1

s memorandum

Environment and Resources

55 Wheeler Street, Cambridge, MA  Tel: 617 492-7100  www.abtassociates.com

Date September 26, 2011

To Tom Born, U.S. EPA/OW/OST

From Emily Giovanni, Elena Besedin, Molly Cohen, David Mitchell

Subject Literature Review of the Impacts of Mercury from Dental Amalgam in Wastewater

This memo summarizes dental mercury discharges, impacts, and potential benefits of reductions resulting 

from the installation of mercury separators in dental offices in support of EPA’s associated rulemaking. It 

complements and expands on the previous Detailed Study (U.S. EPA, 2008). 

1 Introduction

Mercury is a well-known environmental toxicant ubiquitously distributed at low levels in most terrestrial 

and aquatic ecosystems. Humans and wildlife are exposed to mercury primarily via the consumption of 

fish from mercury-contaminated waterbodies, although other exposure pathways exist (such as through 

accidental inhalation at the workplace). When exposed to mercury, humans can experience a wide range 

of health effects, including damage to the brain, kidneys, heart, and especially nervous system. These 

impacts are particularly harmful for children, who can experience profound and permanent developmental 

and neurological delays as a result of exposure in-utero. Similar neurological effects as well as adverse 

impacts on reproduction have been documented in many wildlife species.

Because mercury is extremely toxic even at very low concentrations dangerous levels in United States 

water bodies are increasingly being detected (Scarmoutzos and Boyd, 2003). As of 2008, 50 states, 1 

territory, and 3 tribes had issued mercury advisories, with 16.8 million lake acres and 1.3 million river 

miles under advisories (U.S. EPA, 2009a). Today, mercury can be detected in virtually every area of the 

world, including the remote Polar Regions and other areas without emission sources of their own, as well 

as the upper reaches of the atmosphere (Scarmoutzos and Boyd, 2003).

Several major sources of anthropogenic mercury emissions have been identified and are increasingly 

regulated, including power-plant emissions, industrial and manufacturing activities, household products, 

agricultural applications, and waste disposal. As larger sources are increasingly regulated and controlled, 

regulatory attention has focused on identifying and reducing smaller sources (U.S. EPA, 2006). This has 

been concurrent with advances in the analytical measurement of mercury at increasingly refined detection 

limits and adoption of increasingly stringent emission restrictions and water quality standards (Stone, 

2004; U.S. EPA, 2006). 

Publicly owned treatment works (POTWs) are a known source of dilute amounts of anthropogenic 

mercury to surface waters. To reduce mercury discharges to surface waters, some states have begun 

implementing discharge limitations for POTWs in the past 15 years (CDPHE, 2004). Thus, many POTWs 

have begun to work to identify and reduce discharges of mercury entering their treatment systems. Recent

research suggests that dental offices are a primary source category of mercury entering POTWs (Larry 

Walker Associates, 2002; NACWA, 2006; Vandeven and McGinnis, 2005; U.S. EPA, 2008).

However, a detailed assessment of environmental impacts of mercury discharges from dental offices is 

complicated due to the difficulty of establishing definitive estimates of several factors, including:
2

 The exact amount discharged by dental offices,

 The amount that POTWs are capable of removing,

 The extent of methylation in the environment,

 The uptake of mercury from the environment to fish,

 Fish consumption rates, particularly by species,

 Effects of mercury to infants, children, and adults.

Additionally, the presence of naturally-occurring mercury throughout the environment can further 

complicate attribution of mercury contaminations to particular sources. Given the increased sensitivity of 

detection technology, a major challenge to scientists and regulators is to develop an appropriate 

background or baseline level against which to establish discharge goals (Scarmoutzos and Boyd, 2003), 

especially given the relative dearth of historical data. Further, estimating the share of mercury emissions 

attributable to any individual source is complicated by the fact that reported emissions are frequently 

based on widely divergent methodologies and assumptions. For example, some use direct measurement 

while others use emission factors (Scarmoutzos and Boyd, 2003). However, there are a variety of 

scholarly, governmental, industry, and other literature sources that characterize mercury in dental 

wastewater discharges and its effects.

Given available information, this literature review summarizes mercury discharges from dental offices, 

the pollutant’s fate and transport, potential human health and wildlife effects, and the reduction potential 

from dental offices. This literature review will be used as a basis for qualitative assessment of benefits of 

the Effluent Limitations Guidelines (ELGs) for mercury-containing wastes from dentist offices. The 

literature review is divided into sections that include overviews of mercury forms, discharges from dental 

offices, effectiveness of the POTW treatment and discharges to surface waters, impacts to human health 

and wildlife, reduction potential and examples of existing state and municipal separator requirement 

programs, and the potential value of benefits associated with mercury discharge reductions. The final

section provides a list of literature sources, and the appendices provide supplemental information.

2 Overview of Dental Amalgam Discharges

Dental amalgam has been used as the primary restorative component (i.e. fillings) in dentistry for over 

150 years (WDNR, 1997). It is an alloy consisting of 50% mercury, plus powdered silver, tin, zinc, and 

copper (Vandeven and McGinnis, 2005; Quicksilver Caucus, 2008a). Previously, dentists mixed their 

own amalgam from bulk liquid mercury and metal powders, but now it comes as ready-to-use capsules 

varying in size from 100 to 1,000 milligrams of mercury (IMERC, 2010).

Dental amalgam contributes to mercury pollution in United States water bodies through two pathways: 

 Air deposition. Assuming that the average person has 7 fillings, about 3 tons of mercury is

released from crematoria each year, eventually being deposited domestically or internationally 

(Scarmoutzos and Boyd, 2003). 

 Indirect discharge to surface waters via POTWs. The primary route by which dental amalgam 

reaches surface waters is wastewater discharge from dental offices to publicly-owned treatment 

works (POTWs). Between 36% and 50% of all mercury reaching POTWs is attributable to dental 

office amalgam discharge, far exceeding all other commercial or residential facilities (Vandeven 

and McGinnis, 2005; U.S. EPA, 2006). One Colorado study found that dental clinics contribute 

9.26 pounds of mercury to the Littleton-Englewood Wastewater Treatment Plant each year, or 

62% of the total (CDPHE, 2004).
3

This section 1) characterizes the different forms of mercury, 2) briefly summarizes a pass-through 

analysis previously conducted by EPA to estimate the current annual discharge rates from dental offices 

in the United States and 3) characterizes the share of total anthropogenic mercury emissions that are 

attributable to various sources including dental offices. 

2.1 Forms of Mercury 

Particulate and dissolved mercury can take many forms, most commonly elemental, ionic/oxidized, 

inorganic, or organic (also known as methylmercury; ATSDR, 1999). 

 Elemental mercury is the pure form of the metal, which is rarely found in nature. It is the 

substance used in dental amalgam, as well as other commercial products such as thermometers 

and electrical switches. Elemental mercury is volatile, displaying appreciable vapor pressure 

(Scarmoutzos and Boyd, 2003). As such, it is susceptible to absorption through the human 

respiratory tract. It is lipid soluble, and so it can cross the blood-brain barrier causing 

neurological effects, and it can cross the placenta, facilitating impacts to infants and children 

exposed in-utero after maternal exposure.

1 Elemental mercury is primarily of concern when it is 

inhaled due to its ability to absorb into the bloodstream; when it is ingested, less than 0.01% of it 

is absorbed into the digestive tract. Elemental mercury is a liquid under ambient conditions 

(Scarmoutzos and Boyd, 2003). However, when it is combined with other metals, elemental 

mercury dissolves and stabilizes, forming amalgam. Ionic mercury is elemental mercury that has 

been oxidized. 

 Inorganic mercury is elemental mercury combined with other elements such as chlorine or sulfur. 

Common species are mercuric chloride, mercuric acetate, mercuric sulfide, and others. This type 

of mercury is usually made up of white powders or crystals, although mercuric sulfide is red and 

is sometime used as a dying agent (including for tattoos). Inorganic mercury is found primarily in 

batteries, some disinfectants, home health remedies, or skin-bleaching creams (U.S. EPA, 2007).

Inorganic mercury has low lipid solubility, and, on its own, is not as serious a health concern as 

elemental or organic mercury.

 Organic mercury, or methylmercury, is mercury combined with carbon. This form is primarily 

produced from other forms of mercury by anaerobic bacteria in the aquatic environment (usually 

in water body sediments). Methylmercury is the primary concern associated with mercury 

discharges to the environment. It is highly lipid-soluble, and it primarily resides within fatty 

tissues in the body so it is eliminated from the body slowly (Scarmoutzos and Boyd, 2003). 

Mercury discharged from dental offices primarily remains in elemental, amalgamated form throughout 

wastewater conveyance and treatment (Vandeven and McGinnis, 2005), during which it is typically comingled with other sanitary wastes. However, once it reaches surface waters, it can be transformed into 

methylmercury which is the primary health and ecological concern associated with mercury discharges.

Exhibit 2-1 summarizes forms of mercury, their primary sources, and the relevant physical characteristics.

