1989). This difference was expected because voles are herbivorous and would accumulate less Pb than common shrews which are carnivores.

Carnivores are exposed to Pb through their prey (e.g., earthworms that accumulate Pb) and soil, while herbivores are only exposed to Pb through soil and soil dust on diet items.

Another possible major difference between the sites is distribution of Pb on the site. Ma

(1989) found a range of Pb concentrations in soil from 360 mg/kg to 70,000 mg/kg, while

Pb concentrations in soil from the shooting range in this study soil were consistently over

1,000 mg/kg. Lower Pb levels in tissues of bank voles in the study by Ma (1989) may suggest that voles were using less contaminated areas more frequently for foraging.

Furthermore, the site studied by Ma (1989) was closed approximately 20 years before the study; while the present study site is still in continuous operation.

No differences in body weight, organ weight, SKI, or SLI were found between the shooting range and reference sites. These results do not support hypothesis two. Results from a similar study (inside and outside a shooting range) by Ma (1989) did not find significant differences in body weight for the bank vole or Sorex araneus (common shrew). However, the body weights of Apodemus sylvaticus (wood mouse; an herbivorous small mammal) were significantly lower at the shooting range than at the

88 reference area. Ma (1989) also found that the SKI ratio was higher in the bank vole and common shrew collected at the shooting range.

Lead concentrations in kidney and liver tissue in the present study were also approximately three times higher than values reported by Stansley and Roscoe (1996) for

Peromyscus leucopus (white-footed mouse) of a trap and skeet range in operation for greater than 30 years. Stansley and Roscoe (1996) did not find significant differences in

SKI or SLI ratios between the shooting range and reference site. Values from the present study were below Pb levels for Northern short-tailed shrews (Blarina brevicauda) kidney

(1,506 mg/kg Pb) and liver (34.1 mg/kg) reported by Stansley and Roscoe (1996).

Northern short-tailed shrews are carnivorous small mammals and therefore higher organ

Pb residues are expected because of bioaccumulation from prey items (e.g., earthworms that accumulate Pb) and soil ingestion. Total Pb at this site was reported as 75,000 mg/kg with a soil pH of 6.3. However, this site differed from the present study site because it was covered in water most of the time. Inundation by water can have impacts on Pb bioavailability and subsequent exposure of small mammals to soil Pb. An interesting conclusion of Stansley and Roscoe (1996) was the estimation of soil ingestion based on the total body burden (estimated at 374 µg Pb). They estimated that the mean body burden could be accounted for by soil ingestion alone if only 0.06% of the diet was soil.

This was calculated by dividing the body burden (measured) by the products of food intake rate, soil Pb concentration (measured), fraction of Pb absorbed, and exposure time in days. The authors used standard literature values for food intake (4.6 g/day), exposure time (300 days), and fraction of Pb absorbed (0.65%). In the present study, approximately

4% of the diet of voles was estimated to be soil. This could explain the higher organ values in the present study (total body burden was not measured in the present study). In the present study, total Pb in fecal samples was higher than total Pb in stomach content samples. Not all Pb dissolved in the stomach will be absorbed in the small intestine, and therefore much Pb is excreted as waste in feces.

While Pb in diet items of M. pennsylvanicus was not estimated, it was estimated that 0.47 g soil is ingested per day by each vole at the shooting range site. With an average soil concentration of 3,876.7 mg/kg soil Pb, small mammals at the shooting range site consume approximately 1.8 mg/kg Pb per day (assuming that food item Pb is negligible and soil is the only source of Pb). This value is very close to values from Ma et al. (1991) for Microtus agrestis that were collected from an area (Budel site; soil Pb =

130 mg/kg) contaminated with Pb and Cd from a smelter. The estimate of soil in the diet

(4%) in our study is higher than in a study conducted by Beyer et al. (1994) which used the acid-insoluble ash method for estimating soil in the diet. The tracer method used in the current study required utilizing assumptions from the literature. In the future, it would be useful to do an actual comparison of both methods at the same site to determine if there are differences.

