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.
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.
Lead has been added to paint for centuries—distinctive colors are achieved with the addition of metals to paints. In the USA, a boom in residential housing development in the early part of the 20th century resulted in national-scale advertising blitz for “white lead paint” and the application of Pb-based paints in millions of new homes (Fig. 1). The addition of Pb, in practice up to 15% by weight, enhanced durability and flexibility of paints. In the early part of the 20th century, most single- and multi-family dwelling units had Pb-based paints in their walls, window sashes, and doorways. Even brick and stone houses employed Pb-based paints in window frames and doors. Although Pb enhanced durability, paint has its functional limits, and the degradation around friction points (doorways, window sashes) combined with the exploratory nature of young children, and malnutrition and poverty, resulted in the first widespread tragedy of Pb. As children were being admitted to hospitals with symptoms of severe Pb poisoning, the link to Pb-based paints became apparent. In the late 1940s, pressure from the health profession and consumer advocate groups succeeded in legislation prohibiting the addition of Pb to house paints. Although still allowed in the U.S. to this day in industrial applications such as use in bridge paints, with potential implications for train trestle bridges in urban settings (Weiss et al., 2006), the phase-out of Pb in house paints beginning in the 1950s gave hope for a Pb-free future for children. Because of the rapid, continuing, and pervasive growth in the use of automobiles, a growth fueled with leaded gasoline, and the legacy of degrading Pb-based paints on homes, this future has been a long time coming. Indeed, Pb additives for gasoline were developed as an anti-knock engine formula in the 1920s, and the proliferation of motor vehicles in the middle part of the century was fueled by gasoline enriched with tetra-ethyl Pb (Fig. 1; Mielke, 1994).
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).
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).
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).
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).
In the U.S. alone, about 5 million metric tons of Pb was deposited in the environment as a result of the combustion of leaded gasoline (Fig. 1). Almost all of that Pb was originally deposited very close to roadways, with aerosolized combustion products containing Pb initially deposited within about 50 m of a roadway if no obstructions were present. The fate of deposited Pb then depended on the conditions of the depositional area. Although intersections of busy streets may have received over 1 metric ton of Pb per year (Mielke, 1994), their impervious surfaces led to continual runoff of Pb-enriched particulates down storm drains, and from there directly into rivers. If the particulate Pb was deposited instead on a grassy fringe, like a front yard or park, the Pb was effectively retained. In such a setting, the insolubility of Pb leads to its concentration in the surface layer of soils, which can reach levels above 1000 parts per million (ppm) in cities (e.g., Filippelli et al., 2005). Thus, surface soils became the repositories of Pb deposited over decades—in the case of older roadways, the proximal soils retained almost all of the Pb deposited on them over a period of about 60 years.
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).
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).
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).
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).
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.
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.
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.
The strategy of mitigating Pb risks in entire cities is of course different than doing such at a single contamination site or mine. In many ways, it seems more daunting because of the scale, but it is perhaps more feasible given a newer understanding of urban Pb exposure sources. As indicated above, surface capping of Pb-contaminated soils is effective, especially if done on a neighborhood rather than individual property scale. Mielke et al. (2006) calculated that the entire Pb-affected area of New Orleans could be remediated by capping for a total cost of less than $US 300M, compared to the $76M annual cost to New Orleans of Pb poisoning. Simply put, the national cost of doing nothing about soil Pb is staggering—Gould (2009) calculates that Pb poisoning costs the nation $US 30–146M in special education, $US 267M in attention deficit-hyperactivity disorder, and $US 1.7B in the direct cost of crime and recidivism. The cost:benefit ratio of Pb reduction is favorable from a financial standpoint, but the necessity of reducing Pb burdens to urban youth goes far beyond dollars and cents, and continues to be a critical component of enhancing public health.
An interesting phenomenon has occurred in many larger U.S. cities over the past decade—an explosion of urban agriculture and a new awareness of sustainable urban food systems. This movement likely has several contributing causes, including newer availability of large tracts of property in cities after the global recession of 2008–2010, a shift in public perception of cities as desirable places to live and work, and the influx of creative, innovative, and sustainability-focused young people, and their energy and resources, to cities. For example, the number of registered urban farms has increased from 20 to 110 in Indianapolis over the past five years, a trend which shows no signs of slowing down.
Growing food in cities and distributing that food locally has a number of benefits, including enhanced access to fresh and nutritious food, employment of local farmers and distributors, reuse of otherwise vacant land, and a generally decreased carbon footprint of food. But urban soils have environmental legacies, not least of which is soil Pb (e.g., Clark et al., 2008). Given its geochemistry and past sources, Pb is most enriched in surface soils, and indeed those soils acted to concentrate up to a century of Pb deposition in the surface 20 centimeters of soil, exactly where gardeners work and where plants grow (Oka et al., 2014).
