## Abstract

The past and future of cities are inextricably linked, a linkage that can be seen clearly in the long-term impacts of urban geochemical legacies. As loci of population as well as the means of employment and industry to support these populations, cities have a long history of co-locating contaminating practices and people, sometimes with negative implications for human health. Working at the intersection between environmental processes, communities, and human health is critical to grapple with environmental legacies and to support healthy, sustainable, and growing urban populations. An emerging area of environmental health research is to understand the impacts of chronic exposures and exposure mixtures—these impacts are poorly studied, yet may pose a significant threat to population health.

Acute exposure to lead (Pb), a powerful neurotoxin to which children are particularly susceptible, has largely been eliminated in the U.S. and other countries through policy-based restrictions on leaded gasoline and lead-based paints. But the legacy of these sources remains in the form of surface soil Pb contamination, a common problem in cities and one that has only recently emerged as a widespread chronic exposure mechanism in cities. Some urban soils are also contaminated with another neurotoxin, mercury (Hg). The greatest human exposure to Hg is through fish consumption, so eating fish caught in urban areas presents risks for toxic Hg exposure. The potential double impact of chronic exposure to these two neurotoxins is pronounced in cities. Overall, there is a paradigmatic shift from reaction to and remediation of acute exposures towards a more nuanced understanding of the dynamic cycling of persistent environmental contaminants with resultant widespread and chronic exposure of inner-city dwellers, leading to chronic toxic illness and disability at substantial human and social cost.

## Introduction

The Earth is increasingly an urban planet. Urbanization is becoming the global norm; the percentage of global population living in urban settings has increased from less than 30% in 1950 to 47% in 2000; the percentage of urban dwellers is expected to increase to 60% by 2025 (WHO, 2010).

Decades of research into human disease have revealed a host of environmental factors influencing human health, both negatively and positively. Cities are at the epicenter of human-environment interactions, marked by high population density, high concentrations of fixed and mobile sources of human-produced or enhanced emissions, high traffic volumes, and frequent occurrence of industrial operations co-located in proximity to human habitation. These factors have resulted in a number of clear case-study examples of negative impacts of mineral-sourced and soil-based contaminants on human health.

Perhaps the most well-studied of these is lead (Pb) poisoning, where the human-produced sources and the environmental cycling of Pb have been reasonably well-constrained (Hamester et al., 1994; Rabinowitz and Wetherill, 1972), and the human health impacts have been well documented (Pirkle et al., 1998). These health impacts extend even to children with relatively low levels of Pb in their blood (Koller et al., 2004). A clear cause-effect relationship for Pb has resulted in substantial and highly effective mitigation actions, although the limitations in current practices focusing on one particular source (degrading Pb-based paints in older homes; Lanphear et al., 1998) have masked somewhat the widespread problem of highly elevated soil Pb in urban areas (Laidlaw and Filippelli, 2008; Mielke and Reagan, 1998; Zahran et al., 2013; Laidlaw et al., 2014). Indeed, as paint-related sources were mostly eliminated in the U.S. 60 years ago, and other probably more harmful sources such as leaded gasoline and lead solder in plumbing were phased out shortly after, the legacy of these sources have been imprinted into the urban fabric in the form of soil contamination. The very soil under urbanites’ feet is now a primary exposure pathway for Pb, creating pockets of poor health and potentially contributing to trends in violent crime in many cities (Wright et al., 2008).

Even after decades of research and action, the incidence of Pb poisoning remains high in urban areas of the U.S., and globally. At particular risk are urban youth from low income families who inhabit the polluted inner neighborhoods of older cities without the benefits of adequate nutrition, education, and access to health care (Filippelli and Laidlaw, 2010). A newer environmental health model is helping us understand this exposure, and to provide several tools to mitigate the harmful impacts of urban Pb. To transition our cities into safer and more health sustainable systems, and to provide environmental justice (that is, equal access to a safe and healthy environment) for a full spectrum of urban dwellers, newer approaches are needed to assess current Pb exposure mechanisms and to fully understand the health implications of chronic Pb exposure—some of this has to revolve around soil geochemistry and legacies of Pb-enriched urban soils. Components of these soils acted as a highly efficient trap of anthropogenic Pb over about 100 years of urban development, and are now returning that Pb to the next generations of people living in cities.

Mercury (Hg) is another trace metal that can affect the health of urban populations, and we will show a number of factors that can increase Hg in an urban environment. Mercury in urban soils can have concentrations substantially higher than concentrations in soils outside cities (Hatcher and Filippelli, 2010). Wet and dry deposition add atmospheric Hg to urban storm runoff in wastewater discharges (e.g., Risch et al., 2010). Streams draining predominantly urban land cover have higher dissolved, particulate, and streambed sediment Hg concentrations than streams with non-urban land cover (Brightbill et al., 2004; Lyons et al., 2006; Risch et al., 2010). When conditions that promote accumulation and magnification of Hg in the food web coincide with high Hg in urban streams, unsafe levels of Hg in fish can occur. As a result, people eating fish caught in these urban waters can have a risk for above-average dietary Hg exposure (e.g. Munthe et al., 2007). Adverse effects on human health from chronic Hg exposure through fish consumption are understood, but enhanced exposures to Pb and Hg in urban populations present the potential for synergistic effects.

We offer a review of Pb and Hg in studies from cities in North America, adding some new research synthesis from Indianapolis, Indiana. This city has a history that includes the use of leaded gas and lead-based paint, along with coal-fired power, incinerators, and metal industries that release Hg to the atmosphere. The findings for this city, in the context of other studies, are likely to be applicable to other urban environments in North America with similar history.

