ENVIRONMENTAL IMPACTS OF EMISSIONS
AND ASH FROM BIOSOLIDS INCINERATION
INTRODUCTION
Incineration and pyrolysis are thermal-chemical processes that reduce the amount (weight
and volume) of waste and sterilize solids. (Pyrolysis is similar to incineration, except that it
involves burning in the absence of oxygen; this process typically occurs in a fluidized-bed
reactor. Unless noted, this summary uses the term “incineration” to describe both processes.)
Incineration also produces energy, emissions, and ash. The quality of the incineration
byproducts is related to the quality of the raw materials, operation of the incinerator, and the
volatility or combustibility of contaminants in the raw materials. Incineration is a partial
biosolids disposal method only, because the ash requires proper disposal. The mass of fly ash
produced by incineration is usually between 15 to 45 percent of the mass of the dry biosolids
burned (Greenberg et al., 1981).
Potential environmental impacts of and concerns regarding incineration include: 1) air
pollution from emissions, 2) water pollution from boiler blowdown and scrubber water
discharge, 3) decreased effluent quality from recirculating loads, and 4) ash handling and
disposal (Dewling, 1980). Denison et al. (1987) even suggested that incinerators could
process trash into hazardous waste, because studies have shown that ash can release metals
and hazardous organics into the environment. Nevertheless, incineration of biosolids remains
an important component of biosolids management programs in the United States.
Metals are neither created nor destroyed by incineration; their amounts in biosolids before
incineration are identical to those present in the sum of emissions, scrubber water, and ash.
Incineration will, however, change the concentration of metals by destroying (combusting)
the organic portion of the biosolids (reducing mass). Incineration may also change the form
of metals, because complexes are broken down and new ones are created.
To meet specific health and regulatory goals for stack emissions, incinerators are
increasingly equipped with pollution abatement equipment such as fabric filters, highefficiency electrostatic precipitators (ESPs), and scrubbers. Such devices increase the amount
of ash and cause both the concentration and leachability of several trace metals in the fly ash
to increase.
The following is a summary of the content of various literature on the environmental
impacts of emissions and ash from biosolids incineration. The summary is divided into four
major sections: 1) system design; 2) regulations; 3) emissions; and 4) ash. For emissions, the
literature reviewed was limited to biosolids incinerators only; the section on ash includes
literature on municipal incinerators, because the properties and characteristics of ash are
similar regardless of the source of the ash. Abstracts of the articles that are referenced in this
NBMA Summary—Incineration
summary are included and are available for further review by contacting the Northwest
Biosolids Management Association.
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NBMA Summary—Incineration
SYSTEM DESIGN
The type of incinerator and the pollution abatement equipment at an incinerator both
influence the quality of incineration byproducts. This section describes the types of
incinerators and pollution abatement equipment found in the United States.
Incinerators
Approximately 200 biosolids incinerators operate in the United States. Most are multiplehearth incinerators, while most of the remaining incinerators are fluidized-bed incinerators.
Multiple-hearth incinerators are seldom installed currently, because fluidized-bed incinerators
are less expensive to build and operate, more flexible to operate, and can meet emission
requirements without an afterburner (Hemphil, 1988).
Multiple-hearth incinerators were originally designed for processing mineral ores and
have been used on biosolids since the 1930s (Dykes et al., 1987). The multiple-hearth
incinerator is a staged process consisting of 4 to 12 hearths. The upper hearths are used for
biosolids drying, combustion occurs in the central hearths, and ash cooling takes place in the
lower hearths. An afterburner is sometimes used at the exit of the furnace to reduce
hydrocarbon emissions and eliminate odors. The multiple-hearth incinerator produces
demister effluent, bottom ash, stack emissions, and a wet scrubber slurry, which yields a fly
ash and scrubber water (Fradkin et al., 1987).