In an effort to both inform the public and to provide opportunities for citizen scientists, we launched the Safe Urban Gardening Initiative in Indianapolis, assisted with funding from the Indianapolis Foundation. This initiative called on citizens to collect samples of soils from several locations in their yards (under the roof dripline, near a roadway, in the garden or potential garden sites) and deliver these samples to the laboratory at IUPUI for geochemical analysis for Pb, other anthropogenic metals, and organic matter. The citizens were provided with instructions, and sometimes sampling kits. Data were then provided back to the citizens with recommendations on remediation based on the levels of Pb that were found, and a guide to safe urban gardening (Fig. 5; Filippelli, 2012). The citizens received data and solutions, and we received geolocated samples from a broad expanse of neighborhoods. Although ongoing, with several potential intriguing findings about the fine-scale nature of metals sources and behavior, we have analyzed over 2,000 samples and have provided that data back to citizens. This is not the only such program—Dr. Mark Tayler runs a VegSafe program out of Macquarie University in Australia and Dr. Howard Mielke runs a similar program out of Tulane University in the USA—and we hope that by engaging community members in the process, we are creating a more scientifically-informed populace, with the necessary tools to take action on various issues related to environmental justice.
Mercury is another neurotoxic heavy metal with different sources than Pb, and comprehensive regulatory control of anthropogenic Hg emissions to the air is a very recent development. Air emissions from fixed sources such as coal-fired power plants and incinerators are regulated in the U.S., but human activities remain the most significant sources of atmospheric Hg loading to aquatic and terrestrial ecosystems in most places. The environmental cycle of Hg includes air, water, sediment, soil, and biota. It is important to understand the components of this cycle because past and current global Hg pollution impacts human health (Park and Zheng, 2012). Deleterious impacts of Hg include permanent effects on the development of the brain and nervous systems of fetuses and children (Mergler et al., 2007), on human endocrine systems (Tan et al., 2009), and on the nervous and cardiovascular systems of adults (Hightower, 2009).
Mercury is routinely measured in biologically-relevant concentrations in the urban environment. Mercury in urban soils can have concentrations substantially higher than ambient background concentrations in soils outside cities (Crewe, 2012). Man-made surfaces such as buildings and pavement in cities enhance the dry deposition of atmospheric Hg that can enter storm runoff and cities that have large numbers of outfalls for wastewater effluent can have high levels of stream Hg (Risch et al., 2010). Streams draining predominantly urban land cover have higher streambed Hg concentrations (Wentz et al., 2014, Hatcher and Filippelli, 2010). Lakebed sediments in locations draining urban land cover have higher Hg concentrations than rural areas (Van Metre, 2012; Engstrom et al., 2007). Streams draining predominantly urban land cover have higher concentrations, loads, and yields of stream Hg than non-urban streams (Brightbill et al., 2004; Lyons et al., 2006; Risch et al., 2010). Where environmental conditions promote the conversion of the large urban Hg loads to methylmercury, food web accumulation will occur. As a result, unsafe levels of Hg in fish can occur in urban streams with high Hg concentrations in the water and streambed sediment, which means that individuals eating fish caught in urban waters risk above-average dietary Hg exposure.
The contributions to atmospheric Hg deposition from local, regional, and global emissions sources is different than that for Pb, which is dominated by local sources. Model-based attributions of Hg deposited at a specific location from local-regional Hg emissions versus the global pool and international sources will vary depending on the location. Urban locations with a greater number of nearby Hg emission sources will be affected differently than remote locations with more distant sources. Grant et al. (2014) presented a model simulation for 2005 that suggested strong localized inputs to wet and dry Hg deposition in the southern Great Lakes region, closer to emissions sources. Cohen et al. (2004, 2007) reported similar findings, with regional sources in the Ohio River valley contributing Hg deposited to the Great Lakes. Pirrone and Mason (2009) reported that Hg wet deposition in urban areas can be three to five times higher than surrounding areas, with total deposition fluxes up to 10 times greater in the immediate vicinity of a local source and up to five times greater than background levels. Expected reductions in Hg emissions from local and regional sources may be offset by increasing global Hg emissions and transport, especially from Asia (Driscoll et al., 2013).
Calls for controls on Hg sources to the environment have resulted in international and national policies. The Minamata Convention on Mercury is a global treaty to protect human health and the environment from the adverse effects of mercury, and has been signed or endorsed by well over 100 countries. The U.S. EPA Mercury and Air Toxics Standards (MATS) rule phases in limitations to Hg emission from power plants and other stationary sources beginning in 2015, although it has recently been challenged in court. Record low prices of natural gas in recent years have also motivated some power companies to replace coal with natural gas as the fuel at younger plants and to retire old plants rather than add the maximum achievable technology for reducing Hg emissions, such as activated carbon and a combination of flue gas desulfurization, selective catalytic reduction, and electrostatic precipitators or bag houses. Risch et al. (2014) reported 11% of the coal-fired capacity will be converted during 2010-2019 in the Great Lakes states. Many current power plants are refitting equipment to comply with these newer standards, at significant costs to local electricity producers and consumers. To date, however, there is still a need to identify how much current Hg emissions impacts local environments; little actionable information exists about to what extent, and how quickly, the benefits of reducing emissions will be realized by the communities downwind of coal-fired power plants. Without such information, it is difficult to balance power producers’ arguments of increased electricity costs to local consumers from compliance versus counter-arguments of environmental and human health savings from compliance.