## The long encounter with lead in cities

Lead has been used by humans for thousands of years, and its toxicity has been known for centuries, but it was not until the Industrial Age that this issue became a widespread problem. Lead is a soft and malleable metal easily extracted from galena ore through simple heating. The Romans established a metal-based society early and intensively. Using the newly-conquered Iberian Peninsula, a target of the Romans in large part because of its rich metal ore deposits, the Romans developed the first large-scale quarrying and working operations for Pb, using the finished product in containers, water pipes, and as a sweetener in wines to counteract high tannin levels (Gilfillan, 1965). The environmental legacy of Roman mining in Spain still plagues a number of regions with severe contamination problems. Lead refining and use reduced drastically following the Roman era, with some use in alloys, soldering, glazes, and containers, into the present.

Figure 1.

History of Pb usage in paints and in gasoline (Filippelli et al., 2005), and US geometric mean blood lead levels (BLL; vertical bars; from NHANES, 2008). Lead usage shows the early dominance of Pb-based paints followed by the boom in transportation resulting in a high use of leaded gasoline. The decline after the 1940s in Pb-based paints and after the 1970s in Pb from gasoline points to the environmental controls put in place on both of these sources. The earliest US population subsample for Pb levels in blood occurred in the mid-1970s. In subsequent years, the mean BLL decreased along with the decrease in Pb usage, until sometime in the mid-1980s, when the decline in BLLs flattened, indicating continued sources of Pb from the legacy of Pb usage.

Figure 1.

History of Pb usage in paints and in gasoline (Filippelli et al., 2005), and US geometric mean blood lead levels (BLL; vertical bars; from NHANES, 2008). Lead usage shows the early dominance of Pb-based paints followed by the boom in transportation resulting in a high use of leaded gasoline. The decline after the 1940s in Pb-based paints and after the 1970s in Pb from gasoline points to the environmental controls put in place on both of these sources. The earliest US population subsample for Pb levels in blood occurred in the mid-1970s. In subsequent years, the mean BLL decreased along with the decrease in Pb usage, until sometime in the mid-1980s, when the decline in BLLs flattened, indicating continued sources of Pb from the legacy of Pb usage.

It took the efforts of Clair Patterson, a Caltech geochemist, to bring this hazard to the public’s attention. In the 1950s, Patterson was conducting experiments designed to pinpoint the age of various rocks – but found that his results were skewed by consistent Pb contamination. Further studies showed that Pb levels were elevated in water, soil, even arctic ice – and most troubling, in organisms (e.g., Settle and Patterson, 1980). Over the next three decades, Patterson waged a crusade against Pb that attracted the vociferous opposition of industry groups (as documented in an excellent essay by Bryson, 2003). He ultimately contributed to convincing lawmakers and regulators to outlaw Pb in pipes, solder, and finally in gasoline. As a result of Patterson’s efforts and those of other public health advocates, human-produced sources of Pb in the environment have been significantly reduced (Fig. 1).

### Effects of Pb on humans

Compared to other chemicals of environmental concern, the uptake mechanisms and toxic effects of Pb are relatively well understood. The primary pathway of Pb uptake in humans is via ingestion, where Pb is absorbed in the intestine and incorporated in the body (Manton et al., 2001). Human absorption potential for Pb is dependent mainly on age—the proportion of ingested Pb that is taken up in the body is typically less than 5% for adults whereas it is as high as 50% for children (Roberts et al., 2001). The presence of elevated blood Pb in infants and children leads to permanent neural differentiation defects resulting in lowered Intelligence Quotient (IQ), learning disorders, and attention deficit hyperactivity disorder (Nevin, 2000; Nigg et al., 2008). Because of their high ingestion efficiency and the rapid neural differentiation during early brain and nervous system development, children are especially vulnerable to permanent effects of Pb poisoning. A majority of blood Pb becomes incorporated into bone, which itself becomes a longer-term source of Pb to the biological system—bone is regenerated on timescales of months to years, continually leaking additional Pb into the system (some evidence suggests that elderly suffering from osteoporosis can have elevated blood lead levels from bone-loss related sources, decreasing cognitive function; Needleman, 2004). For this reason, children treated by medical interventions like blood chelation may continue exhibiting toxic levels of Pb in their blood. Furthermore, as neither the placenta nor mammary glands are a perfect barrier to Pb, pregnant and lactating mothers with elevated blood Pb levels may themselves pose a health risk to babies and fetuses.

In some inner-city neighborhoods of Indianapolis, a typical older USA mid-western city, approximately 8% of youth from 1–6 years old exceed the earlier screening standard for the “safe” blood lead level (BLL) of 10 micrograms/dL (Morrison et al., 2012). But with the reduction in this screening level by the U.S. Center for Disease Control and Prevention (U.S. CDC) to 5 micrograms/dL in May 2012 (in response to numerous studies which find significant neurologic and cognitive effects at lower BLLs; Schnaas et al., 2006; Canfield et al., 2007; Lanphear et al., 2005; Chiodo et al., 2007; Jusko et al., 2007; Surkan et al., 2007; Miranda et al., 2007; Nigg et al., 2008), this percentage of Pb-affected children increases to 27% of children in these neighborhoods. Thus, Pb exposure continues to be a public health threat, largely from legacy sources of Pb, the product of a century of Pb use in urban areas. Additionally, most areas with elevated soil Pb also include elevated levels of other metals, such as cadmium (Cd), manganese (Mn), and arsenic (As). Individually, each of these metals poses certain neuorological and developmental risks, but collectively as metal mixtures, their toxic effects may be significantly increased, particularly in utero and in young children (Wright et al., 2006; Hu et al., 2007; Yorifuji et al., 2011). Based on previous work (Filippelli et al., 2005; Laidlaw et al., 2005; Laidlaw and Filippelli, 2008; Laidlaw et al., 2012), soils with elevated Pb and the periodic resuspension of dust particles from these soils plays a major role in Pb exposure to urban children.