Fluidized-bed incinerators were originally developed to recover catalysts from oil
refining. The first fluidized-bed biosolids incinerator was installed in Lynnwood, Washington
in 1961 (Dykes et al., 1987). In 1988, 70 fluidized-bed incinerators burned biosolids in the
United States (Hemphil, 1988). A fluidized-bed incinerator has a single combustion chamber
in which biosolids drying and combustion occur simultaneously. Combustion occurs in two
zones within the chamber: 1) water evaporates and pyrolysis of organics occurs in the bed,
and 2) free carbon and combustible gases are burned in the freeboard area above the bed
(Dykes et al., 1987). Because its combustion products exit though the top of the furnace, a
fluidized-bed incinerator does not produce bottom ash (Fradkin, 1987). Fluidized-bed
incinerators operate best at peak capacity and are easily shut down and restarted (Hemphil,
1988). Therefore, they can be operated efficiently without continuous operation.
System design of incinerators has been influenced by the development of biosolids
dewatering techniques, which reduced the energy required to evaporate the water from the
biosolids. However, as pretreatment of biosolids can reduce fuel requirements, it usually
requires increased energy in other unit processes (Dewling, 1980).
Incinerators are designed to be autogenous; auxiliary fuel, usually natural gas or fuel oil,
is used only for startup; burning the organics in the biosolids keeps the fires going. Coincineration (burning biosolids with solid waste) is another means of reducing auxiliary fuel
needs. Simple methods of reducing energy requirements through waste heat recovery include
using waste heat to preheat incoming air. Waste heat can also be used for heating systems or
to produce electricity. Waste heat recovery is economically attractive for communities over
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NBMA Summary—Incineration
200,000 that co-burn biosolids with municipal solid waste. A community of this size
produces approximately 76,000 m3 day-1 wastewater, 82 Mg day-1 biosolids, and 450 Mg day1 burnable refuse.
The importance of incinerator maintenance and efficient burning is apparent from a case
reported by Wall (1985). The City of Indianapolis, Indiana installed $250,000 worth of
instrumentation to improve the efficient operation of its incinerator and has since saved
$1,000,000 in fuel and an undetermined amount in operation and maintenance costs annually.
The instrumentation helped to regulate excess air and temperatures in the incinerator and
improve operation. The incinerator operated at 30 percent of maximum allowable particulate
emissions without additional pollution abatement equipment, as compared to 400 percent
before the instrumentation was installed. This reduction in emissions was realized merely by
improving incinerator efficiency. Incinerators in Nashville, Tennessee, Hartford, Connecticut,
and Buffalo, New York are using similar strategies to improve fuel efficiency.
Pollution Abatement Equipment
The type of pollution abatement equipment used in incinerators is important in
determining emissions. Bennett and Knapp (1982) studied four incinerators and found that
emissions of cadmium (Cd) were far greater at the incinerator with a wet single-pass scrubber
system. This type of system is considerably less efficient for the removal of submicron
particles than other scrubber systems. This is particularly important because Cd emissions are
predominantly associated with small particles.
Before 1978, only 20 percent of all multiple-hearth incinerators were equipped with
Venturi/impingement-tray scrubbers, but 14 out of 17 new incinerators built since 1978 used
this technology, including a Venturi scrubber followed by impingement-tray scrubber/cooler
compartment (Dykes et al., 1987).
ESPs and wet scrubbing devices control particles, and wet alkaline scrubbers control acid
gases. However, volatile arsenic (As), Cd, lead (Pb), and mercury (Hg) adsorb on fine
particles that are difficult to capture (Frost, 1988). A study by Takeda and Hiraoka (1976)
found that scrubber water contained relatively high concentrations of trace metals. However,
the total amount of metals was quite small compared with the total mass balance. The
particulate collection efficiencies of the scrubber for the vaporized trace metals were 0.6 to
2.7 percent for Cd, 0.6 to 1.6 percent for Pb, 0.9 to 2.3 percent for manganese (Mn), 3.9 to
4.7 percent for copper (Cu), and 1.8 to 85 percent for zinc (Zn). Thus, the trace metals were
scarcely caught in the scrubber water.
REGULATIONS GOVERNING INCINERATION AND ASH DISPOSAL
Incineration of biosolids and the disposal of incinerated biosolids ash are covered by
various state and federal regulations.