Urban environments may include Hg emissions to the air from a variety of human activities. These sources include coal-fired power plants, waste incinerators, steel mills and foundries, cement kilns, utility and industrial boilers, chlor-alkali plants, brick refractories, refineries, landfills, and asphalt plants. Some sources have tall stacks, whereas others have shorter stacks. A discussion of coal-fired power plant emissions should consider the potential chemical reactions that occur downwind. Mercury concentration in coal is not particularly high compared to average rocks, but the high volatility of Hg combined with the large amounts of coal used to fuel power plants yields a high mass of Hg entering the flue gas stream. Although many power plants currently utilize flue gas desulfurization (scrubbers), these are not primarily designed for the removal or control of Hg (US EPA 1997); however, scrubbers do remove sulfur complexes, consequently removing some of the Hg that is chemically bonded to the complexes (Chu and Porcella, 1995). Other emission control devices also have an additional beneficial side effect of removing Hg from the flue gas. Selective catalytic reduction for nitrogen oxides and electrostatic precipitators for particulates, when combined with flue gas desulfurization, can remove approximately 90 percent of the Hg in the flue gas (Chu and Porcella, 1995).
Interchanges of Hg species also can occur in the stacks themselves. Mercury can react easily with Cl- and sulfite groups, which are both common constituents of flue gas; therefore, it may oxidize and reduce several times in the stacks depending on height of the stack and concentration of those anions (Chu and Porcella, 1995). Elemental Hg (Hg0), reactive gaseous Hg (RGM), and particulate-bound Hg (Hgp) are the most relevant species released from coal-fired power plant stacks and the heights of the stacks have an inverse relationship with dry deposition (Pleijel and Munthe, 1995; US EPA, 1997). That is, the lower the stack height, the higher the rate of local deposition.
Hg0 is the most common form of Hg in the atmosphere and usually has the longest mean residence time, approximately one year (Lindberg et al., 2007; US EPA, 1997; Pleijel and Munthe, 1995). RGM and Hgp have a mean residence time ranging from only a few hours to several months. Both of these species are influenced in the atmosphere by sulfite and ozone (Lamborg et al., 1995; Lindberg et al., 2007; Slemr et al., 2011). RGM will readily complex with sulfur, chloride, and hydroxyl ligands; however, Hg has a particular affinity for sulfites (Pleijel and Munthe, 1995; Lamborg et al., 1995). Mercury can also complex with soot particles and fine particulates in the air to form Hgp, a common form emitted from coal-fired power plants and other combustion sources (US EPA, 1997; Pleijel and Munthe, 1995; Wängberg et al., 2008).
Atmospheric Hg is transported to terrestrial and aquatic ecosystems by wet and dry deposition. RGM and Hgp can be scavenged from the air by rainout and washout with precipitation and the resulting wet deposition will most likely occur locally (US EPA, 1997; Pleijel and Munthe, 1995; Lindberg et al., 2007, Driscoll et al., 2013). Generally, wet and dry deposition rates are influenced by the Hg amount present in the atmosphere. RGM is more soluble and reactive than Hgp, and therefore, will be removed from the atmosphere at a higher rate (US EPA, 1997). Local and regional scale RGM concentrations in soil and waterways will be increased due to human-influenced sources such as coal-fired power plants because there is more RGM available for removal from the atmosphere (Lindberg et al, 2007). Conversely, Hg0 can remain in the atmosphere and circulate and deposit as part of a global Hg pool (US EPA, 1997). However, Hg0 can be dry deposited at high rates in the forest canopies because it is taken into the leaves and needles as a gas and incorporated into the plant material. RGM also is dry deposited in forests and urban areas. Annual litterfall transfers the accumulated canopy Hg0 (and some RGM) to the forest floor, thus adding to the local soil Hg concentration levels (Lindberg et al., 2007; Driscoll et al., 2013). Forest canopies, forest floors, and leaf litter are a sink for Hg (Grigal, 2002; Gustin et. al., 2008). A portion of both Hg0 and RGM deposited to the earth’s surface or oceans can become Hg0 that is re-emitted to the atmosphere (US EPA, 1997; Gustin et al., 2008; Driscoll et al., 2013). In aquatic ecosystems, some of the Hg is converted to organic methylmercury (MeHg) by sulfate-reducing or iron-reducing bacteria in anoxic conditions stimulated by dissolved organic carbon. The water soluble MeHg enters phytoplankton at the base of the food chain. Food web accumulation and bioconcentration occurs as the successively higher trophic levels conserve MeHg (Fig. 6; Rudd, 1995; Munthe et al., 2007; Scudder et al., 2009). In this way, large, long-lived and top-level predator fish have the greatest amounts of MeHg. These relatively complicated processes involving Hg chemistry result in a highly variable spatial distribution of environmental mercury that is difficult to model and predict.