The full range of toxic effects of Pb in the human system is still not known, and deserves further study. But the persistent presence of Pb in children is a public health issue of a first order (Karr, 2008). Sources of Pb that could contribute to acute Pb poisoning have been highly publicized in the media, with the focus on consumer product safety (e.g., toys with Pb-based paints) and seriously degraded Pb-based paints in dilapidated homes. The continued source of chronic low levels of Pb to children, however, is not always easy to constrain, and must be assessed by examining environmental loading of Pb from multiple sources.

Lead exposure is listed among the factors that contribute to the global human burden of disease (WHO, 2013). Exposure to Pb is related to a wide range of adverse health effects with high exposures being related to death and serious conditions such as encephalopathy (Needleman, 2004). However, lower, chronic levels also result in impairments of cognition, motor skills, behavior and the immune system and there appears to be no lower threshold below which there are no adverse effects (Binns et al., 2007; Needleman, 2004). The neurobehavioral toxicity caused by Pb places great economic burdens on families and societies. An economic analysis conducted in the United States found the current costs of childhood lead poisoning to be U.S. $43 billion per year. A recent cost–benefit analysis undertaken in the United States found that for every U.S.$1 spent to reduce lead hazards, there is a benefit of U.S. $17–220 (HUD, 1999). This cost–benefit ratio is better than that for vaccines, which have long been described as the single most cost-beneficial medical or public health intervention (WHO, 2010). ### Acute exposure versus chronic exposure Pb-based paints continue to pose acute exposure risks for children. Although paint applied after 1950 was Pb-free, older high-Pb remained in many homes (e.g., Ter Haar and Aronow, 1974). Anybody who has refinished an older home is aware of the problem—what do you do with the Pb paint on the walls, sills, and doorways? The popular way to refinish trimwork and windows is the most problematic. Sanding of Pb-based paints converts the paint from a glue-type solid with limited bioavailability into millions of fine particulates with relatively high Pb content AND very high bioavailability, due to the high surface area/mass ratio of these particles. To confront this problem, many health and environmental agencies at the national, state, and local levels have been waging a campaign of remediation and education about the hazards of Pb. Most of the remediation efforts have been focused on safely removing or covering Pb-based paints in homes—approximately 26% of all US housing stock was built before 1950, and 24 million homes still contain Pb-based paint (HUD, 1999). These remediation efforts continue to this day, with almost$120 million allocated by the U.S. Department of Housing and Urban Development for Pb remediation projects in 2009 alone. The agencies involved have touted these efforts as a success, holding up the improvement in the number of children affected by Pb over the past 25 years as clear evidence. In a national health assessment survey in the late 1970s, 88% of children in the USA had blood-Pb levels above that deemed safe by recent standards; in a follow-up survey in the 1990’s, that number was down to 2.2%, with annual improvements continuing to be seen in interim surveys (NHANES, 2008).

This conventional wisdom, that Pb-based paints still constitute the biggest risk to children with respect to Pb, and that the remediation of Pb-based paint sources has in the past and will continue in the future to provide the chief benefits to children’s health, is firmly entrenched (e.g., Filippelli and Laidlaw, 2010). Even recent litigation in the USA reflects the threat of Pb-based paint, where several high-profile cases brought before juries revolve around large paint producers (Rabin, 2006).

### Urban legacies and soils as a primary chronic exposure source of Pb to children

A practical limit may have been reached in terms of improving the Pb-poisoning outlook for some children, particular those living at or below the poverty level in older cities. Even after decades of active intervention, these urban youth have Pb-poisoning rates that are up to 10 times the national average (Morrison et al., 2012). Some socio-economic risk factors include poor nutrition, pica behavior (a subconscious desire to ingest soil and dust to overcome nutritional deficits), and inadequate pediatric health care (Bernard and McGeehin, 2003; Karr, 2008). Additionally, and of critical importance for improving the health outcome of urban youth, these risk factors also include poor home maintenance with high rental percentages, significant proportion of urban housing with high dust and dirt exposure, and relatively low awareness of the links between health and behavior (Filippelli et al., 2005). This evidence is the key to a new emerging paradigm—namely, that a major source of Pb to children is Pb-enriched soils that are prevalent in cities, especially older ones (Filippelli and Laidlaw, 2010). The source of Pb to the soils includes degraded Pb-based paints, but also Pb deposited from tailpipes, the result of 60 years of combustion of leaded gasoline, and Pb from stationary sources, such as industry. In fact, much of the blame for chronic Pb poisoning and credit for the improvement in the national average of blood-Pb may be the banning of Pb as an additive in gasoline in 1980, effectively preventing substantial amounts of combustion-related Pb deposition from entering the U.S. environment (Mielke, 1994).

The roadway Pb is generally bioavailable, being present in carbonate and oxyhydroxide soil fractions, while the Pb in natural soils is relatively inert. Therefore, dust originating from urban soils contaminated by anthropogenic Pb is more toxic than Pb in naturally occurring dust (Chlopecka et al., 1996; Lee et al., 1997). Lead from the combustion of leaded gasoline is preferentially enriched in the more readily windblown fine size fraction of soils, and so Pb in dusts derived from urban soils is likely to be more potent and concentrated than in the bulk soils (Young et al., 2002).