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NBMA Summary—Incineration
Incineration
Biosolids incineration must meet the requirements in the federal regulation 40 CFR Part
503 Subpart E, with the following exceptions:
•
If municipal solid waste accounts for more than 30 percent (by dry weight) of the
mixture of biosolids and auxiliary fuel (for example, wood chips, coal, municipal
solid waste, grit, scum, and screenings) following incineration, the process is covered
by 40 CFR Parts 60 and 61.
•
Nonhazardous incinerator ash generated during the firing of biosolids is not covered
by 40 CFR Part 503 when it is used or disposed. It must be disposed according to the
solid waste disposal regulations in 40 CFR Part 258. However, if the ash is applied to
land or placed on other than a municipal solid waste landfill, 40 CFR Part 257 must
be followed.
•
Hazardous waste is not considered to be auxiliary fuel under 40 CFR Part 503. Thus,
an incinerator that burns hazardous waste with biosolids is considered to be a
hazardous waste incinerator and is covered by 40 CFR Parts 261 through 268.
Subpart E of the 503 rule covers requirements for biosolids incineration. §503.40 defines
Subpart E §503.40 (a) as applying to a person who fires biosolids in a biosolids incinerator,
to a biosolids incinerator, and to biosolids fired in a biosolids incinerator. §503.40 (b) applies
to the exit gas from a biosolids incinerator stack. Pollutant limits for seven metals—
beryllium (Be), Hg, Pb, As, Cd, chromium (Cr), and nickel (Ni)—are regulated by §503.43
(a) through (d). Total hydrocarbon concentration in biosolids emissions is covered in §503.44
(a) and (b). Management practices (§503.45), frequency of monitoring (§503.46), recordkeeping (§503.47), and reporting (§503.48) are also covered.
State and local regulations also apply to biosolids incineration and are often more
stringent or define biosolids differently than the federal regulation. These other regulations
can be obtained at the appropriate state biosolids permitting authorities.
In Europe, the Federal Republic of Germany is at the forefront of air pollution control for
incinerators, with regulations covering the principles of protection, precaution, minimum
residue formation, and waste heat recovery. Specific requirements include flue gas
temperature above 800˚ C (not possible in multiple-hearth incinerators) and oxygen content
of flue gas at 6 percent. In the United Kingdom, the Economic Commission Directive of 1984
requires incinerators to get approval during the design stage by the appropriate government
authority (Frost, 1988).
The type of thermal process can also affect emissions and regulatory compliance. Majima
et al. (1977) compared noncombustion pyrolysis and partial combustion pyrolysis to
incineration and found that HCN, NH3, HCl, and odor evolved in pyrolysis and could be
removed in secondary combustion or gas scrubbing. For both pyrolysis processes, the total
exhaust gas volume and total fuel required was 67 percent and 65 to 70 percent respectively
5
NBMA Summary—Incineration
of that required in incineration. Furthermore, waste scrubbing water was below Japanese
drainage water pollutant standards.
Testing is typically performed during smooth operations. Metal and organic emissions
during startup, shutdown, temperature excursions, feed rate changes, and downtime are rarely
tested, but are likely to exceed emissions during smooth operations and emission standards
(Dykes et al., 1987).
Disposal of Incinerated Ash
Disposal of incinerated ash from biosolids incineration is regulated by 40 CFR Parts 257,
258, or 261–268, as appropriate. Data from the EPA and other sources indicate that
hazardous levels, as defined by the EPA’s extraction procedure (EP) toxicity test, of Pb, Cd,
and other toxic substances have been found in ash. The EP toxicity test identifies the water
solubility of metals and the concentration of contaminants in the leachate. Regulations are
then based on the results of the test. Regulatory levels for dangerous waste are 5.0 to 500 mg
L-1 Pb and 1.0 to 100 mg L-1 Cd; the levels for extremely hazardous waste are greater than
500 mg L-1 Pb and greater than 100 mg L-1 Cd (Knudson, 1986). Washington and California
have designated fly ash as hazardous waste that must be disposed of under stricter regulations
(Denison et al., 1987). In Washington, fly ash is handled as a dangerous or extremely
hazardous waste, while bottom ash is excluded from hazardous waste designation (Knudson,
1986).