With the exception of occupational or industrial Hg exposures, the main human health threat is from methylmercury (MeHg) (Rudd, 1995; US EPA, 1997; National Research Council, 2000; Grandjean et al., 2010; CDC, 2013). The consumption of fish containing MeHg is the primary pathway of MeHg exposure in humans, and the overwhelming majority of human exposure to MeHg is through this pathway (CDC, 2013). MeHg can build up and accumulate in the body causing severe neurological disorders in children and adults, and can cross the placenta and can harm the unborn fetus if the mother already has a high MeHg level in the body or consumes large amounts of upper-level food chain fish while pregnant (Ronchetti et al., 2006). High MeHg levels in children can cause learning disabilities, psychological disorders, and other neurological disorders (Ronchetti et al., 2006). Since methylmercury bioaccumulates in fish and is a neurotoxicant to humans, especially to developing fetuses, the U.S. has issued warnings to pregnant women and women of child bearing age to limit their intake of high-Hg species of fish. Children, who are affected by methylmercury exposure as fetuses, can have lower IQ’s, developmental delays, and other permanent motor and neurological deficiencies as a result of the inability of their bodies to metabolize and excrete the mercury. Indeed, these intellectual impairments can be translated into economic values using a cost-of illness approach. The US EPA (2011) estimate the lifetime social value of a lost IQ point to be approximately U.S. $16,500 (year 2000 dollars). The EPA concludes that mercury reduction measures from coal-fired power plants will result in between U.S. $75 million and $288 million of IQ point loss averted due to reduced mercury emissions.
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.
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).
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:(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.
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).
The current paradigm for the human Hg exposure pathway in urban environments (Fig. 6) includes:
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.
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.
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.
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.
Contributed to conception and design: GF, MR, JC, DN, ML
Contributed to acquisition of data: GF, JC, DN, ML
Contributed to analysis and interpretation of data: GF, MR, JC, DN, ML
Drafted and/or revised the article: GF, MR
Approved the submitted version for publication: GF, MR, JC, DN, ML
The authors have declared that no competing interests exist.
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.
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.
Bernard SM, McGeehin MA. 2003. Prevalence of blood lead levels =5 µg/dL among US children 1 to 5 years of age and socioeconomic and demographic factors associated with blood of lead levels 5 to 10 µg/dL, Third National Health and Nutrition Examination Survey, 1988–1994. Pediatrics 112: 1308–1313.
Binns HJ, Campbell C, Brown MJ, Centers for Disease Control, Prevention Advisory Committee on Childhood Lead Poisoning. 2007. Interpreting and managing blood lead levels of less than 10 microg/dl in children and reducing childhood exposure to lead: Recommendations of the centers for disease control and prevention advisory committee on childhood lead poisoning prevention. Pediatrics 120: 1285–1298.
Brightbill RA, Riva-Murray K, Bilger MD, Byrnes JD. 2004. Total mercury and methylmercury in fish fillets, water, and bed sediments from selected streams in the Delaware River Basin, New Jersey, New York, and Pennsylvania, 1998–2001. U.S. Geological Survey Water-Resources Investigations Report 03–4183 . 30 p.
Bryson B. 2003. A short history of nearly everything . New York: Broadway Books. 560 p.
Butler T, Likens G, Cohen M, Vermeylen F. 2007. Final report on mercury in the environment and patterns of mercury deposition from the NADP/MDN Mercury Deposition Network. http://www.arl.noaa.gov/documents/reports/MDN_report.pdf.
Canfield RL , Henderson CR , Cory-Slechta DA , Cox C , Jusko TA , et al. 2007. Intellectual impairment in children with blood lead concentrations below 10 µg/dL: The Rochester cohort study. New Eng J Med 348: 1517–1526.
Centers for Disease Control and Prevention. 2013. Mercury biomonitoring summary. http://www.cdc.gov/biomonitoring/Mercury_BiomonitoringSummary.html.
Chiodo LM , Covington C , Sokol RJ , Hannigan JH , Jannise J , et al. 2007. Blood lead levels and specific attention effects in young children. Neurotoxicol Teratol 29: 538–546.
Chlopecka A, Bacon JR, Wilson MJ, Kay J. 1996. Forms of cadmium, lead, and zinc in contaminated soils from southwest Poland. J Environ Qual 25: 69–79.
Chu P, Porcella DB. 1995. Mercury stack emissions from U.S. electric utility power plants. Water Air Soil Poll 80: 135–144.
Clark H, Haustaden D, Brabander D. 2008. Urban gardens: Lead exposure, recontamination mechanisms, and implications for remediation design. Environ Res 107: 312–329.
Cohen AJ , Ross Anderson H , Ostro B , Pandey KD , Krzyzanowski M , et al. 2005. The global burden of disease due to outdoor air pollution. J Toxicol Environ Health Part A 68(13–14): 1301–1307.