#### Industrial sources of Pb to soils—the “Ghost Factory” syndrome

Industry has been the backbone of urban growth in the U.S. since the mid-1800s and many neighborhoods were built near industrial sites for the ease of transport for factory workers Indeed, many of these “factory neighborhoods” thrived with the worker’s income supporting markets, restaurants, and myriad retail establishments (USA Today, 2012). This co-location of industry and community had some negative effects, including of course the emission of harmful products into the air, water and soil and subsequent human exposure. This situation is seen clearly with Pb, which was actively processed, recycled and reused in secondary Pb smelting operations. Workers at these operations were exposed to significant amounts of Pb, as were their families when the workers would come home wearing their Pb-laden work clothes. Additionally, off-site air releases of Pb were common—for example, several Superfund sites in the city of Indianapolis are former Pb smelting operations, which contaminated entire neighborhoods with Pb (Morrison et al., 2012). The industrial facilities of many of these contaminating industries are often no longer present, and the only way to know where these “Ghost Factories” (USA Today, 2012) were located, or the type of facilities they were, is through old property records. They have also been found when performing soil or water testing (USA Today, 2012). As many of these sites were not properly “closed” after operations ceased, they continue to be sources of fugitive dust contamination (Morrison et al., 2012).

#### Diffuse soil Pb and children’s health

The original sources of Pb to the environment were directly tied to the spatial characteristics of the product itself, with Pb-based paints linked to pre-1950 structures, gasoline-related Pb linked to roadways and traffic volume, and Pb emitted from smelters linked to the smelter location, stack height, and wind direction. Lead does not deposit far from its source, and its geochemical characteristics promote rapid sequestration onto surface soil particulates (usually via surface complexation of Pb and Pb oxides with soil organic matter). An analysis of many urban areas reveals that these point sources have, to some extent at least, been redistributed to produce regions of Pb enrichment. Several factors can lead to redistribution of Pb-enriched particles and soil, but the recurrence of a general urban enrichment of soil Pb has been documented in many regions, and is termed diffuse soil Pb.

One of the characteristics of Pb distribution in surface soils of several older cities is a distinct decreasing trend from city center to suburban surroundings, a legacy both of Pb deposition, redistribution and smearing of original point sources, and less Pb deposition in newer suburban neighborhoods due to recent Pb controls (e.g., Mielke et al., 1984; Filippelli et al., 2005; Johnson and Bretsch, 2002; Roux and Marra, 2007). The urban roadway example shows both the impact of the point source of Pb deposition from leaded gasoline as well as the diffuse soil Pb that blankets urban regions. In other words, even at distances away from the roadway beyond where direct Pb deposition occurs (and far away from structures using Pb-based paint), the background level for Pb is significantly higher in urban areas (∼500 ppm) than in suburban areas (∼ 60 ppm; Laidlaw and Filippelli, 2008).

Our recent work in Indianapolis, Indiana (USA) illustrated many of these issues. Urban areas far from proximal Pb sources (i.e., house-side paint or roadway gasoline) typically had soil Pb concentrations below 500 ppm (Laidlaw and Filippelli, 2008), whereas roadway and house-side soil sampling revealed Pb concentrations well above 1000 ppm (Fig. 2). The lowest Pb concentrations averaged about 50 ppm, which is a typical value for soils in this region. As expected, the highest soil Pb concentrations were focused in a bulls-eye pattern directly over the old urban areas of Indianapolis, where the diffuse soil Pb content averaged over 200 ppm (Filippelli et al., 2005). Beyond this central hot spot, Pb concentrations decreased systematically toward the suburban outskirts of the city, ultimately falling to background values in the rural fringes of the city. The central peak is consistent with the long history of Pb use in the downtown, but the generally high values even away from the urban core supports the argument of a redistribution of Pb over time. This pattern of soil Pb is a common feature of the spatial distribution of urban Pb, and is likely related to the wind-driven redistribution of fine Pb-enriched particulates over decades. A “plume” of deposition doesn’t exist that can be ascribed to the northwestward prevailing winds in Indianapolis, likely because, unlike releases of particulates at higher elevations (i.e., smokestacks), wind direction has little influence on Pb-dust depositional patterns in the turbulent near-surface environment of a cityscape (Laidlaw et al., 2005).

Figure 2.
Soil lead levels in Indianapolis, Indiana, USA.

Diffuse soil Pb (measured more than 40 m from roadways and structures) in Indianapolis, Indiana (lower left) and detailed blow-up of higher resolution yard-scale sampling in a neighborhood with high incidence of elevated blood Pb levels (upper right). The concentration of diffuse soil Pb in surface soils of Indianapolis (in colored regions) displays a characteristic pattern of urban enrichment trending toward lower values in suburban and agricultural regions. Note the high soil Pb values on the south portion of the neighborhood map, proximal to a shuttered Pb smelting plant.

Figure 2.
Soil lead levels in Indianapolis, Indiana, USA.

Diffuse soil Pb (measured more than 40 m from roadways and structures) in Indianapolis, Indiana (lower left) and detailed blow-up of higher resolution yard-scale sampling in a neighborhood with high incidence of elevated blood Pb levels (upper right). The concentration of diffuse soil Pb in surface soils of Indianapolis (in colored regions) displays a characteristic pattern of urban enrichment trending toward lower values in suburban and agricultural regions. Note the high soil Pb values on the south portion of the neighborhood map, proximal to a shuttered Pb smelting plant.