Knudson (1986) subjected fly and bottom ash samples from seven incinerators in the
United States to dangerous waste criteria tests. The tests showed that all five of the fly ash
samples were considered to be hazardous waste under federal standards and three samples
were considered to be extremely hazardous waste under Washington State standards. None of
the bottom ash samples were considered to be hazardous waste under federal standards;
however, under Washington State standards, four samples were considered to be dangerous
waste and one was extremely hazardous waste, due to the presence of metal carcinogens. The
two combined ash samples were considered to be dangerous waste due to metal carcinogens,
and one was considered to be hazardous waste, due to Pb concentration. The EPA and the
Environmental Defense Fund (EDF) have grown increasingly concerned with the lack of
regulatory compliance by incinerator operators concerning hazardous waste and ash disposal.
The EDF notified 104 municipal incinerators that they were required to test their ash to
determine if it is hazardous under EPA guidelines. The results of Pb and Cd levels from fly
ash and combined ash were reported (Denison et al., 1987). Once the testing was completed,
it was found that fly ash values exceeded acceptable Cd and Pb limits in all but one case,
according to the EP toxicity test. Cd ranged from 2.1 to 100 mg L-1, and Pb ranged from 2.5
to 230 mg L-1 in fly ash (acceptable test limits are 1.0 and 5.0 respectively), and 0.02 to 5.30
mg L-1 Cd and 0.5 to 31.0 mg L-1 Pb in combined ash (Denison et al., 1987).
The EDF recommends: 1) testing fly and bottom ash frequently and thoroughly; 2)
imposing strict handling, transportation, and disposal methods for all ash; and 3) listing fly
ash as a hazardous waste (Denison et al., 1987).
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NBMA Summary—Incineration
The EPA has decided not to publish the technical guidance for ash disposal until the risks
of ash are evaluated. The EDF currently has lawsuits pending against municipal incinerators
in Chicago and New York, alleging that incinerator ash from these facilities should be
considered to be hazardous waste even though the EPA has designated incinerator ash as
nonhazardous under the Resource Conservation and Recovery Act (RCRA).
EMISSIONS FROM BIOSOLIDS INCINERATION
Emissions from incinerators include particulates, water vapor, heat, CO, CO2, SO2, NOx,
hydrocarbons, volatile organics, and metals. The concerns over emissions from incinerators
include not only the immediate pollution having air quality impacts, but also eventual
deposition to plants and soils, leading to potential long-term detrimental effects to the
environment.
High temperature, humidity, and chlorides are corrosive air pollutants. High chlorine
content in biosolids also results in complexation of metals as metal chlorides, which are more
soluble in water than other metals (Denison and Silbergeld, 1988). Nitrogen oxides form in
most combustion reactions, but are not formed at high enough concentrations to pose
emission problems in temperatures below 982˚ C (Dewling, 1980).
Blown sand from fluidized-bed incinerators and particulates from both fluidized-bed and
multiple-hearth incinerators are erosive (Hemphil, 1988). Particulates above 20 microns can
be completely separated from gaseous emissions (Liao and Pilat, 1972). However, smaller
particles are not easily captured in pollution abatement equipment and thus enter the
environment in emissions.
Asbestos fibers have been found in biosolids of several cities. However, asbestos was
absent from all ash samples tested by Patel–Mandik et al. (1988), suggesting potential escape
in emissions. It is believed that temperatures of 550˚ C or higher dehydroxylate the asbestos
lattice, resulting in alteration or destruction of the mineral. Unfortunately, the toxicology
associated with altered asbestos structures is unknown.
Sulfur oxides are formed by combustion of pyritic and organic sulfur. Although
approximately 75 percent of total sulfur content in municipal biosolids is in the form of sulfur
oxides, this is usually not a problem in incineration, according to Dewling (1980). Acidic
gases (HCl and SO2) are formed in municipal solid waste incinerators, but can be removed
effectively by alkali wet scrubbers (Lowe, 1988; Frost, 1988).