Cohen M, Artz R, Draxler RR. 2007. Report to Congress: Mercury Contamination in the Great Lakes. National Atmospheric and Oceanic Administration Air Resources Laboratory, Silver Spring, Maryland, USA.
Cohen M , Artz R , Draxler RR , Miller P , Poissant L , et al. 2004. Modeling the atmospheric transport and deposition of mercury to the Great Lakes. Environ Res 95(3): 247–265.
Crewe J. 2012. Analysis of mercury concentrations in Indiana soils to evaluate patterns of long-term atmospheric mercury deposition. (MS thesis) . Indiana University.
Dreher GB, Follmer LR. 2004. Mercury content of Illinois soils. Water Air Soil Poll 156: 299–315.
Driscoll CT, Mason RP, Chan HM, Jacob DJ, Pirrone N. 2013 Mercury as a global pollutant: Sources, pathways, and effects. Environ Sci Technol 47(10): 4967–4983.
Engstrom DR, Balogh SJ, Swain EB. 2007. History of inputs to Minnestota lakes—Influences of watershed disturbance and localized atmospheric deposition. Limnol Oceanog 52: 2467–2483.
Evers DC. 2007. Biological mercury hotspots in the northeastern United States and southeastern Canada. BioScience 57: 29–43.
Filippelli GM. 2012. Garden Safe, Garden Well. Center for Urban Health. 22 pp.
Filippelli GM, Laidlaw MAS. 2010. The Elephant in the Playground: Confronting lead-contaminated soils as an important source of lead burdens to urban populations. Perspect Biol Med 53: 31–45.
Filippelli GM, Laidlaw MAS, Latimer JC, Raftis R. 2005. Urban lead poisoning and medical geology: An unfinished story. GSA Today 15: 4–11.
Filippelli GM, Morrison D, Cicchella D. 2012. Urban geochemistry and human health. Elements 8: 439–444.
Gilfillan SC. 1965. Lead poisoning and the fall of Rome. Journal Occupational Medicine 7: 53–60.
Gould E. 2009. Childhood lead poisoning: Conservative estimates of the social and economic benefits of lead hazard control. Environ Health Persp 117: 1162–1167.
Grandjean P , Satoh H , Murata K , et al. 2010. Adverse effects of methylmercury: Environmental health research implications. Environ Health Persp 118: 1137–1145.
Grant SL, Kim M, Lin P, Crist KC, Ghosh S, Kotamarthi VR. 2014. A simulation study of atmospheric mercury and its deposition in the Great Lakes, Atmospheric Environment . doi: 10.1016/j.atmosenv.2014.05.033.
Gray JE, Theodorakos PM, Bailey EA, Turner RR. 2000. Distribution, speciation, and transport of mercury in stream-sediment, stream-water, and fish collected near abandoned mercury mines in Southwestern Alaska, USA. Sci Total Environ 260: 21–33.
Grigal DF. 2002. Inputs and outputs of mercury from terrestrial watersheds—a review. National Research Council of Canada Environmental Reviews 10: 1–39.
Gustin MS, Lindberg SE, Weisberg PJ. 2008. An update on the natural sources and sinks of atmospheric mercury. Appl Geochem 23: 482–493.
Hamester M, Stechmann H, Steiger M, Dannecker W. 1994. The origin of lead in urban aerosols—a lead isotopic ratio study. Sci Total Environ 146: 321–323.
Harris RC , Rudd JWM , Amyot M , Babiarz CL , Beaty KG , et al. 2007. Whole-ecosystem study shows rapid fish-mercury response to changes in mercury deposition. P Natl Acad Sci 104(42): 16586–16591.
Hatcher C. 2009. Mercury Distribution in Soils and Stream Sediments of Central Indiana, USA. MS Thesis . Indiana University.
Hatcher CL, Filippelli GM. 2010. Mercury cycling in an urbanized watershed: The influence of wind distribution and regional subwatershed geometry in Central Indiana, USA. Water Air Soil Poll 219: 251–261.
Hightower JM. 2009. Diagnosis: Mercury . Island Press: Washington, D.C.
Hissler C, Probst J. 2006. Impact of mercury atmospheric deposition on soils and streams in a mountainous catchment (Vosges, France) polluted by chlor-alkali industrial activity: The important trapping role of the organic matter. Sci Total Environ 361(1): 163–163.
Hu H, Shine J, Wright RO. 2007. The challenge posed to children’s health by mixtures of toxic waste: The Tar Creek superfund site as a case-study. Pediatr Clin N Am 54: 155–175.
HUD. 1999. Economic analysis of the final rule on lead-based paint: Evaluation and reduction of lead-based paint hazards in federally owned residential property and housing receiving federal assistance. U.S. Department of Housing and Urban Development. http://portal.hud.gov/hudportal/documents/huddoc?id=DOC_25478.pdf.
Johnson D, Bretsch J. 2002. Soil lead and children’s BLL Levels in Syracuse, NY, USA. Environ Geochem Health . 24(4): 375–385.