OSU-IVG bioaccessibilty extractions yielded values from 69–75% for shooting range soils. These numbers are comparable to another study by Bannon et al. (2009) that used the Drexler and Brattin (2007) method for estimating Pb bioaccessibility and estimated bioaccessibility between 83–100% for shooting range soils. . The method by

Drexler and Brattin (2007) differs from the OSU-IVG method in pH (1.5 vs. 1.8), gastric

90 solution (glycine HCl vs. 0.15 M NaCl and 1% porcine pepsin), and soild-to-liquid ratio

(1 g soil to 400 mL gastric solution vs. 4 g soil to 600 ml gastric solution). These parameters alone could result in slight differences in bioaccessibility values. Regardless, soil samples from both shooting ranges resulted in high bioaccessibility.

Hazard quotients calculated for soil ingestion at the site both exceeded one, even when IVG bioaccessibility was incorporated into the estimate. This does not support hypothesis one. It was expected that since voles were easily collected at the site that once site-specific parameters were incorporated into the calculation that the HQ would fall below one. If an organism is plentiful at a site, it can be assumed that the population is doing well and capable of reproduction. However, reproduction or population-level effects for meadow voles were not measured at the site. It is possible that the assumptions made when calculating the HQ (e.g., bioaccessibility, soil ingestion) were still very conservative estimates and it may be useful to refine these further for future estimates.

The bioaccessibilty estimate was conservative because it estimates the amount of Pb dissolved in the stomach and does not consider absorption in the small intestine (which is only a fraction of the amount dissolved in the stomach). Also, the estimates of soil ingestion in this paper were much higher than previous studies (Beyer 1994) which also adds conservatism to the estimate. The reader should take caution in using IVG bioaccessibility estimates for this purpose in the future. This part of the study was meant to be a demonstration of the potential use for bioaccessibility estimates if they are validated for ecological receptors. At this time, the OSU-IVG bioaccessibility assay has not been validated for use with M. pennsylvanicus or shooting range contaminated soils.

Conclusions

The elevated levels of Pb in kidney and liver tissue in Microtus pennslyvanicus trapped from the shooting range site suggest that these mammals are chronically exposed to Pb. Although voles are mostly herbivorous, they are still exposed to Pb through ingestion of soil or contaminated diet items. Even with corrections for bioaccessibility of

Pb in soil, a HQ greater than one was calculated for the shooting range site, suggesting that the Pb in soil may have negative impacts on reproduction, growth, or survival of the small mammals that reside there and further investigation may be necessary. While kidney Pb accumulation was at a level related to renal edema and other negative impacts,

M. pennsylvanicus remain abundant at the site. Future studies should attempt to understand how Pb accumulation may impact reproduction or other population-level effects in meadow voles. Since M. pennsylvanicus is such a common small mammal across the landscape, it makes up a significant diet component for carnivorous birds and mammals. Future studies should attempt to understand not only how Pb may be transferred via the food web to carnivores that prey upon voles, but also how negative impacts to reproduction could impact food availability for carnivores that live at or near the shooting range site.

Table 4.1: Mean (± standard error) for body, liver, and kidney weights for Microtus pennsylvanicus collected at the off-site reference area (n = 22 individuals) and the shooting range (n = 18 individuals) field sites. SLI = somatic liver index and SKI = somatic kidney index. Similar superscript letters within columns indicate no significant differenence between means (t-test; p > 0.05; R version 3.1.0). Body (g) Liver (g) Kidney (g) SLI (%) SKI (%) n Shooting range a 32.10 ± 2.50 a 0.64 ± 0.12 a 0.18 ± 0.03 a 1.87 ± 0.26 a0.53 ± 0.06 22 Off-site reference a 33.51 ± 2.44 a 0.67 ± 0.10 a 0.16 ± 0.03 a 2.12 ± 0.29 a0.51 ± 0.10 17