Trace Metals, Particle Size, and Temperature
Emissions of toxic metals from biosolids incinerators can present a risk to human health
and the environment (Barton, 1991; Vancil et al., 1991). Barton (1991) examined the
behavior of metals in municipal incinerators to identify the mechanisms responsible for
metals emissions and to determine the effect of key operating parameters. The primary escape
mechanism identified for metals involved the vaporization of metals in the combustion zone
and subsequent condensation in the exhaust system. This study found that combustion
7
NBMA Summary—Incineration
significantly increased the concentration of toxic metals on submicron particles. Combustion
chamber temperatures were found to have a strong effect on metals emissions. No
relationship was observed between total particulate emissions and metals emissions.
Incineration breaks down the matrix binding trace metals in biosolids, and leaves metals
in more soluble and volatile forms (Denison and Silbergeld, 1988; Segall et al., 1992). These
metals either remain in ash in soluble forms or are lost from the incinerator in emissions. The
emissions of metals from biosolids incinerators are not solely dependent on the quality of the
feed biosolids; the emissions also depend on volatility, emission particle size distribution,
and incinerator operating temperature. Since metals are conserved, those metals not found in
ash must have escaped in emissions.
In a study by Stephens et al. (1972), metal losses (mass entering minus mass in ash) from
biosolids incineration ranged from 10 to 640 g Cd day-1, 64 to 499 g Cr day-1, 0 to 290 g Cu
day-1, 54 to 64 g Ni day-1, 54 to 145 g Pb day-1, and 163 to 3,541 g Zn day-1. However,
pollution abatement equipment was not described, and from the age of the study it can be
assumed that the equipment was less effective at that time than it is today.
Bennett and Knapp (1982) found that several elements (selenium [Se], vanadium [V], Cu,
Zn, Cd, tin [Sn], and Pb) were at higher concentrations in the particles emitted from four
incinerators compared to their concentrations in the biosolids feed. This enrichment is
attributed to the volatility of the elements and their association with extremely fine particles
that are not efficiently removed by scrubbers. A second study by Bennett et al. (1984) found
that the ratios of elements in emissions differed considerably from their ratios in biosolids,
apparently due to the volatility of some elements, notably Cd, Pb, and Zn. The elements with
the largest enrichment ratios (percent in emissions divided by percent in biosolids) were: Cd
(31), Zn (14), Pb (9), and sulfur (8). Because of their vapor pressures, Cr, Ni, Cu, Zn, or Pb
should not be emitted during pyrolysis when the temperature is less than 750˚ C. However,
metals with lower vapor pressures, such as Cd and Hg, can be emitted during pyrolysis or
incineration (Kistler et al., 1987).
The volume of unfiltered/unscrubbed particulate emissions varies from 4 to 110 mg kg-1
dry biosolids input, depending on the incinerator type and the biosolids composition
(Dewling, 1980). Bennett and Knapp (1982) found that particles emitted from three multiplehearth and one fluidized-bed incinerator had small average mass median diameters, ranging
from 0.28 to 1.1 µm, with a few larger than 2.0 µm; small particle size promotes dispersion.
Particles smaller than 10 µm in diameter are respirable and easily inhaled (Denison and
Silbergeld, 1988). In addition, higher concentrations of some metals are often found on small
particles, due to the larger surface area to mass ratio of small particles. Bostian et al. (1988)
observed that metal emissions are primarily associated with finer particles and noted that this
association was likely due to the greater surface area per unit volume (or mass) available for
adsorption or condensation on the smaller particles. This characteristic was demonstrated by
leaching experiments, in which the increased surface area of finer particles and the presence
of metals at or near the surface of such particles exposed to leaching mediums resulted in
greater release of metals (Denison and Silbergeld, 1988). In experiments with a pilot-scale
multiple-hearth incinerator, between 2 and 11 percent of the feed silver (Ag), Cr, Cu, iron
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NBMA Summary—Incineration
(Fe), Mn, and Ni were found on particulate emissions (Nichols Engineering and Research
Corporation, 1981).