Jusko TA , Henderson CR Jr , Lanphear BP , Cory-Slechta DA , Parsons PJ , et al. 2007. Blood lead concentrations less than 10 micrograms per deciliter and child intelligence at 6 years of age. Environ Health Persp . doi: 10.1289/ehp.10424.
Karr C. 2008. Reducing childhood lead exposure: Translating new understanding into clinic-based practice. Pediatr Ann 37: 748–756.
Keeler GJ, Landis MS, Norris GA, Christianson EM, Dvonich JT. 2006. Sources of mercury wet deposition in eastern Ohio, USA. Environ Sci Technol 40: 5874–5881.
Koller K, Brown T, Spurgeon A, Levy L. 2004. Recent developments in low-level lead exposure and intellectual impairment in children. Environ Health Persp 112: 987–994.
Laidlaw MAS, Filippelli GM. 2008. Resuspension of urban soils as a persistent source of lead poisoning in children: A review and new directions. Appl Geochem 23: 2021–2039.
Laidlaw MAS, Mielke HW, Filippelli GM, Johnson DL, Gonzales CR. 2005. Seasonality and children’s blood lead levels: Developing a predictive model using climatic variables and blood lead data from Indianapolis, Indiana, Syracuse, New York, and New Orleans, Louisiana (USA). Environ Health Persp 113: 793–800.
Laidlaw MAS, Zahran S., Mielke HW, Taylor MP, Filippelli GM. 2012. Re-suspension of lead contaminated urban soil as a dominant source of atmospheric lead in Birmingham, Chicago, Detroit and Pittsburgh, USA. Atmos Environ 49: 302– 310.
Laidlaw MAS, Zahran S, Pingitore N, Clague J, Devlin G, Taylor MP. 2014. Identification of lead sources in residential environments: Sydney, Australia. Environ Pollut 184: 238–246.
Lamborg CH, Fitzgerald WF, Vandal GM, Rolfhus KR. 1995. Atmospheric mercury in Northern Wisconsin: Sources and species. Water Air Soil Poll 80: 189–198.
Lanphear BP , Hornung R , Khoury J , Yolton K , Baghurst P , et al. 2005. Low-level environmental lead exposure and childrens intellectual function: An international pooled analysis. Environ Health Perspect 113(7): 894–899.
Lanphear BP , Matte TD , Rogers J , Clickner RP , Dietz B , et al. 1998. The contribution of lead-contaminated house dust and residential soil to children’s blood lead levels. Environ Res 79: 51–68.
Lee P-K, Touray JC, Baillif P, Ildefonse JP. 1997. Heavy metal contamination of settling particles in a retention pond long the A-71 motorway in Sologne, France. Sci Total Environ 201: 1–15.
Lindberg S , Bullock R , Ebinghaus R , Engstrom D , Feng X , et al. 2007. A synthesis of progress and uncertainties in attributing the sources of mercury in deposition. Ambio 36(1): 19–33. doi: 10.1579/0044-7447(2007)36[19:ASOPAU]2.0.CO;2.
Lyons WB, Fitzgibbon TO, Welch KA, Carey AE. 2006. Mercury geochemistry of the Scioto River, Ohio—Impact of agriculture and urbanization. Appl Geochem 21(11): 1880–1888.
Macalady JL. 2000. Sediment microbial community structure and mercury methylation in mercury-polluted Clear Lake, California. Appl Environ Microbiol 66(4): 1479–1488.
Manton WI, Angle CR, Stanek KL, Reese YR, Kuehnemann AJ. 2001. Acquisition and retention of lead by young children. Environ Res 82: 60–80.
Mason RP. 1994. The biogeochemical cycling of elemental mercury: Anthropogenic influences. Geochim Cosmochim Acta 58(15): 3191–3198.
Mergler D , Anderson HA , Chan LHM , Mahaffey KR , Murray M , et al. 2007. Methylmercury exposure and health effects in humans—A worldwide concern. Ambio 26(1): 3–11.
Mielke HW. 1994. Lead in New Orleans soils: New images of an urban environment. Environ Geochem Hlth 16: 123–128.
Mielke HW, Barroughs S, Wade R, Yarrow T, Mielke PW Jr. 1984/85. Urban lead in Minnesota: Soil transect results of four cities. Journal of the Minnesota Academy of Sciences 50: 19–24.
Mielke HW, Powell ET, Gonzales CR, Ottesen RT, Langedal M. 2006. New Orleans soil lead (Pb) cleanup using Mississippi River alluvium: Need, feasibility, and cost. Environ Sci Technol 40: 2684–2789.
Mielke HW, Reagan PL. 1998. Soil is an important pathway of human lead exposure. Environ Health Persp 106: 217–229.
Miranda ML , Kim D , Galeano MA , Paul CJ , Hull AP , et al. 2007. The relationship between early childhood blood lead levels and performance on end-of-grade tests. Environ Health Persp 115: 1242–1247.