 

1 Discussed further in Section 4.2
4

Exhibit 2-1. Forms of Mercury

Form Species Primary Sources/Load Solubility Physical State

Water Lipid

Elemental 

Mercury Hg0

Vapor from dental amalgam; 

Incineration of dental and medical 

waste; air emissions; vast majority of 

mercury in atmosphere (95-97%)

Low High Heavy, mobile, 

liquid metal

Oxidized or 

Ionic 

Mercury

HgII

Produced via oxidation of elemental 

mercury or demethylation of MeHg; 

Water emissions; May be dissolved in 

POTW effluent

High Low Crystals, granules, 

or powder

Inorganic 

Mercury

Cl-Hg-Cl

Hg-S

Used in a variety of products 

(fungicides, insecticides); 

combination of ionic mercury with 

sulfur, oxygen and chlorine

LowMedium1

Low

Powders or 

crystals – mostly 

white but can be 

red or black

Methylmercury

CH3-Hg+

Few anthropogenic sources, aside 

from indirect via methylation of 

inorganic and ionic mercury; Stored 

in macroinvertebrates and fish and 

mammal tissue; most common form 

of organic mercury

Low High White powder or 

crystalline solids

1. Water and lipid solubility depend on the compound in question. Mercuric chloride, one of the most 

common inorganic mercury compounds, is soluble in water.

2.2 Dental Amalgam Use and Waste

Given the availability of new alternatives, improved dental hygiene, and concerns about the health 

implications of mercury, use of mercury amalgam by dentists has decreased since the 1980s. For example, 

the American Dental Association (ADA) estimates that the number of restorations in 1999 was 71 

million, down from 99.5 million in 1990 (Vandeven and McGinnis, 2005). Nonetheless, dentistry remains 

one of the largest mercury consuming industries in the United States, largely because amalgam is less 

expensive than other types of restorative materials such as resin composite, glass ionomer, resin ionomer, 

porcelain, and gold alloy (Vandeven and McGinnis, 2005; Quicksilver Caucus, 2008a). Dental amalgam 

is the largest remaining dissipative uses of mercury, since many other dispersing uses of mercury2have 

been decreased or eliminated (Swain, et al., 2007).

Vandeven and McGinnis (2005) estimate that there were over 120,000 private dentists in the United 

States in 2001, with 76% using dental amalgam. An estimated 31.3 million tons of waste amalgam is 

generated by United States dentists annually (30.5 tons of suspended solid mercury and 0.8 tons of 

dissolved mercury).

3

Not all of the waste amalgam is discharged to receiving sewer systems. Currently, all dentists in the 

United States use some mechanisms to capture dental amalgam, including chair-side traps and vacuum 

filters. About 80% use vacuum filters combined with chair-side traps, with an overall capture efficiency 

of 81%, while most others use chair-side traps only, with an overall capture efficiency of 68% (Vandeven 

and McGinnis, 2005). As summarized in Appendix A, 11 states and several municipalities require that 

dental offices also use dental amalgam separators, which remove a higher percentage of amalgam

 

2

Including fungicides, paper manufacturing, and skin-lightening soaps

3

For placements, assumes that the average mercury per filling is 450 mg, 25% of it is not used, and that 9% of amalgam from 

placements is discharged as waste; for removals, assumes that the average removal generates 300 mg of amalgam. For a 

more detailed description, see the Detailed Study (U.S. EPA, 2008)
5

particles (usually over 96%) using sedimentation, filtration, centrifugation, and/or ion exchange 

(McManus and Fan, 2003; CDPHE, 2004). 

However, there are no commercially available separators that remove all dental mercury from the waste 

stream. Specifically, dissolved mercury or molecularly dispersed mercury species can pass through 

separators and other mechanisms because it is too fine to settle out with other particles (Scarmoutzos and 

Boyd, 2003). Additionally, surfactants and disinfectants commonly discharged in the dental wastestream 

can increase mercury dissolution, causing a larger share of it to pass through traps, filters, and/or 

separators to be discharged to POTWs. The amalgam that is not captured by in-house control technologies

flows out of the dental facility to POTWs (Vandeven and McGinnis, 2005).

2.3 Dental Amalgam Share of Total Mercury Loading

Overall, 60% to 70% of mercury contaminating United States waters is attributable to anthropogenic 

sources, while the remaining 30% to 40% is from natural sources such as volcanoes, geologic or mineral 

deposits, and volatilization from the oceans (Scarmoutzos and Boyd, 2003; Trasande, et al., 2005).

Exhibit 2-2 shows the approximate anthropogenic shares of mercury in the environment attributable to 

different sources as of 1994 and 1995 (the date of the most recent comprehensive data set; as cited in 

Scarmoutzos and Boyd, 2003).4 The combustion of coal and other fossil fuels is the largest source of

mercury emissions (Trasande, et al., 2005; Swain et al., 2007).Combined incineration of municipal and 

medical waste is the second largest (Mohapatra, et al., 2007). 

Based on these data, dental offices appear to represent a relatively small share of total anthropogenic 

mercury in the environment (less than 1%). However, some of the dental amalgam discharged to POTWs

is settled out with biosolids during wastewater treatment.

5

In some cases (about 22% of the time), these 

biosolids are incinerated or otherwise volatilized into the atmosphere. Because of this, as Scarmoutzos 

and Boyd (2003) point out, dental mercury emissions also contribute to municipal and medical waste 

incineration categories. As such, they account for a larger share of total mercury in the environment –

possibly up to 4% to 19% – and they are responsible for discharges directly to water bodies, as well as 

atmospheric emissions that will eventually be deposited. Dental offices also represent a unique source in 

that they are numerous and very diffuse across the nation, further complicating efforts to curtail emissions 

(compared with less numerous sources such as coal burning utility plants).

Exhibit 2-2. United States Annual Anthropogenic Mercury Discharges by Source

Source Discharge Medium

(air/water)

Amount

(tons per year) Percentage of Total

Coal Burning Utility 

Plants Air 52.0 33.7%

Boilers Air 32.1 20.8%

 Commercial Air 28.8 18.6%

 Residential Air 3.3 2.1%

Municipal Waste 

Combustors Air 29.6 19.2%

Medical Waste 

Incinerators Air 16 10.4%

Hazardous Waste 

Combustors Air 7.1 4.6%

 

4 Although it is the most recent comprehensive data available, the data shown in Exhibit 2.2 suggests a lower contribution by 

dental offices than most updated estimates from and by the dental industry.

5 Treatment of biosolids is discussed in more detail in Section 3.2.
6

Exhibit 2-2. United States Annual Anthropogenic Mercury Discharges by Source

Source Discharge Medium

(air/water)

Amount

(tons per year) Percentage of Total

Chlor-Alkali Industry Air 6.5 4.2%

Portland Cement Air 4.4 2.8%

Lamp Breakage Water 3.1 2.0%

Pulp and Paper Industry Air 1.7 1.1%

General Laboratories Water 1.4 0.9%

Dental Offices Water 0.6 0.4%

Total -- 154.5 100%

Adapted from Scarmoutzos and Boyd (2003); based on EPA data from 1994 and 1995.

With regard to air emissions, mercury can have long atmospheric residence, eventually being deposited 

far from the emission source. As such, only about 33% of anthropogenic air emissions released in the 

United States are deposited domestically; the rest enters the global mercury pool (Trasande, et al., 2005);

with anthropogenic emissions from the United States accounting for about 3% of the total global mercury 

pool (U.S. EPA, 2006). Likewise, much of the mercury deposited domestically was emitted 

internationally and came from the global mercury pool.6 This global-scale intermixing adds complexity to 

any effort to track emissions attributable to any particular domestic source, including dental amalgam. 

3 POTW Treatment, Pass-Through, and Discharges to Surface Waters

Once dental amalgam is discharged with wastewater, it enters sewer systems or septic tanks. This section 

describes the waste treatment methods (e.g., primary, secondary and tertiary treatment) and removal 

efficiencies typically associated with POTWs, how much mercury is likely to be discharged as effluent to 

receiving waters, and how mercury is converted to methylmercury once it reaches surface waters.

3.1 Treatment at POTWs and Surface Water Discharges

Once wastewater from dental facilities reaches one of 17,345 POTWs in the United States, it undergoes 

one or several treatment processes. Preliminary treatment consists of passing the wastewater through 

filtering bars or screens to remove large pieces of debris or trash, and the primary treatment consists of 

sedimentation tanks, where solid materials (including some amalgam particles) settle to the bottom of the 

tank. The solid materials are usually called sludge or biosolids, and are pumped out of the tank, often for 

further treatment. Secondary treatment of the wastewater consists of “activated sludge process,” or 

aeration, in which bacteria consume the remaining organic materials in the wastewater. In some cases, the 

water then goes through disinfection (usually via chlorine, sometimes UV) before it is discharged to the 

receiving waterbody. POTWs have a 95% capture rate for mercury (Larry Walker Associates, 2002), with 

most of the capture occurring during preliminary treatment, when particulate mercury settles out with 

other solids.