The combustion temperature during incineration affects metal emissions. The range of
combustion temperature is limited on the low end by the need to eliminate odorous matter in
the flue gas, and on the high end by the melting temperature of the ash. Therefore, the
combustion temperature is usually maintained between 700 and 1,000˚ C (Takeda and
Hiraoka, 1976). Importantly, the combustion temperature and the volatility of the metal are
the critical factors in determining the amount of the metal emitted from the incinerator.
Although high temperature incineration destroys organic compounds, these high temperatures
increase the potential for greater metals emission rates (Gerstle and Albrinck, 1982).
Emissions of Specific Metals
Arsenic. There is limited data available on the fate of As during biosolids combustion.
However, As is an element of concern because it is a known human and animal carcinogen as
well as a neurotoxin (Denison, 1988). Gerstle and Albrinck (1982) believed As and its
compounds should be considered volatile because they may be emitted in larger amounts
with increased combustion temperature. As can form the volatile oxide As2O3 (Novak et al.,
1977). However, Bostian et al. (1988) did not find that temperature affected enrichment of As
in emissions. In mass balance studies, 26 percent of the As in the feed biosolids was lost to
stack gas, but it was not discharged to the atmosphere. Campbell et al. (1982) found that As
in the scrubber water accounted for all As lost to stack gas, indicating that As associates with
larger particles.
Cadmium. Because of its low vapor pressure, Cd must be considered a volatile metal
(Gerstle and Albrinck, 1982). Kistler (1987) studied biosolids pyrolysis and found that at
505˚ C Cd was retained in the ash; at 750˚ C it was essentially transferred to the off-gas. At
625˚ C, the percent of Cd in the gas or ash depended primarily on the residence time. When
the residence time was increased from 1 to 4 hours, the Cd remaining in the ash decreased
from 52 percent to 18 percent. Kistler (1987) concluded that “during the pyrolysis of
biosolids at T>600˚ C the transfer of Cd to the off-gas cannot be prevented.” He illustrated
the mechanisms causing this transfer using Cd carbonate as an example:
CdCO3 —> T >400˚ C —> CdO + CO2
CdO + Cchar —> T>625˚ C—> Cd(1) + CO
Cd(1) —> T> 625˚ C —> Cd (g)
The CdCO3 in the feed biosolids is decomposed to the oxide form and then reduced to the
metal. Thus, Kistler attributed the emission of Cd at temperatures exceeding 600˚ C to
physiochemical factors and not process design.
In pyrolysis experiments conducted by Stammbach et al. (1988), Cd emissions started
below 600˚ C and were promoted by small particle size and high turbulence in the fluidizedbed incinerator.
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NBMA Summary—Incineration
Takeda and Hiraoka (1976) observed Cd to be readily vaporized at 450˚ C, with a
residual ratio (the amount remaining in the ash to that in the biosolids feed) of 80 percent at
450˚ C and 25 percent at 800˚ C. The authors reported that since most Cd is not caught by the
pollution abatement equipment, it is lost in emissions. The enrichment of Cd in emissions
was highest at an incinerator with a wet single-pass scrubber system. In the scrubber inlet of
this incinerator, 72 percent of the Cd was associated with particles less than 1 µm in diameter
(Bennett and Knapp, 1982).
Cd emissions increased from 1.0 to 8.1 mg g-1 when incinerator temperature increased
from 650 to 950˚ C (Dewling, 1980). Campbell et al. (1982) also stressed that Cd
volatilization was a function of temperature with the loss of Cd from fly ash ranging from
25 percent at 760˚ C to 56 percent at 928˚ C in their studies. Bostian et al. (1988) also found
Cd enrichment but at insignificant levels.
Chromium. The potential for Cr to become volatile during biosolids combustion is low. In
Gerstle and Albrinck’s literature review (1982), the authors indicated that less than 1 percent
of the Cr in the feed biosolids is associated with fine particles or fume. Tests by Takeda and
Hiraoka (1976) showed that over 80 percent of Cr was maintained in ash. Bostian et al.