Morrison D , Lin Q , Wiehe S , Liu G , Rosenman M , et al. 2012. Spatial relationships between lead sources and children’s blood lead levels in the urban center of Indianapolis (USA). Environ Geochem Health 35(2): 171–183. doi: 10.1007/s10653-012-9474-y.
Munthe J , Bodaly RA , Branfireun BA , Driscoll CT , Gilmour CC , et al. 2007. Recovery of mercury-contaminated fisheries. Ambio 36(1): 33–44.
National Atmospheric Deposition Program. 2015. Mercury Deposition Network. http://nadp.sws.uiuc.edu/mdn/.
National Research Council. 2000. Toxicological effects of methylmercury . Washington, D.C.: National Academy Press. 344p.
Needleman H. 2004. Lead poisoning. Annu Rev Med 55: 209–222.
Neumann K, Lyons WB, Graham EY, Callender E. 2005. Historical backcasting of metal concentrations in the Chattahoochee River, Georgia: Population growth and environmental policy. Appl Geochem 20: 2315–2324.
Nevin R. 2000. How lead exposure relates to temporal changes in IQ, violent crime, and unwed pregnancy. Environ Res Section A 83: 1–22.
NHANES (National Health and Nutrition Examination Survey). 2003–2006. U.S. Centers for Disease Control and Prevention. National Center for Health Statistics. Hyattsville, MD: U.S. Department of Health and Human Services. Available: http://www.cdc.gov/nchs/nhanes.htm [accessed 10 October 2008].
Nigg JT , Knottnerus GM , Martel MM , Nikolas M , Cavanagh K , et al. 2008. Low blood lead levels associated with clinically diagnosed attention deficit/hyperactivity disorder and mediated by weak cognitive control. Biol Psychiat 63: 325–331.
Oka GA, Thomas L, Lavkulich LM. 2014. Soil assessment for urban agriculture: A Vancouver case study. J Soil Sci Plant Nutr 14(3): 657–669.
Park J-D, Zheng W. 2012. Human exposure and health effects of inorganic and elemental mercury. J Prev Med Public Health 45: 344–352.
Pirkle JL , Kaufmann RB , Brody DJ , Hickman T , Gunter E , et al. 1998. Exposure of the US population to lead, 1991–1994. Environ Health Persp 106: 745–750.
Pirrone N, Mason RP, eds. 2009. Mercury fate and transport in the global atmosphere: Measurements, models and policy implications . United States: Springer.
Pleijel K, Munthe J. 1995. Modeling the atmospheric chemistry of mercury. Water Air Soil Poll 80: 317–324.
Rabin R. 2006. The Rhode Island lead paint lawsuit: Where do we go from here. New Solutions 16: 353–363.
Rabinowitz MB, Wetherill GW. 1972. Identifying sources of lead contamination by stable isotope techniques. Environ Sci Technol 6: 705–709.
Risch MR, Baker NT, Fowler KK, Egler AL, Lampe DC. 2010. Mercury in Indiana Watersheds: Retrospective for 2001–2006. U.S. Geological Survey Professional Paper 1780 , 66p. plus appendixes.
Risch MR, Fowler KK. 2008. Mercury in precipitation in Indiana, January 2004–December 2005. U.S. Geological Survey Scientific Investigations Report 2008-5148 . 76 p.
Risch MR , Gay DA , Fowle KK , Keeler GJ , Backus SM , et al. 2012. Spatial patterns and temporal trends in mercury concentrations, precipitation depths, and mercury wet deposition in the North American Great Lakes region, 2002–2008. Environ Pollut 161: 267–271.
Risch MR, Kenski DM, Gay DA. 2014. A Great Lakes atmospheric mercury monitoring network: evaluation and design. Atmos Environ 85: 109–122.
Roberts JR, Reigert JR, Eberling M, Hulsey TC. 2001. Time required for blood lead levels to decline in nonchelated children. J Toxicol-Clin Toxic 39: 153–160.
Ronchetti R , Zuurbier M , Jesenak M , Koppe JG , Ahmed UF , et al. 2006. Children’s health and mercury exposure. Acta Pediatr 95 (Suppl 453): 36–44.
Roux KE, Marra PP. 2007. The presence and impact of environmental lead in passerine birds along an urban to rural land use gradient. Arch Environ Con Tox 53: 261–275.
Rudd JWM. 1995. Sources of methyl mercury to freshwater ecosystems: A review. Water Air Soil Poll 80: 697–713.
Schnaas L , Rothenberg SJ , Flore MF , Martinez S , Hernandez C , et al. 2006. Reduced intellectual development in children with prenatal lead exposure. Environ Health Persp 114: 791–797.
Scudder BC , Chasar LC , Wentz DA , Bauch NJ , Brigham ME , et al. 2009. Mercury in fish, bed sediment, and water from streams across the United States, 1998–2005. U.S. Geological Survey Scientific Investigations Report 2009–5109 , 74 p.
Settle DM, Patterson CC. 1980. Lead in albacore: Guide to lead pollution in Americans. Science 207: 1167–1176.