Previously, one of the greatest difficulties associated with enforcement of stringent limitations on mercury 

discharges, including those from POTWs, was that monitoring technology was not advanced enough to 

detect mercury at or below the designated water quality criteria. With regard to trace metals, it proved 

particularly challenging to preclude sample contamination during collection, transport, and analysis (U.S. 

EPA, 2002). However, method 1631E, approved by EPA in 1999, is designed specifically to preclude 

contamination,7allowing regulators to produce reliable results even at the lowest water quality 

 

6

For example, EPA (2006) estimates that only 17% of mercury deposited in the United States originated domestically or in 

Canada; the remaining 83% came from international sources.

7 Via oxidation, purge and trap, and cold-vapor atomic fluorescence spectrometry (CVAFS).
7

certification levels published by EPA (between 0.5 ng/L and 100 ng/L; Hanlon, 2007; U.S. GAO, 2008).

Currently, about 6% of POTWs have permits with mercury limitations, and another 10% have monitoring 

requirements (Larry Walker Associates, 2002).

The Great Lakes, which have particularly stringent mercury criterion for wildlife at 1.3 ng/L and human 

health at 3.1 ng/L, provide an example (U.S. GAO, 2008). Canada and the United States jointly developed 

the Great Lakes Water Quality Agreement of 1972 and the Great Lakes Binational Toxics Strategy of 

1996 and 1997 with the goal of the virtual elimination of persistent toxic chemicals including mercury in 

the Great Lakes (U.S. EPA, 2000). However, prior to the deployment of test method 1631E, EPA was not 

able to detect mercury at the level of the established water quality criterion (U.S. GAO, 2008). The 

approval and use of the new method thereby facilitated the identification of exceedances in pursuit of the 

United States’ commitments. Thereafter, the number of permits for discharges to Great Lakes waterbodies 

with mercury limitations increased from 185 in 2005 to 292 in 2007.

8

 

Permits for POTWs discharging to other water bodies are often based on the Great Lakes Initiative 

criterion. However, many of them are having difficulty meeting their discharge requirements, since 

dissolved mercury can pass through standard POTW treatments (i.e. it is not settled out during 

sedimentation as with most of the particulate mercury) and enter surface waters. In order to address 

mercury discharges, POTWs are primarily relying on source control in an effort to meet stringent permit 

limitations for discharges to surface waters (for example, see U.S. EPA, 2004).

3.2 POTW Sludge Disposal and Potential Discharges

Much of the mercury that is captured by POTW treatments is settled out along with biosolids during the 

sedimentation phase (Mohapatra, et al., 2007). However, the mercury content in biosolids is typically 

very low. For example, data from Pennsylvania shows that the mean and 95th percentile concentrations of 

mercury in sewage sludge is well below standards for metals allowable for land-applied biosolids (NRC, 

2002). However, the fate and transport of mercury settled out with biosolids depends on the treatments 

and disposal methods employed by the particular POTWs.

Many POTWs treat sludge in digesters, or heated oxygen-free tanks, where anaerobic bacteria consume 

organic material (NYCDEP, nd.). Although the evidence is mixed, some research suggests that mercury 

can be converted to methylmercury within sludge digesters, especially in summer months 

(Pavlogeorgatos, et al., 2006; Gbondo-Tugbawa, et al., 2010). After digestion, sludge can then be dewatered to reduce its volume, creating biosolids that are disposed of via one of several potential methods, 

including land application, composting, incinerating, or landfilling. If the sludge is particularly high in 

mercury (or other toxics) municipalities may require disposing of it as hazardous waste, which is very 

expensive (Stone, 2004).

EPA has a longstanding policy of promoting the beneficial uses of biosolids, including composting 

(whereby biosolids are mixed with a bulking agent and decompose) or land application (U.S. EPA, 1999; 

NYCDEP, nd.; NRC, 2002; Mohapatra, et al., 2007). Currently about 60% of the 5.6 million tons of 

biosludge generated annually in the United States is used for land application and is applied to about 0.1% 

of all agricultural lands in the United States (U.S. EPA, 1999; NRC, 2002). In addition to agricultural 

application, sludge is applied to forests, reclamation sites, or public sites such as golf courses and parks. 

Biosolids used as fertilizer have stringent standards for nine inorganic chemicals including mercury (U.S. 

EPA, 1999; 2010), and so higher mercury levels may preclude this beneficial use. However, mercury 

disposed of via land application is well below the threshold (U.S. EPA, 2009b).

 

8 EPA estimated that it would take at least two 5-year permit cycles to completely implement the stricter permit limitations 

associated with the improved monitoring (U.S. GAO, 2008).
8

Landfilling is another common disposal method (U.S. EPA, 1999). NYDEC (2006) indicates that 26% of 

biosolids in New York are landfilled by 50% of the state’s POTWs. This represents an increase from 17% 

in 1998. This may be due to multiple factors, including that the number of landfills accepting biosolids 

has increased over the same time period, and that tipping fees make the option attractive to communities 

(NYDEC, 2006).

Beneficial disposal methods and landfilling can result in the risk of erosion and runoff of mercury 

contained in the sludge to local waterbodies (NRC, 2002). Watershed characteristics such as the size and 

topography can affect mercury loading via runoff (Munthe et al., 2007). Very large watersheds may be 

less responsive to loading reductions because of reduced yield per unit area but larger overall yields, and 

forest disturbances can increase loadings (Munthe et al., 2007). In addition to concerns about runoff, there 

is some evidence that sunlight and heat can cause mercury to volatize from the soil into the atmospheric 

pool (NRC, 2002).

Incineration is another common disposal method of biosolids. This process reduces biosolids to residue 

that consists of ash taking up about 20% of the original volume (U.S. EPA, 1999). U.S. EPA (1999) 

estimates that about 22% of biosolids are incinerated nationally. There is no method that can feasibly 

remove trace mercury from the biosolids in advance of the incineration process (WERF, 2009), and as 

such this method may result in the discharge of mercury to the atmosphere (CDPHE, 2004). Vandeven 

and McGinnis (2005) estimate that about 1 ton of dental amalgam is incinerated in sewer sludge 

incinerators. Of this amount, 0.2 tons are emitted to the atmosphere, with 0.1 tons being atmospherically

deposited within the United States and ultimately reaching domestic surface water. 

The Water Environment Research Foundation (WERF, 2009) summarizes actions by incinerators to 

remove mercury from atmospheric emissions, which vary depending on the operation and the physical 

properties of the mercury species present. Elemental mercury, which is the predominant form from dental 

amalgam, is difficult to capture, and presents the greatest challenge to control strategies because it is 

water insoluble. This property makes wet scrubbing ineffective at controlling elemental mercury, and it is 

also resistant to electrostatic precipitation (WERF, 2009). This challenge is compounded by the fact that 

other types of mercury can be converted to elemental mercury by the high temperatures in incinerators. 

Other types of mercury can be partially removed by wet scrubbing processes, and further reductions have

been demonstrated via the use of fabric filters and wet electrostatic precipitators. To date, only two 

biosolid incinerators have installed mercury-specific control systems call activated carbon fixed-bed 

polishing systems: 1) Ypsilanti, Michigan, which reports that the systems removes over 99% of mercury 

from exhaust gas, and 2) St. Paul, Minnesota, which reports 94% efficiency with that system, and an 

additional 4% from a wet electrostatic precipitator.

Although it depends on the particular species, most types of mercury are highly soluble (Lindberg, et al., 

2007). Because of this, once it enters the atmospheric pool, it has a long residency and can often be 

deposited far distances away from the original discharge point, including places that have few sources of 

emissions such as the Arctic (Hylander and Goodsite, 2006). However, there is also evidence that 

mercury can be deposited relatively close to the emission source. In Seattle, deposition decreased 

significantly coincident with decreased emissions from local incineration sources (Lindberg, et al., 2007). 

3.3 Methylation in Surface Waters

Methylmercury exposure via fish consumption is the primary health and ecological concern associated 

with mercury pollution. However, all types of mercury discharges can contribute to this exposure 

pathway, since mercury can be transformed into methylmercury by bacteria present in the sediments of 
9

water bodies.9 Microorganisms and decomposers in the top layers of sediments and soils (such as bacteria 

and fungi) will, as a by-product of natural sulfate-reduction metabolism, convert elemental or other

inorganic forms of mercury to methylmercury (USGS, 1997; WDNR, 1997; ATSDR, 1999; Chin, et al., 

2000; U.S. EPA, 2005a; Mohapatra, et al., 2007). Once mercury is converted to methylmercury, it can 

enter the food chain through ingestion by bacteria and phytoplankton which are consumed by invertebrate 

grazers, with subsequent transfer to smaller forage fish and eventually to larger game fish and piscivorous 

waterfowl and wildlife (USGS, 1997).