Table 4.2: Mean (± standard error) total liver and kidney Pb (mg/kg dw) for Microtus pennsylvanicus collected at the off-site reference area (n = 22 individuals) and the shooting range (n = 18 individuals) field sites. Superscript letters within columns indicate significantly different means (t-test; p < 0.05; R version 3.1.0). Site Total Liver Pb (mg/kg) Total Kidney Pb (mg/kg) Off-Site Reference a0.15 ± 0.03 a0.34 ± 0.081 Shooting Range b16.51 ± 2.2 b104.59 ± 16.68

Table 4.3: Soil characteristics for the shooting range and off-site reference area field soils, and laboratory soils. Mean ± SE for soil characteristics (n = 3 for each soil). The same letter within rows indicates no significant difference between means at p > 0.05, using a Tukey’s post-hoc test.

Off-Site Reference Negative Control On-Site Reference Extracted Shotfall Positive Control Overall ANOVA

-6 Total Pb (mg/kg) a23.0 ± 0.3 a 13.4 ± 0.5 b3405.5 ± 2686.9 b 3230.3 ±1113.9 b 7925.7 ± 3655.3 b 1869.9 ± 110.3 p = 1.3 x 10 -4 OSU-IVG Pb* (%) a18.7 ± 1.2 a19.3 ± 1.5 b75.0 ± 3.9 b74.8 ± 3.0 b69.2 ± 5.6 b82.7 ± 1.7 p = 7.9 x 10 * Ohio State University in vitro gastrointestinal method (OSU-IVG) according to Schroder et al. (2004)

Table 4.4: Total dry weight (µg) range, mean ± standard error, and n for stomach contents and fecal pellets for shooting range and off-site reference Microtus pennsylvanicus. For stomach contents, no significant difference was found between shooting range and off-site reference area in dry weight stomach contents (p = 0.237; t- test; R 3.1.0). Total dry weight (µg)

Range Mean ± SE n Shooting range

stomach contents 25.0–96.0 42.1 ± 9.3 7 fecal pellet 2.5–10.0 6.58 ± 1.4 5

Off-site reference

stomach contents 9.6–45.9 28.0 ± 6.2 5

Table 4.5: Total Pb (mg/kg) range, mean ± standard error, and n for stomach contents and fecal pellets for shooting range and off-site reference Microtus pennsylvanicus. Different superscript letters within columns for stomach contents indicates significant difference (T-test; p = 0.026). Total Pb (mg/kg)

Range Mean ± SE n

Shooting range

stomach contents 20.0–410.0 a168.6 ± 56.5 7

fecal pellet 70.0–1140.0 460.0 ± 194.7 5

Off-site reference

stomach contents 0.0–10.0 b2.0±2.0 5

Table 4.6: Soil equivalent values (total element concentration in feces (mg/kg, dw) / total element concentration in soil (mg/kg, dw)) for fecal samples obtained from voles trapped at the shooting range.

025F 011F 012F 026F 009F

Al 0.032 0.015 0.321 0.095 0.012 Y 0.025 0.025 0.254 0.094 0.013 Zr 0.024 0.012 0.148 0.040 0.011 Ce 0.024 0.024 0.411 0.154 0.018 Pr 0.020 0.026 0.419 0.156 0.017 Nd 0.025 0.026 0.410 0.154 0.018 Sm 0.023 0.030 0.360 0.139 0.019 Eu 0.000 0.000 0.307 0.117 0.000 Gd 0.013 0.020 0.318 0.119 0.013 Tb 0.001 0.001 0.234 0.064 0.001 Dy 0.000 0.010 0.220 0.083 0.007 Ho 0.000 0.000 0.199 0.066 0.000 Er 0.000 0.000 0.201 0.071 0.000 Tm 0.001 0.001 0.148 0.037 0.001 Yb 0.000 0.000 0.215 0.066 0.006 Lu 0.000 0.000 0.148 0.000 0.000 Average soil 1 1 27 9 1 equivalent (%)