One example of a scaled-down approach to the study of urban Pb is a neighborhood-scale analysis of soil Pb, blood Pb levels in children, industrial sources, and socio-economic patterns (Filippelli et al., 2012). In this study, a detailed soil Pb survey was performed in a neighborhood on the west side of Indianapolis. Several interesting findings arose from this analysis. First, significant patterns existed in soil Pb concentrations (Fig. 3), even though the housing stock was of the same general age (i.e., pre-1920’s) across the neighborhood and thus the influence of lead-based paint in homes could be considered uniform. Second, the area proximal to a large Pb smelting operation on the south side of the neighborhood had relatively low soil Pb concentrations, due to remediation activities in 1994 involving the removal of surface soils that were contaminated with Pb from yards of these homes (Figs. 2, 3). Third, the average Pb concentration in this remediated area on the south side of the neighborhood was 115 ppm approximately 17 years after clean soil with Pb concentrations less than 50 ppm was imported, indicating that a continued source of elevated Pb dust influenced the neighborhood soils. Fourth, average blood Pb levels varied across the neighborhood; this variation was not well predicted by soil Pb concentrations, and not explainable by socio-economic status as this entire neighborhood was uniformly at or below the family poverty level. The strongest predictor of blood Pb level in this neighborhood was race, with high relative proportions of African American children coinciding with the areas of high blood Pb levels (Morrison et al., 2012).

Figure 3.
Soil lead and blood lead levels on the neighborhood scale in Indianapolis, Indiana, USA.

Mis-matches between soil Pb (here, shown in ppm Pb noting that the background value for soils is 30 ppm) and BLL (here aggregated at the census block level and ranging from <2 microgram/dL in dark green to >6 microgram/dL in dark red). This example is from the NearWest community of Indianapolis, showing pockets of low BLL even in the face of high soil Pb.

Figure 3.
Soil lead and blood lead levels on the neighborhood scale in Indianapolis, Indiana, USA.

Mis-matches between soil Pb (here, shown in ppm Pb noting that the background value for soils is 30 ppm) and BLL (here aggregated at the census block level and ranging from <2 microgram/dL in dark green to >6 microgram/dL in dark red). This example is from the NearWest community of Indianapolis, showing pockets of low BLL even in the face of high soil Pb.

The interplay between environmental burden and human health in the case of Pb is critical. On a city-wide scale in Indianapolis, Indiana, average blood Pb levels in children display highest values near the city core, and decreasing values away from the urban center (Fig. 4). This pattern is largely related to the distribution of elevated background soil Pb levels in the city (Filippelli et al., 2005), although some modifications in the relationship are observed, such as the lack of elevated blood Pb levels in the central city core even in the face of elevated soil Pb there. This mismatch is driven by the lack of home addresses in the city core, which is dominated by commercial buildings and corporate campuses. The general relationship between elevated soil Pb and elevated blood Pb levels in children is driven by the strong exposure link between Pb-rich soils, periodic atmospheric transport of Pb-rich dust generated from those soils, and uptake in children via behavior (hand-to-mouth behavior, crawling, etc.) and subsequent inadvertent ingestion (e.g., Filippelli and Laidlaw, 2010). To ascertain the major risk factors involved beyond the environmental ones of soil Pb concentration and dust generated from soils, the analysis needs to be reduced to smaller spatial scales with more meaningful population differences.

Figure 4.
Blood lead levels of children in Indianapolis, Indiana, USA.

Average blood lead levels by US Census Block for the period February 2002–December 2008 (n=12,431) for children between the ages of 0 and 5 years old in Indianapolis, Indiana (area = 1,044 km2). BLL = blood lead level, given in micrograms per deciliter. (From Filippelli et al., 2012).

Figure 4.
Blood lead levels of children in Indianapolis, Indiana, USA.

Average blood lead levels by US Census Block for the period February 2002–December 2008 (n=12,431) for children between the ages of 0 and 5 years old in Indianapolis, Indiana (area = 1,044 km2). BLL = blood lead level, given in micrograms per deciliter. (From Filippelli et al., 2012).

In summary, a complicated pathway leads from legacy Pb sources to Pb exposure impacts on human health in cities. Tracking this pathway is critical for reducing the Pb burden in cities, and thus also enhancing the future health, intellectual development, and economic welfare of urban dwellers in Indianapolis and many other similar cities across the globe.

### New views of lead mitigation approaches in cities

A new paradigm of urban Pb loading is emerging, one that helps to explain continued chronic Pb poisoning and seasonal patterns in blood Pb levels of children. Unlike discrete point sources such as Pb paint and industrial contact, which are still responsible for most cases of acute Pb poisoning, diffuse soil Pb is the main avenue for urban Pb loading of children. The diffuse soil Pb comes from several sources, including leaded gasoline and degraded Pb-based paints, but in a sense the source no longer matters—because of the ability of surface soils to retain Pb, these soils themselves have become the new risk factor for children’s health in Pb-contaminated cities.

Widespread contamination of urban soils creates a different challenge for mitigation of Pb risks for children, one based on removing surface soils from human contact. Most mitigation efforts for heavily-contaminated soils have involved soil removal and replacement, a disruptive and expensive option for controlling Pb sources in urban areas. Another approach was tested which was simply to cover the contaminated yard soils with about 15 cm of Pb-free soil, which in the case of New Orleans came from the nearby Mississippi levee (Mielke et al., 2006). At a fraction of the soil removal cost, this clean soil is simply graded over the old soil layer, hydroseeded, and left to grow a lawn. This approach caps the Pb-contaminated soils, removing them from contact by children. The result of this approach has been a substantial reduction in the blood Pb levels of children living in the affected homes (Mielke et al., 2006). Zahran et al. (2010) report on how nature natural processes did this same experiment, seeing substantially lower blood Pb levels for New Orleans children after Hurricane Katrina, due to the capping of much of the Pb-contaminated soil with flood-related sediments. Mielke et al. (2006) observed that, over the course of several months after treatment, soil Pb levels in the treated sites began increasing. This increase was due to dust generated from soils from adjacent, untreated yards and neighborhoods that still had high soil Pb concentrations. This finding agrees with results from an urban gardening study in Boston (Clark et al., 2008), which revealed that raised beds experienced substantial increases in soil Pb values over as little as four years after bed construction, indicating the need to control dust-transported Pb at the neighborhood scale. Collectively, these findings not only confirm the new paradigm of diffuse soil Pb as a culprit in urban areas, but also indicate that a comprehensive treatment approach is required to provide a long-term benefit.