Many water body-specific factors such as carbon sources and sulfur species present can affect the rate and 

degree of methylation (WDNR, 1997; U.S. EPA, 2006; Munthe, et al., 2007). The biological cycle of 

mercury is delicately balanced, and small changes in input rates and the chemical form of mercury may 

result in increased methylation rates in sensitive systems (Eisler, 1987). Waters that are already impaired 

have higher potential for anaerobic methylation. For example, higher concentrations of dissolved organic 

carbon (DOC) and lower pH levels are likely to increase the mobility of mercury in the aquatic ecosystem 

(USGS, 1997; WDNR, 1997). Also, increased water sulfate levels are likely to increase methylation 

(WDNR, 1997). Munthe, et al. (2007) state that anoxic conditions (especially in wetlands), warm, 

shallow-organic rich lake sediments, and increased decomposition due to flooding are conditions that 

favor methylation.

According to Chan, et al. (2003) and Scheuhammer, et al. (2007) ecosystems that are especially 

vulnerable to methylation are low alkalinity, low pH surface waters with high organic carbon 

concentrations or large upstream or adjacent wetlands, dark-water lakes and streams, and waters with 

adjoining or upstream terrestrial areas subject to flooding. According to the authors, piscivorous wildlife 

species living in inland freshwater habitats have higher mercury exposure rates than the same species 

living in nearby estuarine or marine habitats. Also, human activities can increase rates of methylation. For 

example, flooding of soils to create reservoirs increases the decomposition of vegetation, increasing the 

dissolution of organic carbon, which causes increased release of mercury that is bound to organic material 

(Chan, et al., 2003). Eisler (1987) found close correlations between mercury content of zooplankton and 

suspended particulate matter and sediment and fish muscle, suggesting that suspended matter and 

zooplankton are a pathway between sediments and fish.

4 Environmental Impacts

The two main categories of concern with respect to mercury in the environment are human health effects 

(most commonly attributed to the consumption of mercury-contaminated fish) and wildlife effects. There 

have been several incidences of high exposures and poisonings demonstrating that high methylmercury 

exposures can cause acute, immediate, and in some cases severe symptoms among both humans and 

wildlife. 

One of the most well-known and well-studied incidences contributing to the current knowledge regarding 

effects from mercury exposures via fish consumption occurred in Minamata Bay, Japan. This incident 

was caused by a chemical plant that released mercury directly into the bay for several years. Among 

infants and children, effects included cerebral palsy, mental deficiency, primitive reflexes, limb 

deformities, altered physical growth, sensory impairments, dysarthria, hyperkinesias, delays reaching 

developmental milestones, increased incidents of stillborn male infants (Mergler, et al., 2007; U.S. EPA, 

2005a; 2005b; Mohapatra, et al., 2007; CDC, 2009). Among adults, effects included paresthesia, ataxia, 

dysarthria, hearing impairment, progressive constriction of the visual field, sensory disturbance, tremors, 

 

9 Methylation can also occur aerobically in certain environments (such as soils), but primarily it occurs anaerobically in surface 

water sediments.
10

difficulty walking, irritability, death (Carrington, et al., 2004; Mergler, et al., 2007; U.S. EPA, 2005b; 

CDC, 2009)

In addition to human health impacts, a wide variety of animals were affected by this incident. Abnormal 

mercury content was found among many fish and shellfish, as well as necropsied cats who had 

succumbed to what was known as Minamata disease. Among others, observed wildlife effects in 

Minamata included poisoned cats, dogs, rats, waterfowl, and pigs, who behaved erratically (such as 

walking in circles) and died, flying crows and grebes that suddenly fell into the sea and drowned, and a 

large number of dead fish that were found floating on the surface of the sea (Eisler, 1987).

However, although the Minamata Bay incident and others significantly contributed to the current 

knowledge concerning the effects of methylmercury on human health and wildlife, exposures attributable 

to dental mercury discharges are likely to be subtler and cause human health and wildlife effects more 

consistent with chronic, low-level exposure. These levels of exposure can also cause cognitive and 

developmental effects in populations with moderate fish consumption patterns. This section will explore 

the relatively extensive body of literature addressing these topics, touching on exposure pathways, 

mercury behavior within organisms, levels of exposure considered to be harmful and specific effects on 

human health and wildlife.

4.1 Bioaccumulation

Mercury concentrations in most surface waters are well below the drinking water maximum concentration 

limit of 2 ppb (ATSDR, 1999; Scarmoutzos and Boyd, 2003). However, concentrations of mercury in fish 

are generally on the order of a million times the concentration in the water, due to bioaccumulation (U.S. 

EPA, 2005a; 2006). Methylmercury is highly lipid-soluble, meaning that organisms (including wildlife 

and humans) sequester mercury in body tissues much faster than they lose it (USGS, 1997; WDNR, 1997; 

Scarmoutzos and Boyd, 2003). Organisms at the lowest trophic levels such as phytoplankton and bacteria

absorb mercury from the aquatic environment through a combination of passive diffusion and facilitated 

transport (Chin, et al., 2000; U.S. EPA, 2005b). These organisms are consumed by organisms at 

intermediate trophic levels (e.g., zooplankton, young-of-year fish) which are, in turn, consumed by higher 

trophic level organisms. This process is biomagnification, or the incremental increase in the concentration 

of a contaminant at each level of the food chain, since each link is progressively more concentrated 

(USGS, 1997).

Due to this bioaccumulation and biomagnification, organisms at the highest trophic levels (such as 

predatory fish, piscivorous mammals or birds, and humans) reach the highest concentrations of mercury

(WDNR, 1997). This accumulation can be profound, with biomagnifications of 100,000 times from algae 

to top predators having been documented (Chin, et al., 2000). Because of this, additions of relatively 

small amounts of mercury that reach methylation zones can lead to disproportionate effects at higher 

levels of the food chain (Munthe, et al., 2007). Although bioaccumulation also occurs in terrestrial 

environments, it is much more pronounced in aquatic ecosystems, possibly because aquatic chains 

generally have more trophic levels.

Because of the complex pathways by which mercury loading, methylation, and bioaccumulation occur, 

there is a wide range of ecosystem response rates, from months to decades (Munthe, et al., 2007). 

Oftentimes, there is a long lag between emissions and effects, making identification and quantification of 

damages difficult (Hylander and Goodsite, 2006). The timing and magnitude of changes in mercury 

impacts from increased or decreased loading depends on case-specific factors such as timing and extent of 

loading, the chemical form of mercury, ecosystem characteristics, and the role of the terrestrial 

environment in delivering atmospherically deposited mercury to waterbodies (Munthe, et al., 2007).
11

4.2 Human Health Effects

Bioavailability (i.e. the fraction of mercury that is absorbed into the blood stream and it able to exert toxic 

effects) depends in part on the type of mercury, since each type has a different route of entry, absorption 

rate into circulatory system, biostability, and capacity for biotransformation. For example, while 

elemental mercury is easily absorbed into the circulatory system upon inhalation, it is not easily absorbed 

when it is ingested (with less than 0.01% being absorbed in the gastrointestinal tract). Methylmercury, in 

contrast, is readily absorbed upon ingestion, with an absorption rate of 95% (Scarmoutzos and Boyd, 

2003). Methylmercury also has a longer half-life (i.e. the time it takes for the body to eliminate half the 

body burden) than elemental mercury, at 70 to 80 days rather than 58. Additionally, mercury can change 

forms inside the body. Once it reaches the brain, methylmercury is slowly transformed chemically to 

mercuric mercury and is thereafter trapped in the brain.

Because elemental mercury is not easily absorbed in the circulatory system, it is considered a safe dental 

filling material. Although some mercury from fillings is ingested, the elemental form of it readily passes 

through the digestive system to be excreted (Scarmoutzos and Boyd, 2003). The average person receives 

much higher exposure to mercury from eating fish due to the process of bioaccumulation. In this way, 

methylmercury, which is readily absorbed through the digestive system, represents a more serious health 

threat than elemental mercury with regard to dental use of mercury (Scarmoutzos and Boyd, 2003).

The human health effects associated with methylmercury have been relatively well-documented via indepth studies of high exposures as well as population-level studies of relatively lower, but more chronic 

exposures. Multiple episodes of disease among numerous exposed populations that are widely separated 

geographically provide a basis for high confidence in the association of methylmercury exposure and 

adverse health effects (U.S. EPA, 1997). The primary exposure route is fish consumption. Birds and 

mammals can also be a dietary source, but they tend to store methylmercury in feathers and fur where it is 

less easily digestible, while fish accumulate it in tissue. Since methylmercury binds to proteins and amino 

acids, it cannot be removed by cooking or cleaning since the muscle is retained (Mergler, et al., 2007).

Once it is ingested, mercury passes to the liver and quickly enters the bloodstream to go to other parts of 

the body (ATSDR, 1999; Goldman and Shannon, 2001; Rice and Hammit, 2005; U.S. EPA, 2005a). It 

combines with amino acids, so it can easily cross the blood-brain barrier (WDNR, 1997; Goldman and 

Shannon, 2001). Methylmercury causes a wide range of effects, but target organs are the brain, liver, and 

kidneys (Chin, et al., 2000; Rice and Hammit, 2005). It has a strong affinity for sulfur and as such can 

affect cell metabolism by interfering with membrane and enzyme functions of individual cells (Chin, et 

al., 2000). 