Chapter 5: Conclusion

This study comprised a site-specific assessment to understand how ecological receptors at a shooting range site were impacted by Pb deposited in soil from shotgun pellets, utilizing various measures of bioavailability and bioaccessibility to assess exposure. Receptors at the shooting range that were examined included, ground beetles, earthworms, and meadow voles (conceptual model; Figure 1.1). A summary of findings for each of the receptors is provided below:

1) Earthworm laboratory tests revealed that bioaccumulation was the highest and

reproduction was significantly reduced in the shooting range soils, relative

earthworms exposed to the negative control soil and off-site reference soil. Field-

collected shooting range earthworms (Lumbricus spp. juveniles) accumulated

significantly higher concentrations of Pb that laboratory-exposed Eisenia fetida

earthworms. Furthermore, non-depurated earthworms from the most

contaminated soils contained the highest concentrations of Pb.

2) Ground beetles communities at the various locations of the shooting range site

were significantly different, however, it was concluded that this was likely due to

the disturbance of Pb removal by wet screening and subsequent slice-seeding of

the field and not from Pb in soil. Certain indicator taxa were identified as driving

differences between communities. Total Pb concentrations in ground beetles

were highest for beetles collected from the shooting range sites. Non-depurated

ground beetles from the extracted area contained higher concentrations of Pb

than depurated beetles from the same area.

3) Meadow voles appeared to be chronically exposed to Pb at the shooting range

site and had kidney and liver tissue residues above thresholds for organ damage.

Incidental soil ingestion of meadow voles at the site was estimated at 4% of the

diet, which combined with IVG bioaccessibility estimates, resulted in a HQ of

greater than 1.

Based on laboratory and field-based analysis, it appears that receptors at the site may be under stress from elevated Pb concentrations in the soil. Future studies of earthworms from shooting range sites should focus on determining if field earthworms are tolerant and/or resistant to high Pb concentrations. This study showed that earthworms at the field site store Pb in MRGs; future studies should examine if there are differences in MRG formation between naïve earthworms and earthworms collected from contaminated sites. Furthermore, conducting similar studies with multiple generations may shed light on true adaptations to Pb or mechanisms for tolerance.

Researchers should continue to consider the research goals when deciding to use depuration for field-collected organisms. Differences between depurated and non- depurated organisms were significantly different in the Pb-impacted area and that can have an impact on the results of a study depending on the questions and goals.

Ground beetle communities appeared to be affected more by physical disturbance of habitat due to management practices of the site rather than total Pb concentrations at the site. Ground beetle laboratory toxicity studies should be conducted

100 in the future to determine critical Pb concentrations in soil and in tissue residues for acute and chronic endpoints. These studies should choose taxa from different trophic levels, as well as ones that were indicators of the different shooting range areas. Before ground beetles can be used as indicators of contaminated sites, more studies need to be conducted to monitor trends in taxa abundance and richness at multiple sites.

Meadow vole tissue samples exceeded literature threshold values for organ damage, however, this study did not link high Pb tissue values to any type of histological change. Future studies should include histological analysis of tissue samples or some other valid endpoint to link organ concentrations with adverse effects. Future studies should also consider population-level effects such as adaptation of shooting range mammals to high Pb. Validation of IVG extractions for ecological mammalian receptors

(for food and soil) should be pursued in the future. Finally, this study focused on organisms that live in close association with the soil, but future studies should include higher trophic level consumers (e.g., carnivorous birds or mammals) that may feed on meadow voles or other small mammals from shooting range sites.

This study advances many areas of ecotoxicology research, especially in the area of comparing field- and laboratory-based assessments in the exposure of wildlife receptors to Pb. Future ecotoxicology studies should attempt to include both lab and field assessments so that the predictive capacity of laboratory tests for field conditions can be better compared and understood. Future studies should continue to develop new ideas and surrogate measures for bioavailability and should validate these methods for ecological receptors.

101

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