### Mercury in the urban environment—identifying sources, reducing exposures

#### Atmospheric Hg deposition in urban areas

The National Atmospheric Deposition Program Mercury Deposition Network (NADP-MDN) is the largest, long-term monitoring system for Hg wet deposition in North America (National Atmospheric Deposition Program, 2015). The network has included a small number of urban sites, including a site in Indianapolis which, for this paper, is a source of data for comparison with Hg in soils, streambed sediment, stream water, and fish. Mercury emissions were approximately 160 kg/yr within 50 kilometers of the Indianapolis NADP-MDN site, according to inventory data such as the Regional Air Pollutant Inventory Development System (Risch and Fowler, 2008). The largest emissions sources included a coal-fired power plant, a waste incinerator, a wastewater sludge facility, and a foundry, all located in the south-southwestern part of the city, which has prevailing winds from the south-southwest (Risch and Fowler, 2008). NADP-MDN data for the Indianapolis site had average weekly Hg wet deposition among the highest in the Great Lakes Region for 2002-2010 (Risch et al., 2014). An analysis of spatial patterns and temporal trends in Hg wet deposition in the Great Lakes region by Risch et al. (2012) indicated a substantial localized net increase in Hg concentrations and deposition in central Indiana near Indianapolis, 2002-2008. On a larger scale, a synthesis of national Hg wet deposition monitoring data and related studies by Butler et al. (2007) identified examples of substantial Hg deposition in urban areas while noting the limited number of urban Hg monitoring locations, and Keeler et al. (2006) demonstrated that local emissions sources contribute to local Hg deposition, thus reinforcing the importance of local components in the regional Hg cycle.

#### Hg in urban soils and streambed sediments

A case study by Hatcher and Filippelli (2010) indicated a strong relationship between soil Hg concentrations and stream bed sediment Hg and a potential link between soil Hg distribution and anthropogenic sources. In this study, the unique relationship between several emission sources in the SW part of Indianapolis, a predominant wind direction from southwest to northeast, and a stream flow pattern from the NE to the SW revealed internal cycling of Hg in this setting. Plumes from local emissions sources delivered wet and dry Hg deposition to soils downwind, while particulate Hg in stream water from upstream watersheds delivered Hg to bed sediments in the same area (Hatcher and Filippelli, 2010). Risch et al. (2010) found that unfiltered water Hg concentrations in streams were related most to levels of particulates in the water. These findings suggest that soil and its attached Hg, if it is eroded and transported to a stream, will increase the stream Hg concentrations.

The White River in central Indiana flows from the rural/suburban northeast, through the urban core of Indianapolis to the rural/suburban southwest. Streambed sediment samples from individual impoundment stretches of the White River were examined spatially with respect to mean total Hg, organic matter content, and Hg normalized to percent organic matter (Hatcher and Filippelli, 2010). The normalized Hg calculation was designed to control for that aspect of Hg content associated with variations in dilution by terrigenous material, which presumably has little Hg associated with it compared to organic matter. A correlation did indeed exist for total Hg and organic matter as a function of distance from stream edge (Hatcher and Filippelli, 2010), and thus normalizing total Hg to organic matter content does “stabilize” the Hg record at given sampling sites. Also, the stream bank sediments nearest the waterline are typically wetter and thus have a higher potential to contain reducing sub-environments. Given the combination of high organic matter, high total Hg, and low free oxygen, these are exactly the environments where methylation of Hg is most likely to occur (e.g., Evers, 2007; Skyllberg et al., 2003).

The mean total Hg content ranged from a low of 6 ppb in the far northeastern portion of the study area to a high of 830 ppb in the urban core (Hatcher and Filippelli, 2010). But the most striking aspect of the downstream record of Hg in the White River is seen in the discrete sample record (Fig. 6). The rural upstream stretches exhibit Hg concentrations between 3 and 45 ppb, a range that persists as the White River flows into the northern part of the urban core of Indianapolis over several impoundments. The Hg values increase markedly, however, in the southern part of the urban core, marked by the input of several tributaries and proximal to many of the listed Hg emission sources in the SW portion of the city, reaching values several-fold higher than incoming sediments (Fig. 6) and indicating significant additions of Hg in this region. These elevated Hg values persist for tens of kilometers downstream and south of the urban/industrial core of Indianapolis, into rural stretches of the White River.

The urban core of Indianapolis certainly sees high Hg values in streambed sediments, and unfortunately also sees a large number of low income urban anglers who might note the fish consumption advisories posted along the White River but do not always comply with them, based on extensive angler surveys conducted by us previously (Hatcher, 2009). That the high Hg values persist well downstream of the urban core into rural lands is also of concern given the natural human tendency to equate the local surrounding environments with the health of the air and water (and in this case, the fish). The identification of this urban pollution memory for streams is not new (e.g., Gray, 2000; Neumann et al., 2005), but it might be especially problematic for Hg. Many conditions need be present to cause high methylation rates, but the combination of high Hg, high organic matter, and numerous impoundments all contribute to a higher potential for MeHg production (Mason et al., 1994; Macalady, 2000; Seigneur et al., 2004; Sorensen et al., 2005) and thus a higher risk to anglers if they consume certain fish from even these rural stretches of the river (Munthe et al., 2007; Scudder et al., 2009).