There are several measures of mercury exposure. Hair and blood concentrations (including cord blood) 

are valid biomarkers, but convey different information about exposure (Mergler, et al., 2007). For 

example, hair analysis is best for examining dynamic changes in exposure, while blood concentration 

presents a more immediate and full picture of current exposure. EPA sets a reference dose10 (RfD) based 

on blood concentrations. Currently, this is set at 0.0001 mg/kg by weight/day (U.S. EPA, 2005a), and is 

associated with a national ambient water quality criteria for human health protection of 0.3 parts per 

million. However, this criterion is not a bright line; rather, it is intended as a broad frame of reference for 

evaluating methylmercury levels (U.S. EPA, 2005a). 

Exposures among women of childbearing age are of particular importance in establishing RfDs, since 

infants exposed in-utero are especially vulnerable to the effects. Because of this vulnerability, EPA has 

set a benchmark dose level (BMDL) for cord blood mercury of 58 ug/L, although effects have been 

 

10 an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily oral exposure to the human population 

(including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects during a lifetime
12

documented at significantly smaller exposures (Trasande, et al., 2005). The National Health and Nutrition 

Examination Survey (NHANES) reported blood concentrations and mercury intake from fish and 

shellfish among this population segment. The results of the survey indicate that 6% of women of 

childbearing age have blood concentrations above the RfD, with higher concentrations among women 

who reported eating 3 or more servings of fish during the previous 3 months (U.S. EPA, 2005a). Women 

who eat fish or shellfish at least twice per week had 7 times higher blood concentrations of mercury 

(Mahaffey, et al., 2004).

4.2.1 Fish Consumption

Consumption of fish, particularly predatory fish, is the largest source of human mercury exposure 

(WDNR, 1997; Stone, 2004; U.S. EPA, 2005b), accounting for 87% of exposures (Mahaffey, et al., 

2004). As such, patterns of human exposure to methylmercury are largely determined by global 

distribution of fisheries, trade in fish, and fish consumption (Swain, et al., 2007), and blood mercury 

levels increase with higher levels of consumption (Rice and Hammit, 2005; CDC, 2009). 

In the United States, between 70% and 90% of the population consumes seafood from the commercial 

market (Rice and Hammit, 2005). According to Goldman and Shannon (2001), the top ten commercial

fish species – canned tuna, shrimp, pollock, salmon, cod, catfish, clams, flatfish, crabs, and scallops –

represent 85% of the seafood market and contain an average mercury concentration of 0.1 ug/g. Half the 

harvest of developing nations (including tuna and other piscivorous species) is exported to the United 

States or other industrialized nations (Swain, et al., 2007). In the United States, 42% of edible fish sold 

are imported (Trasande, et al., 2005), and most fish consumed are ocean species (U.S. EPA, 2006). As 

discussed above, however, some of the imported and ocean fish are contaminated by United States 

contributions to the global mercury pool (Trasande, et al., 2005). 

In recent years, fish consumption advisories for United States water bodies and for entire states have 

sharply increased (Stone, 2004). In 2008, 16.8 million lake acres and 1.3 million river miles were 

impaired for mercury, representing increases of 19% and 42% respectively from 2006 levels (U.S. EPA, 

2009). This represents a total of 3,361 total impairments, up from 3,080 in 2006. There are several 

explanations for this upward trend, including that:

 States are increasingly issuing advisories to increase awareness of human health risk of mercury 

exposure from fish consumption (U.S. EPA, 2009a). 

 States such as Idaho, Oregon, and Wyoming have issued new statewide advisories, driving the 

increased acreage and mileage. 

 Increased frequency and sensitivity of monitoring are facilitating faster and more accurate 

identification of mercury concentrations that may pose a threat, suggesting that the mercury 

impairments themselves are not increasing as rapidly as our abilities to detect them and distribute 

the necessary information.

Attributing mercury concentrations in water bodies or fish tissue to any particular source of mercury 

discharges is extremely difficult. First, some mercury is naturally occurring in the environment, and it can 

be deposited to water bodies via volcanoes, mineral deposits, or other sources. As such, it is difficult to 

establish a baseline concentration against which to measure anthropogenic contributions. This challenge 

is complicated by the fact that adequately sensitive detection methods to identify potentially harmful 

concentrations have only recently been approved and deployed. Because of this, there is little historical 

data for establishing past concentrations and discharges. 

Additionally, anthropogenic emissions can be deposited locally or very distantly from the original source, 

depending on site-specific circumstances and atmospheric conditions. Mercury in any particular water 

body may have originated locally, regionally, or internationally. There is also a temporal factor 
13

contributing additional complexity to attribution. Once mercury reaches surface waters through

atmospheric deposition, runoff from surrounding land areas, or direct discharges, it can be stored for a 

long time (on the scale of years) before it is methylated and/or taken up into the aquatic food chain. After 

it enters the food chain, methylmercury from one source can be combined with mercury from multiple 

sources through bioaccumulation. Because of these and other factors, attributing mercury content in fish 

tissue to any particular source is virtually impossible.

4.2.2 Infant and Child Health Effects

Infants and children can be affected by mercury when they are exposed in-utero after maternal fish 

consumption. Because mercury disturbs cell divisions and neural migrations, they are particularly 

susceptible to the effects of mercury. Effects can be documented in children at intake levels 5 to 10 times 

lower than adults (WDNR, 1997). In the incidences of high exposures, asymptomatic mothers bore 

children who were severely affected, further suggesting that developing fetuses are more vulnerable to the 

effects of mercury than adults (Rice and Hammit, 2005; U.S. EPA, 2005b). Fetal mercury levels are at 

least equal to or higher than maternal levels (Goldman and Shannon, 2001), with some estimates placing 

the average cord blood mercury concentration 70% higher than the maternal blood concentration 

(Trasande, et al., 2005; Mergler, et al., 2007). Also, mercury can be transferred to breast milk, extending 

infant exposure post-birth (Goldman and Shannon, 2001). 

There have been three major epidemiological studies documenting the effects of methylmercury on 

children exposed in-utero: Seychelles (Davidson, et al., 1997), Faroe Islands (Grandjean, et al., 1997),

and New Zealand (Kjellstrom, et al., 1989). In 2000, the National Research Council (NRC) was 

commissioned by EPA to review available information on methylmercury effects; in its final report, the 

NRC summarized and evaluated these epidemiological studies (NRC, 2000). Although the Seychelles 

study was unable to establish a statistically significant relationship between mercury and childhood 

developmental test performance, the Faroe Islands and New Zealand studies did, and the Faroe Islands 

study is generally considered to be the most reliable with regard to the effects of methylmercury on 

children exposed in-utero. Based on the weight of evidence of these three studies, NRC (2000) concluded 

that low doses of mercury in-utero will likely lead to subtle but measurable neurological effects in 

children.

At low exposures (those likely to be associated with methylmercury exposure in-utero from maternal fish 

consumption), documented effects on infants and children include abnormal cognitive performance, 

delays in the development of language, memory, attention, visual-spatial and motor functions, overall 

decrease in IQ, and elevated blood pressure (WDNR, 1997; Jakus, et al., 2002; Trasande, et al., 2005; 

Carrington, et al., 2004; Mergler, et al., 2007; CDC, 2009). Also, the effects of prenatal mercury exposure 

may not become apparent until years after birth (WDNR, 1997; Jakus, et al., 2002; Mergler, et al., 2007). 

For example, Mergler, et al. (2007) did a study of 9-year-olds and found that adverse effects on higherorder cognitive functions were not apparent until the child reached a higher level of maturity.

4.2.3 Adult Health Effects

Adult health impacts from mercury consumption are less studied (Mergler, et al., 2007), but there is still a 

considerable body of evidence consistently documenting negative cognitive, neurological, and 

cardiovascular effects (U.S. EPA, 1997; WDNR, 1997). Effects of overt poisoning have been documented 

from high-exposure incidences, and several studies have documented the effects of lower, chronic 

exposures consistent with current fish consumption patterns in the United States (and those more likely to 

result from dental amalgam discharges). 

Effects from low exposures (consistent with methylmercury exposure from fish consumption) include 

visual defects such as loss of peripheral vision and chromatic and contrast sensitivity, loss of manual 

dexterity, hand-eye coordination, and grip strength, ataxia, paresthesia, hearing loss, tremors, coma, 
14

death, cerebellar changes, atrophy, focal cortical lesions, motor impairment, reproductive impairment, and 

renal toxicity (Rowe, et al., 1996; WDNR, 1997; Jakus, et al., 2002; U.S. EPA, 2005a; 2005b; Mergler, et 

al., 2007; CDC, 2009). In addition, some studies have suggested a link between methylmercury exposure 

and increased risk for cardiovascular disease, coronary heart disease, and acute myocardial infarction 

(AMI), especially among men (WDNR, 1997; U.S. EPA, 2005a; 2005b; Mergler, et al., 2007; CDC, 

2009). As EPA (2005b) notes, these results are particularly noteworthy since fish consumption has 

traditionally been considered to be protective of cardiovascular health.