Seigneur C, Vijayaraghavan K, Lohman K, Karamchandani P, Scott C. 2004. Global source attribution for mercury deposition in the United States. Environ Sci Technol 38(2): 555–569.
Skyllberg U, Qian J, Frech W, Xia K, Bleam W. 2003. Distribution of mercury, methyl mercury and organic sulfur species in soil, soil solution and stream of a boreal forest catchment. Biogeochem 64: 53–76.
Slemr F, Brunke E, Ebinghaus R, Kuss J. 2011. Worldwide trend of atmospheric mercury since 1995. Atmos Chem Phys Discuss 11(1): 2355–2375.
Sorensen JA, Kallemeyn LW, Sydor M. 2005. Relationship between mercury accumulation in young-of-the-year Yellow Perch and water-level fluctuations. Environ Sci Technol 39: 9237–9243.
Surkan PJ , Zhang A , Trachtenberg F , Daniel DB , McKinlay S , et al. 2007. Neuropsychological function in children with blood lead levels <10 µg/dL. Neurotoxicology 28: 1170–1177.
Tan SW, Meiler JC, Mahaffey KR. 2009. The endocrine effects of mercury in humans and wildlife. Crit Rev Toxicol 39(3): 228–269.
Ter Haar G, Aronow R. 1974. New information on lead in dirt and dust as related to the childhood lead problem. Environ Health Persp 7: 83–89.
U.S. Environmental Protection Agency. 1997. Mercury Study Report to Congress: Fate and Transport of Mercury in the Environment. v. 3, 376 p. http://www.epa.gov/ttn/oarpg/t3/reports/volume3.pdf.
U.S. Environmental Protection Agency. 2001. Water quality criterion for the protection of human health-Methylmercury. Office of Water, EPA-823-R-01-001 .
U.S. Environmental Protection Agency. 2011. Regulatory impact analysis of the proposed toxics rule: Final report. (2009-0234-3051).
U.S. Environmental Protection Agency. 2014. Envirofacts Toxic Release Inventory. http://www.epa.gov/enviro/html/tris/ez.html. Retrieved December 23, 2014.
USA Today. 2012. Ghost factories: Poison in the ground. http://www.usatoday.com/topic/B68DCD3E-7E3F-424A-BDA4-41077D772EA1/ghostfactories/ Accessed December 23, 2014.
Van Metre PC. 2012. Increased atmospheric deposition of mercury in reference lakes near major urban areas. Environ Pollut 162: 209–215.
Wängberg I , Munthe J , Amouroux D , Andersson ME , Fajon V , et al. 2008. Atmospheric Mercury at Mediterranean Coastal Stations. Environ Fluid Mech 8(2): 101–116. doi: 10.1007/s10652-007-9047-2.
Weiss AL, Caravanos J, Blaise MJ, Jaeger RJ. 2006. Distribution of lead in urban roadway grit and its association with elevated steel structures. Chemosphere 65: 1762–1771.
Wentz DA, Brigham ME, Chasar LC, Lutz MA, Krabbenhoft DP. 2014. Mercury in the nation’s streams—Levels, trends, and implications. U.S. Geological Survey Circular 1395 , 90 p.
Wright JP , Dietrich KN , Ris MD , Hornung PW , Wessel SD , et al. 2008. Association of prenatal and childhood blood lead concentrations with criminal arrests in early adulthood. PLoS Med 5: 101:1–101:9.
Wright RO, Amarasiriwardena C, Woolf AD, Jim R, Bellinger DC. 2006. Neuropsychological correlates of hair arsenic, manganese, and cadmium levels in school-age children residing near a hazardous waste site. Neurotoxicology 27: 210–216.
WHO. 2010. State of the Worlds Cities 2010/2011: Bridging the Urban Divide. 204 pp.
WHO. 2013. Lead poisoning and health. Fact sheet no. 379. http://www.who.int/mediacentre/factsheets/fs379/en/ [accessed November 11 2014].
Yorifuji T, Debes F, Weihe P, Grandjean P. 2011. Prenatal exposure to lead and cognitive deficit in 7- and 14-year-old children in the presence of concomitant exposure to similar molar concentration of methylmercury. Neurotoxicol Teratol 33: 205–211.
Young TM, Heeraman DA, Sirin G, Ashbaugh LL. 2002. Resuspension of soil as a source of airborne lead near industrial facilities and highways. Environ Sci Technol 36: 2484–2490.
Zahran S, Laidlaw MAS, McElmurry SP, Filippelli GM, Taylor M. 2013. Linking Source and Effect: Resuspended Soil Lead, Air Lead, and Children’s Blood Lead Levels in Detroit, Michigan. Environ Sci Technol 47: 2839–2845.
Zahran S, Mielke HW, Gonzales CR, Powell ET, Weiler S. 2010. New Orleans before and after Hurricanes Katrina/Rita: A quasi-experiment of the association between soil lead and children’s blood lead. Environ Sci Technol 44(12): 4433–4440. doi: 10.1021/es100572s.