Crewe (2012), utilized a broad soil sampling scheme to characterize Hg concentrations throughout central Indiana and to further explore the link between regional patterns and a cluster of large Hg emissions sources in southwestern Indianapolis. Over 100 surface soil samples were collected, utilizing a spatial grid pattern with high density in Indianapolis and lower density in outlying regions (Crewe, 2012). Results revealed significantly elevated soil Hg concentrations in and slightly to the NE of Indianapolis, and lower concentrations in the suburban and rural regions surrounding the city (Figs. 7, 8). The Hg concentration was not influenced by soil type or organic matter content. Background Hg values for this analysis were about 30 ppb, consistent with results for a study of Illinois soils indicating background Hg values of 20 ± 9 ppb (Dreher and Follmer, 2004).

Fig. 7.
Soil mercury concentrations in different land use types from central Indiana, USA.

Average and 1 sigma standard deviations for soil Hg concentrations in central Indiana from a range of land use classifications.

Fig. 7.
Soil mercury concentrations in different land use types from central Indiana, USA.

Average and 1 sigma standard deviations for soil Hg concentrations in central Indiana from a range of land use classifications.

Fig. 8.
Soil mercury and stream sediment mercury concentrations from Indianapolis, Indiana, USA.

Map of Marion County, Indiana with surface soil Hg concentrations and stream Hg concentrations from the White River. The soil Hg peaks coincide with an area just downwind (dominant wind direction is to the northeast) of multiple Hg emission sources, including the largest Hg point source, the Harding Street Station coal-fired electric utility plant (noted by triangle), which emits approximately 62 kg of Hg/yr. The stream sediment values indicate higher concentrations downstream of the soil Hg peak regions, likely the result of funneling of Hg from the various subwatersheds that transport water and sediments to the south /southwest of the county (e.g., Hatcher and Filippelli, 2010).

Fig. 8.
Soil mercury and stream sediment mercury concentrations from Indianapolis, Indiana, USA.

Map of Marion County, Indiana with surface soil Hg concentrations and stream Hg concentrations from the White River. The soil Hg peaks coincide with an area just downwind (dominant wind direction is to the northeast) of multiple Hg emission sources, including the largest Hg point source, the Harding Street Station coal-fired electric utility plant (noted by triangle), which emits approximately 62 kg of Hg/yr. The stream sediment values indicate higher concentrations downstream of the soil Hg peak regions, likely the result of funneling of Hg from the various subwatersheds that transport water and sediments to the south /southwest of the county (e.g., Hatcher and Filippelli, 2010).

The spatial pattern indicates that local sources of Hg dominate the soil Hg content in this region because they are situated directly “upwind” of the apparent plume in soil Hg (Fig. 8). The significance of this relationship is two-fold. First, it reveals the potential regional impact of Hg emissions sources including coal-fired power plants on Hg geochemical cycles, and related issues with streams and fisheries. Second, it provides some measurable indicators of the distribution of wet and dry deposition intensity in the urban landscape. It is expected that under the newer EPA regulations that aim to reduce Hg emissions (Harris et al., 2007), regional Hg sources and thus deposition will decline, leading eventually to reduced Hg in watersheds and in streams. Indeed, these improvements may be reached far sooner in this particular case from Indianapolis. The operator of this coal-fired power plant has filed a request to convert the entire facility to natural gas by the end of 2016, thus effectively and immediately removing nearly all Hg from the emissions stream.

To determine how quickly this improvement will occur, the Net Atmospheric Deposition (NAD) of the urban area was calculated from the soil Hg concentration results:

$NAD=(MU−MR)(SD)(AU)(PD)$
1

Where MU and MR represent the means of urban and rural soil Hg concentrations (g Hg/kg soil), respectively, SD is soil density (1.3 g/cm3), AU represents the area of urban land (1.1 x 1011 cm2), and PD represents the penetration depth of anthropogenic Hg in cm. In this scenario, MR is assumed to be the regional background for soil Hg, and PD is assumed to be 5 cm (our cumulative soil sampling depth). The calculated NAD is 1020 kg of additional Hg in urban soils. Using emissions of 62 kg/yr from the largest Hg emitter in the region (a coal-fired power plant) and steady state relationships between the urban soils and the input flux, we derive a residence time of Hg in surface soils of 16.5 years. In this model, Soil Hg input would be solely via published local atmospheric emission sources and soil Hg loss would occur via transport to deeper soil layers, runoff, and/or volatilization. Clearly, other sources and other loss functions occur in reality, and our estimates are constrained by the simplifications above, and thus uncertainties are significant. Hissler and Probst (2006) estimated the residence time in the top 0–60 cm of soil to be approximately 70 years, and thus our rough estimate using surface soils only might be reasonably representative. If so, we may see a rapid and measurable reduction in soil Hg values in the next decade from this shift in policy and practice from coal burning to alternative forms of fuel. If other monitoring data from environmental compartments (water, sediment, and fish) and relevant health screening are available, this change may be found to result in improvements in environmental health because of a reduction in Hg from local fisheries.

#### Mercury in urban streams and fish

The average annual Hg concentrations and stream yields for the watersheds of the Indianapolis area were the highest in the state and were higher than the Hg wet deposition loading, indicating some Hg loading from wastewater and stormwater outfalls. Most watersheds in the state had watershed Hg yields that were less than the wet deposition loading (Risch et al., 2010). When statewide fish Hg concentration distributions were assessed, these same watersheds in the Indianapolis area had among the highest percentages (up to 40%) of fish Hg concentrations that exceeded the U.S. Environmental Protection Agency criterion for Hg of 300 ppb (U.S. Environmental Protection Agency, 2001).