4.3 Wildlife Effects

There is also consistent evidence across a number of sources that wild populations of fish, birds, and 

mammals in many regions of North America consume diets with methylmercury concentrations that are 

likely to be toxic (Mergler, et al., 2007). Although they may be significant, there is less evidence 

pertaining to effects on reptiles, amphibians, aquatic invertebrates, and vegetation (U.S. EPA, 2005b; 

Scheuhammer, et al., 2007). This section summarizes and discusses the available body of literature 

addressing the effects of mercury on wildlife.

As described above, methylmercury accumulates in fish tissue, with accumulation patterns differing 

among fish species and higher concentrations at the top trophic levels of the aquatic food chain (WDNR, 

1997; U.S. EPA, 2005b; Mergler, et al., 2007). Predatory fish such as tuna, marlin, swordfish, walleye, 

and northern pike can have concentrations as high as multiple parts per million (Mahaffey, et al., 2004). 

Signs of acute mercury poisoning among fish include flaring of gill covers, increased frequency of 

respiratory movement, loss of equilibrium, and sluggish movement, and signs of chronic exposure include 

emaciation, brain lesions, cataracts, diminished response to change in light intensity, inability to capture 

food, abnormal motor coordination, and various erratic behaviors (Eisler, 1987; U.S. EPA, 2005b). 

However, overt effects on fish occur only at very high mercury concentrations. Most fish have no 

symptoms at concentrations ten times those tolerable for humans (WDNR, 1997).

As with humans, effects of mercury exposure among fish-eating mammals and birds can be serious, often 

resulting in neurological impacts. Mercury can penetrate the blood-brain barrier and cause brain lesions, 

spinal cord degeneration, and dysfunction of the central nervous system (Chan, et al., 2003). Eisler (1987) 

notes that methylmercury primarily affects the central nervous system, particularly the sensory, visual, 

and auditory areas that are associated with coordination, irreversibly damaging the cells of the central 

nervous system. For animals with diets containing mercury concentrations higher than 1 mg/kg (primarily 

fish diets), effects may be as serious as neurotoxicity, mental derangement, coma, and death (Eisler, 1987; 

Scheuhammer, et al., 2007). 

Birds can be killed by mercury exposure, especially fish-eating birds (Eisler, 1987; USGS, 1999).

However, major die-offs caused by mercury have rarely been reported (USGS, 1999), and the dose of 

methylmercury needed to cause severe acute poisoning is higher than what is usually found in the 

environment (Chan, et al., 2003), or what is likely to result from dental amalgam discharges. Rather, the 

effect is an insidious and scattered pattern of mortality (USGS, 1999). In most cases, exposures lead to 

emaciation syndrome and a variety of sublethal effects that may act together to cause eventual death

(Eisler, 1987, Chan, et al., 2003). Chan, et al. (2003) found that birds exposed to lower levels of 

methylmercury were less likely to hunt, seek shade, walk, peck, or fly, while other activities such as 

preening increased. Other sublethal effects can include incoordination, ruffled feathers, droopy eyelids, 

falling, slowness, calmness, withdrawal, hyporeactivity, hypoactivity, weakness, tremors, delayed 

righting time, exaggerated wing response, failure to properly place feet, vertigo, anorexia, paralysis, 

convulsions, abnormal vocalizations, and adverse effects on growth, development, reproduction, hatching, 

blood and tissue chemistry, metabolism, and behavior (Eisler, 1987; WDNR, 1997; USGS, 1999; Chan, et 

al., 2003).
15

Methylmercury exposure can also have serious impacts on reproduction among wildlife at exposures 

lower than 1/5 the level needed to cause neurological symptoms (Chan, et al., 2003). As with humans, 

embryos and newborns are most at risk, since mercury affects cell division. These effects are particularly 

well documented among loons; the reproductive success of loons with high levels of methylmercury is 

about half that of unaffected loons (WDNR, 1997), and decreased reproductive success of loons breeding 

in lakes has been documented in areas that were affected by mercury accumulation from a chlor-alkali 

plant’s discharges (Chan, et al., 2003; U.S. EPA, 2005b). 

5 Reduction Potential and Municipal/State Programs

Due to mercury reductions in other sectors and industries, dental offices are now the largest source of 

mercury in wastewater influent to POTWs (Chin, et al., 2000; Quicksilver Caucus, 2008a). From a 

regulatory standpoint, they represent a unique source because each office sends a relatively very small 

amount of mercury to POTWs and they are very geographically diffuse. However, collectively they 

represent a substantial contribution of mercury because they far outnumber any other significant source.

The dental community is aware of the mercury issue and has taken measures to reduce the amount going 

off-site. Most dentists currently use chair-side traps and in some cases vacuum filters to prevent their 

wastewater systems from plugging as a result of amalgam waste. These mechanisms reduce amalgam 

discharge to some extent; however, much of the amalgam waste passes through them and into the 

wastewater being discharged to POTWs. 

Beyond the capture rates associated with typical chair side traps and filters, the amalgam separator is 

designed specifically to remove dental amalgam from wastewater and minimize the amount of mercury 

discharged to POTWs and ultimately receiving waters. Separators use a variety of methods to remove 

mercury particles (along with other solids) from wastewater, including (McManus and Fan, 2003):

 sedimentation, in which tanks or baffles reduce the speed of wastewater flow to allow particles to 

settle out (removes 90% of particles in a couple of hours),

 filtration, which can remove particles and sometimes dissolved mercury,

 centrifugation, which spins water, drawing particles to the sides of the unit, and/or 

 ion exchange, which uses chemicals that bind with mercury, drawing it to the bottom.

According to Quicksilver Caucus (2008a), properly installed and maintained separators are the most 

effective means of reducing amalgam discharges from dental offices. The International Organization for 

Standardization (ISO) provides standardized metrics and test methods for amalgam separators (ISO 

11143). This is the most commonly-used metric for evaluating separators, and it requires that 95% of 

amalgam particles by weight be removed from the waste stream.11 Most available separators achieve 

higher capture rates. For example, EPA previously reported that properly installed and maintained 

separators typically remove 95% to 99.99% of particles (U.S. EPA, 2005c). As an additional benefit, 

installation of amalgam separators would also decrease discharges of silver and other metals to POTWs 

(NACWA, 2006). 

The Association of Metropolitan Sewerage Agencies (AMSA) commissioned a study (Larry Walker 

Associates, 2002) to determine whether various source control programs (including separator 

requirements) could meaningfully reduce mercury in POTW influent and help them to achieve discharge 

standards. They assessed the current influent rates from several POTWs, and found that discharges from 

dental offices and human waste associated with dental amalgam were the two primary sources of 

 

11 60% of particles between 0.5 millimeters and 3.15 mm, 10% of particles 0.1 mm to 0.5 mm, and 30% of particles 0.1 mm or 

smaller.
16

mercury. They identified source control strategies and estimated potential reductions for them based on 

predicted effectiveness (including participation rates and other factors) and the estimated load. Seven 

plants were used as detailed case studies (including 3 in Ohio, 2 in California, 1 in Maine, and 1 in 

Minnesota). In each of the cases, dentists were the greatest contributors to mercury loads, accounting for 

an average of 35% of mercury flowing into POTWs. The source control outcome modeling also indicated

that dental offices (and hospitals) have the greatest potential for achieving measurable reductions.

Despite this potential, it is difficult to make predictions about mercury fluxes at the local level, given the 

current knowledge on mercury cycles, although it is possible to gain an understanding of fluxes on the 

global, continental, and regional scales (U.S. EPA, 2005a). As such, it may be difficult to accurately 

predict mercury reductions in waterbodies due to local installation of separators. However, there is some 

evidence of reductions attributable to separators. Some dentists have already installed separators, either 

voluntarily or to comply with state or local requirements, and there is a relatively robust literature 

addressing their effectiveness at reducing the amount of amalgam going off-site in wastewater. Many 

states and municipalities (as well as almost all European countries) already have programs requiring 

installation of separators in dental offices. These state and municipal programs are summarized in 

Appendix A. 

Although some separator programs are relatively new (for example, Cleveland’s Northeast Ohio Region 

Sewer District is implementing a separator requirement in which all offices must install a separator before 

November, 2011), some programs have demonstrated measurable reductions after implementing separator 

requirements. For example, Massachusetts and King County programs have realized 50% reductions of 

mercury discharges to POTWs (Quicksilver Caucus, 2008b). Stone (2004) and others document case 

studies of separator effectiveness in Toronto, Minneapolis/St. Paul, Duluth, Great Lakes (in Illinois), and 

Denmark. 

 Toronto: all dental offices were required to install separators by 2002. Since then, there has been 

an estimated a 58% reduction of mercury in POTW sludge, with only 800 out of 1100 facilities

participating. Full compliance was expected to result in a nearly 80% reduction in monthly 

mercury concentrations in sludge. Mohapatra et al. (2007) similarly documented a significant 

decrease in mercury levels following Toronto’s bylaws, with a 40% to 68% decrease in the 

amount going into sewers within 8 months of implementation.

 Minneapolis/St. Paul: MCES did a study of amalgam separators at 24 community dental offices. 

They found mercury reductions between 29% and 44% in POTW sludge after separator 

installation. Mercury in dental wastewater was reduced by 91% to 99%.

 Duluth: 35 regional dental facilities installed separators, and influent to POTWs from these 

facilities decreased from 0.18 pounds per day in 1993 to less than 0.02 pounds in 2002.

 Great Lakes: installation of separators on naval base dental facilities was associated with a 52% 

decrease in mercury levels in POTW influent, and exceedances of POTW mercury discharge 

limitations decreased from 54 per year to 3.

 Denmark: mercury in wastewater was decreased from 57 grams per dentist per year to 8 grams 

when separators were installed (RPA, 2002), and overall POTWs influent reductions of 80% were 

observed (Larry Walker Associates, 2002).

6 Benefits of Reducing Dental Mercury Discharges

Reductions of dental mercury discharged by POTWs into receiving waters are expected to result in 

human health and ecological benefits. The economic value of these benefits will be largely dependent on 

the types of illnesses and environmental impacts, exposure thresholds associated with those effects, the 

monetary value of these effects, and the reduction potential compared with a baseline level of discharge. 
17

The fate and transport of mercury in the environment (including methylation and bioaccumulation) will 

also be an important factor an important factor of reduction impacts, since they determine the timing and 

medium of human and wildlife exposures. 

Several factors complicate analyses of the economic benefits associated with reducing mercury discharges 

from dental offices. For example, the health benefits associated with reductions may be delayed since 

exposure can result from long-ago discharges that have been stored in water body sediments for years 

before being methylated and entering the trophic web. Dental wastewater lines may also serve as 

reservoirs of previously discharge mercury that leak over time (Stone, 2004). According to USGS (1997), 

modeled scenarios predict that if mercury emissions were decreased by 5%, it would take 8 years before 

changes in fish concentrations would be observed. Additionally, there is need to track benefits at a wide 

variety of geographic scales, from local to global, and there are uncertainties regarding mercury fate in the 

environment and the human body (Swain, et al., 2007). 

Despite these and other uncertainties, several studies have shown that decreased point source loading of 

mercury to waterbodies decreases mercury concentrations in fish (Munthe, et al., 2007). This will almost 

certainly lead to some economic benefits associated with the health and wildlife effects discussed above. 

Several studies have attempted to quantify the economic benefits associated with the reduction of mercury 

emissions from other sectors. To some extent, their methods and findings may also pertain to reductions 

from the dental sector. 

 Hagen, et al. (1999) conducted a study of the economic benefits of reduced mercury emissions in 

Minnesota. They administered a mail survey and presented respondents with four options for a 

50% reduction in emissions: 1) in Minnesota, 2) in the Midwest, 3) in the United States, and 4) in 

the world. They found an annual household willingness-to-pay (WTP) was $118.91, $212 million 

statewide or $0.12 per person per day for reductions in the Midwest.

 Jakus, et al. (2002) evaluated the economic and health effects of mercury advisories on the 

Maryland portion of the Chesapeake Bay, specifically with regard to striped bass, which is of 

primary concern in that area. They estimated the welfare loss to recreational anglers, the welfare 

loss to consumers and producers of commercial striped bass, and the health benefits to 

recreational anglers from avoided consumption. They found that the annual welfare loss to 

recreational anglers over all Maryland saltwater fishing days was $8.83 million (in 2000 dollars), 

and that the annual consumer and producer surplus losses from commercial fish consumption 

advisory were $215,800 and $304,500 respectively. Using an estimated value of a statistical life 

of $3.11 million, they calculated that the value of health damages associated with male mortality 

(between the ages of 40 and 59 only) from mercury consumption were approximately $14 million 

annually. They also evaluated the economic impacts of IQ losses due to prenatal exposure to 

mercury from fish consumption. However, the authors did not find a significant IQ loss based on 

consumption of striped bass from Maryland’s portion of the Bay.

 Trasande, et al. (2005) estimated the economic impact of anthropogenic mercury exposure in the 

United States 2000 birth cohort, and calculated the fraction of the cost that may be attributed to 

emissions from the electric sector. They utilize the “environmentally attributable fraction” (EAF), 

which is the portion of disease that would be eliminated if environmental risk factors were 

reduced to their lowest feasible level. They evaluated the cost per case based on lifetime lost 

productivity of affected infants, which they consider to be very conservative (versus other 

methods such as willingness to pay). According to their results, the loss of IQ from exposure to 

methylmercury of anthropogenic origin is $8.7 billion annually (in 2000 dollars), with a range 

between $2.2 billion and $13.9 billion with conservative assumptions. With less conservative 

assumptions, their estimates range up to $43.8 billion annually. Restricting the analysis to effects 

attributable only to United States anthropogenic sources results in a loss of IQ worth $3.1 billion 
18

annually, or up to $15.8 billion. This analysis includes consideration only for IQ loss from 

exposure in-utero, not for other possible effects such as cardiovascular impacts in adults.

 U.S. EPA (2005b) conducted an economic analysis of regulations to reduce mercury emissions

from power plants in the United States under the Clean Air Mercury Rule, using a model that 

provides information about changes in fish tissue mercury concentrations based on changes in air 

deposition. The authors assume that imported fish (accounting for 35% of consumer demand) is 

not likely to be substantially affected, but that methylmercury from domestic freshwater 

recreational and commercial fish is likely to be reduced, that about 16% of the U.S. population 

participates in recreational fishing, that annual demand for fishing product is 42.1 pounds per 

person, that 88% of people eat fish or shellfish at least once per month, that 26% of fish tissue 

samples contain higher than 0.3 ppm mercury, that 5% of samples have concentrations above 1 

ppm, and that there is an IQ decrease of 0.13 points for every ppm increase in maternal hair 

mercury concentration. The study quantifies the change in IQ among children exposed in-utero as 

a result of the regulation, but states that restricting the analysis to this endpoint is likely to 

underestimate the benefits. They found that a child of a typical freshwater fisher lost 0.06 to 0.07 

IQ points due to mercury exposure during 2001, that the child of subsistence farmers lost 0.331

points, and that each IQ point was worth about $8,800.

 Rice and Hammit (2005) assigned economic values to two mercury-related health effects: 

cognitive abilities and cardiovascular events. Their model assumes that changes in the quantity of 

mercury deposited led to linear and proportional changes in fish methylmercury concentrations, 

that one lost IQ point was worth $16,500 (2000 dollars), and that children born to mothers above 

EPA’s RfD lose an estimated average 0.77 IQ points. Based on a cost-of-illness approach, they 

found that the total cost of IQ losses per annual birth cohort was about $19.9 million, with 

359,000 children born to mothers with mercury levels above EPA’s reference dose, affecting their 

year of schooling, workforce participation, and wages. The authors also estimated that the 

lifetime cost of a myocardial infarction is between $37,030 and $83,290 depending on the age 

and sex of the affected individual. 

 Hylander and Goodsite (2006) quantified remediation costs and compared them with preventative 

measures and potential health effects in the Arctic, where they state that about 3⁄4 of children are 

born with unsafe mercury concentrations, even though that region has few or no mercury 

emission sources of its own. Their analysis was restricted to the cost of remediation at 6 sites, the 

cost of preventing and controlling emissions, food damage costs, and loss of IQ in Greenland.

They based their IQ valuation analysis on Trasande, et al. (2005) and studies of mercury 

concentrations in mothers and newborns in Greenland in 1998 and 1999. They estimate that the 

cost of IQ loss due to mercury toxicity in Greenland is about $59.1 million every year, and note 

that effects are more intense with a higher proportion of deficiencies, but with a lower overall 

population compared with the United States.

Each of these studies could contribute to a full economic analysis of the benefits associated with potential 

reductions of mercury discharges from the dental studies. For example, Hagen et al. (1999) provide 

estimates for WTP for reducing mercury emissions. The results of this study could be used to develop a 

benefit transfer approach to valuing reductions in mercury emissions from crematoria. Likewise, Jakus, et 

al. (2002) and Rice and Hammit (2005) provide economic values for some of the more well-documented 

health effects of methylmercury, which could be used in a more detailed analysis of health benefits from 

reduced mercury discharges. 

Trasande, et al. (2005) is the most relevant and easily applicable to estimating the benefits associated with 

reduced discharges of mercury from the dental sector. The authors calculated the entire economic cost of 

IQ loss among children resulting from mercury exposure in-utero, and then calculated the fraction 
19

attributable to the electric sector. This approach may also lend some insight into the potential IQ-related 

benefits of reductions from the dental sector, if the environmentally-attributable fraction (EAF) approach 

were transferred across sectors. Under conservative assumptions, Trasande, et al. (2005) estimated that 

United States anthropogenic mercury emissions result in $3.1 billion loss in IQ annually (in 2000 dollars). 

Applying the EAF of emissions from dental offices of 0.4% (from Exhibit 2-2) results in a total annual

cost of $12.4 million from IQ losses associated with methylmercury exposure in-utero from dental offices 

alone (in 2000 dollars). This could represent potential benefits from reduction of mercury discharges from 

dental offices.