#### Human Hg exposure pathway in urban environments

The current paradigm for the human Hg exposure pathway in urban environments (Fig. 6) includes:

• Atmospheric deposition (wet and dry) of Hg goes to urban soils and other surfaces, with little directly to the streams.

• Some Hg is in streams from overland runoff with some Hg attached to soil particles in the runoff.

• Some Hg in dry atmospheric deposition on soil particles and urban surfaces is transported with Hg wet deposition and precipitation runoff to a storm sewer or combined sewer outfall, ending up in streams.

• Several of those paths take Hg to a stream where some Hg stays in the water, dissolved or attached to particles. Some Hg settles into streambed sediment.

• The Hg in anoxic streambed sediments and riparian wetland soils can be converted to methylmercury (MeHg) by microbial action in the presence of organic carbon and sulfate.

• MeHg enters the stream water and then enters the food web, moving from phytoplankton and zooplankton to primary consumers and eventually to secondary consumers and top level predator fish.

• The highest MeHg accumulation is in the long-lived (and largest) secondary consumers and top level piscivorous fish.

A straightforward approach to eliminating Hg exposure is to eliminate consumption of fish that may potentially have elevated Hg, particularly for children and pregnant and nursing women. Of course, removing a valuable protein source from a population that might not have practical alternatives poses its own nutritional and health concerns. Another approach is to reduce the anthropogenic sources contributing to environmental Hg in the first place, an approach embodied in the EPA MATS regulations and occurring through both intentional redesign and modification of existing major emission sources and the economically- and environmentally-driven shift from coal to natural gas as a fuel stock in utility plants. The former approach of reducing/eliminating consumption will have immediate positive effects to individuals, whereas the latter environmental source approach will have positive effects for broad populations, but of indeterminate timeframe for those effects to make it from source to food web to human.

## Lead and mercury together: Synergistic, antagonistic, or agnostic?

Much is known about the neurological impacts of Pb and Hg independently, but little is known about their combined effects. Their exposure sources to people, although both anthropogenic in nature, are quite different in practice. And whereas Pb is stored in the body for years and even decades, Hg is rapidly eliminated from the body in a matter of months. But even with these exposure route and biochemical differences in the body, the growing body of literature indicates that some populations might be at particular risk for being exposed to both neurotoxins at unsafe levels. One particularly vulnerable population would be urban poor who utilize fishing as a cultural or economic practice. These populations have a higher general risk of having elevated blood Pb levels given their environment. And if pregnant women or young children eat fish caught in these environments, they almost certainly will be exposed to a significant dietary source of Hg. It would be enlightening to examine Pb and Hg co-exposure and co-morbidity in some sample populations to understand (1) if this combination of neurotoxin exposure is manifested in this potentially vulnerable group, and (2) if there seems to be synergistic, antagonistic, or agnostic relationships between Pb and Hg in terms of neurocognitive function. Understanding how real-world environmental mixtures impact human health is one of the next big steps in toxicology, and now that we understand the environmental factors well enough to define study population groups, we may indeed be in a position to answer some of these questions.

## Conclusions

Although the two neurotoxins focused on here, Pb and Hg, have markedly different cycles, soil concentrations of both are highly elevated in the Indianapolis urban area at least, and likely in others. With a shift in focus from acute exposures to chronic exposures, our approach to analyzing these toxins and mitigating the negative human impacts from them should continue to shift as well. For example, soil and dust generated from soil is now seen as an important component for human Pb exposure, and thus actions to reduce exposure to these sources have to take geochemical cycling into account. Additionally, the focus on reducing acute Hg exposure has rightly been on understanding Hg bioaccumulation processes in the fisheries realm, but another important question might be whether chronic low-level exposure of urban populations to Hg is also coming through the soil-dust-ingestion process. Lacking appropriate broad scale population data on chronic Hg levels in humans, this question remains unanswerable, even as we are increasingly aware of the deleterious impacts of exposure mixtures on human health.

This survey has focused almost solely on examples from the U.S. Although geochemical processes are universal, the exposure processes and mitigation approaches to reduce exposure vary widely on a global basis. For example, chronic urban Pb exposure is likely typical of all older cities in the world that have utilized Pb-based paints and leaded gasolines, which some countries still do, but acute exposure also manifests in rural areas from unsafe mining and processing practices. Thus, the whole world has not yet shifted from acute to chronic exposures as the main health concern, and indeed violence, malnutrition, and avoidable diseases are a top health concern for much of the planet’s population. Nevertheless, understanding urban geochemical cycling is one path towards enhancing the health and sustainability of cities, and will play a larger role in environmental health fields than it currently does.

## Data accessibility statement

The data supporting the findings and interpretations presented here are available in source publications by the authors as well as through the United States Geological Survey.

© 2015 Filippelli et al. This is an open-access article distributed under the terms of the Creative Commons Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the original author and source are credited.

## Acknowledgments

We are grateful for the support and feedback of valuable colleagues, including Howard Mielke, Mark Taylor, Sammy Zahran, Sarah Wiehe, Tamara Leech, W. Berry Lyons, and David Long, as well as the feedback of the reviewers of this manuscript. This work was supported by the Center for Urban Health at IUPUI.

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This research was funded by the Indianapolis Foundation, the National Institutes of Health through a Clinical Translational Sciences Institute grant (NIH/NCRR Grant number RR025761), a CTSI Pre-Doctoral Fellowship, and the Center for Urban Health at IUPUI through an Urban Health Fellowship.

## Competing Interests

The authors have declared that no competing interests exist.

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