However, there are many uncertainties associated with this estimate. For example, the 0.4% EAF used is 

based on data from 1994 and 1995, and as such is likely to be outdated and not directly pertinent to 

today’s share of discharges, especially given that many of the larger sources shown in Exhibit 2-2 have 

implemented controls since then. This may result in an underestimate of potential IQ-related benefits if 

the actual EAF for dental amalgam is higher. This estimate also does not consider other health-related 

benefits such as reduced cardiovascular disease or neurological impacts among adults. Additionally, 

implementation of a separator requirement (or equivalent) will not result in a total elimination of 

discharges from dental offices, and as mentioned above it is unclear where and when benefits would 

occur.

In addition to reduced mercury impacts to human health and wildlife, reduction of mercury in dental 

wastewater could have additional benefits such as reduced treatment costs and permit exceedances for 

POTWs. For example, after Massachusetts began requiring installation of amalgam separators in January 

2006, loadings of mercury to the Deer Island Treatment Plant decreased significantly (Rex, 2006).

Massachusetts Water Resource Authority data show that daily influent loading decreased from 0.55 

lb/day in 2002 to 0.43 lb/day in 2006, and effluent loadings decreased from 0.1 lb/day to less than 0.05 

lb/day during the same period.

The Security Sanitation District (District) in El Paso, Colorado is another example (Garcia, 2009). That 

treatment plant was issued a discharge permit for mercury in 2003, limiting effluent concentrations to 

0.044 ug/L until 2006, and to 0.011 ug/L thereafter. In 2006, the District began to systematically violate 

the permit. As such, the Colorado Department of Public Health and the Environment issued a Notice of 

Violation and Cease and Desist order to the District, ordering it to research and address sources of 

mercury. Since dental offices were the primary source of mercury to the District, they were required to 

install separators by September 2006 and initiate specific maintenance, recordkeeping, and other best 

management practices in 2007. Since 2008, the District has consistently been compliant with the 

standards, and the installation of separators by contributing dentists was clearly a significant factor in that 

success.
20

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25

Appendix A: State and Municipal Dental Mercury Reduction Programs

Exhibit A-1 summarizes 11 state and 8 municipal programs aimed at reducing dental mercury discharges through separator requirements. 

Exhibit A-1. State and Municipal Dental Mercury Reduction Programs

State or

Municipality Regulation Date Requirements or Additional Details Documented Reductions Notes Citation

Connecticut Sections 22a-616(d) 

and 22a-622 CGS 10/23/2003

Requires dentists to 

implement BMPs and 

install separators; CT DEP 

performs inspections

CT DEP, 2006

Louisiana

Act 126 of the 

2006 Louisiana 

Legislature

2007 Requires installation of 

amalgam separators

Pre-2007 ADA BMPs 

are required for 1,586 

dentists in the state. 

All should be in 

compliance based on 

Act 126.

ADA, 2007b; 

LDEQ, 2007

Maine CHAPTER 301

H.P. 514 - L.D. 697 5/15/2003

Any dental office, placing, 

removing, or repairing 

dental amalgam must 

install an amalgam 

separator by 2005

Maine DEP reviewed preand post-amalgam 

separator requirement 

mercury levels in sewage 

sludge and did not find as 

significant as those 

observed in a Boston 

study.

Quicksilver 

Caucus, 2008b

Massachusetts 310 CMR 73.00 3/1/2006

Implement BMPs and 

install separators which 

operate at 95% removal 

efficiency

From 2004-2006, amalgam 

separator use increased to 

>80%, and mercury levels 

in MWRA sludge 

decreased by about 48%.

Voluntary program 

than became 

mandatory in 2006

Quicksilver 

Caucus, 2008b

New Hampshire

House Bill 1251 & 

PART Env-Ws 905 

(RSA 485-A:4)

5/28/2005

Implement BMPs and 

install separators which 

operate at 95% removal 

efficiency

Allows recycling --

20.43 lbs recycled 

2006-2007

NHDES, 

undated
26

Exhibit A-1. State and Municipal Dental Mercury Reduction Programs

State or

Municipality Regulation Date Requirements or Additional Details Documented Reductions Notes Citation

New Jersey N.J.A.C. 7:14A-1 10/1/2007

Implement BMPs and 

install separators which 

operate at 95% removal 

efficiency (by 2009)

No information on 

reductions, but dental 

facilities contribute 35-

45% of mercury entering 

NJ's treatment plants, and 

this rule is expected to 

result in ~2,550 lbs of 

mercury removed from the 

environment each year

NJDEP 

Mercury 

Workgroup, 

2009

New York 6 NYCRR Subpart 

374-4

5/12/2006

Require recordkeeping and 

certified separators which 

operate at 99% removal 

efficiency

--

Requires ~15,000 

dentists to install 

separators by 2008 

Quicksilver 

Caucus, 2008b

Oregon

Senate Bill 704, 

revised HB 3611 

(679.520)

1/1/2008 Implement BMPs and 

install separators by 2011

No data available -- recent 

regulation

Estimate of 1,900 

facilities impacted

ODA, undated; 

House 

Oversight 

Subcommittee, 

2010

Rhode Island RI GLC Section 1. 

Chapter 23-24.9 2/26/2008

Implement BMPs and 

install separators which 

operate at 99% removal 

efficiency by 2008

-- RIDEM, 2006

Vermont 10 V.S.A. § 7110 2005

Implement BMPs and 

install separators which 

operate at 95% removal 

efficiency by 2007

--

Vermont 

Statutes Online, 

2005

Washington

WAC 173-303-

071(3)(a)(ii)(D) 

(existing hazardous 

waste rule)

MOU-2003

Implement BMPs and 

install amalgam separations 

which operate at 95% 

removal efficiency by 2005

Statewide fish tissue 

monitoring data showed a 

drop of mercury levels in 

monitored fish 

populations. Drop is too 

small with too few data 

points to call it a trend

~2,600 dental offices 

in Washington state 

are impacted. 

Voluntary memo of 

understanding became 

mandatory

Quicksilver 

Caucus, 2008b; 

Washington 

State 

Department of 

Ecology, 2003
27

Exhibit A-1. State and Municipal Dental Mercury Reduction Programs

State or

Municipality Regulation Date Requirements or Additional Details Documented Reductions Notes Citation

Minneapolis/St. 

Paul, MN N/A 2003

Requires certified 

separators which operate at 

99% removal efficiency

The mercury level in the 

influent has been reduced 

by approximately one half 

since the Program began.

Voluntary -- but those 

without separators 

have to apply for 

permit, test and meet 

stringent limit

Quicksilver 

Caucus, 2008b

King County, 

WA

King County 

Code—Title 28

Implement BMPs and 

install amalgam separations 

which operate at 95% 

removal efficiency (or 

perform BMPs and apply 

for permit)

King County realized a 

50% reduction in mercury 

in biosolids after 95% of 

its dentists installed 

amalgam separators and 

followed BMPs

Quicksilver 

Caucus, 2008b; 

KCIW, 

undated; KC 

DNRP

Boulder, CO 11-1-3 and 11-

3-24 3/12/2007

Implement BMPs and 

install amalgam separations 

which operate at 95% 

removal efficiency

--

Over 90 dental offices 

in the Boulder-area 

impacted

City of Boulder, 

2007

Fort Collins, CO 3/1/2004

All dentists in the city 

must install amalgam

separator

30% reduction Dube, undated; 

CDPHE, 2004

San Francisco, 

CA Order No. 01-072 7/1/2001

Implement BMPs and 

install amalgam separations 

which operate at 95% 

removal efficiency

Collection system sampling 

found >90% reduction in 

total Hg, biosolids 25-32% 

reduction

Studies have shown 

that approximately 

61% of the mercury 

entering WWTP is 

from dental practices.

Dube, undated; 

City of San 

Jose, 2010

Cleveland, Ohio 

(Northeast Ohio 

Region Sewer 

District)

Administrative 

Order No. 8021

Develop, submit and 

implement BMPs, 

including separators with 

99% efficiency by 

November 2011

--

Includes 518 dental 

clinics and 14 hospitals

Foley, 2010; 

NACWA, 

undated

East Bay 

Municipal Utility 

District 

(EBMUD), CA

40 CFR 403.12 (l) 11/1/2001 Implement BMPs and 

install amalgam separators

Data indicate a 30% 

reduction of mercury 

loading to EBMUD’s 

wastewater treatment plant 

as a result of these efforts

Has removed over 

1,200 lbs of waste

NACWA, 

undated; 

EBMUD, 

undated
28

Exhibit A-1. State and Municipal Dental Mercury Reduction Programs

State or

Municipality Regulation Date Requirements or Additional Details Documented Reductions Notes Citation

Milwaukee, WI 

(Milwaukee 

Metropolitan 

Sewage District)

11.214 1/26/2004

Implement BMPs and 

install amalgam separators 

by 2007

--

NACWA, 

undated; 

MMSD, 2004