(1988) did not find temperature to have an impact on Cr enrichment in emissions. The
oxidation of Cr(III) to Cr(VI) does not occur in pyrolysis but it does in incineration (Majima
et al., 1977). However, analytical procedures can yield artificially high concentrations of
Cr(IV); sampling of hexavalent Cr without artifact formation and analysis of the resulting
samples (specifically for hexavalent Cr at low concentrations) was accomplished by Segall et
al. (1992).
Copper. Cu is considered to be nonvolatile at incineration temperatures. Takeda and
Hiraoka (1976) found that Cu vaporized very little up to 600˚ C; at 800˚ C the residual ratio
remained about 85 percent.
Lead. Pb is an element of concern because it is neurotoxic at very low concentrations and
incremental doses (Denison and Silbergeld, 1988). Although Pb is considered to be
nonvolatile, Pb chloride is much more volatile than elemental Pb or Pb oxide. During
cooling, Pb oxide is also known to quickly condense on particles (Gerstle and Albrinck,
1982). Pb emissions increased from 3.1 to 77 mg g-1 when incinerator temperature was
increased from 650 to 950˚ C, and Pb was associated with particles under 3 µm in diameter
(Dewling, 1980). Bostian et al. (1988) reported that Pb enrichment increased with increasing
temperatures. EPA research supports this finding (Dewling, 1980).
Manganese. Takeda and Hiraoka (1976) found that Mn did not vaporize between 450 and
500˚ C. The residual ratio was about 80 percent at 600˚ C and about 75 percent at 800˚ C.
Mercury. Hg is the most volatile of the metals of concern in biosolids (Gerstle and
Albrinck, 1982; Kistler, 1987). Hg is neurotoxic and is thought to be particularly damaging to
fetal development (Denison, 1988). Hg has a high vapor pressure of 0.16 Pa at 20˚ C,
accounting for a saturation concentration of 14 mg of Hg m-3 of air (Kistler et al., 1987).
Campbell et al. (1982) found that more than 99 percent of the Hg in the feed stream was
emitted from a multiple-hearth incinerator. The proportion of Hg recovered in the scrubber
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NBMA Summary—Incineration
was not determined. Greenberg et al. (1981) found that Hg was emitted in larger quantities
than any other toxic compound tested from a Laurel, MD, biosolids incinerator. In a New
Jersey fluidized-bed incinerator, approximately 98 percent of the Hg in the feed biosolids was
emitted to the atmosphere when combustion temperatures averaged 788˚ C. The study found
that only 0.4 percent was retained in the ash, and slightly more than 2 percent was found in
the scrubber water (Dewling et al., 1980). In experiments of pyrolysis at 700 and 900˚ C, and
incineration at 900˚ C, less than 1 percent of the Hg could be accounted for in the ash
(Nichols Engineering and Research Corporation, 1981). Kistler (1987) found that, at varying
pyrolysis temperatures (350˚ C, 505˚ C, 625˚ C, and 750˚ C), more than 97 percent of the Hg
was emitted.
Nickel. Not much is known about the fate of Ni during incineration. Although elemental
Ni is nonvolatile, some Ni salts tend to become volatile at incineration temperatures (Gerstle
and Albrinck, 1982). Bostian et al. (1988) found no evidence that temperature affects
enrichment of Ni in emissions. Segall et al. (1992) found that, within the detection limit of a
wet chemical method, no nickel subsulfide was present in emissions.
Silver. Few investigations have looked at the fate of Ag during incineration or pyrolysis.
However, one experiment found that 60 percent of the Ag was retained in the ash after
incineration, while only 20 percent was retained in the ash after pyrolysis. Retention seemed
to be unaffected by combustion temperatures of 700 or 900˚ C (Nichols Engineering and
Research Corporation, 1981).
Zinc. The vapor pressure of Zn indicates that it can be volatile at combustion temperatures
(Gerstle and Albrinck, 1982). Takeda and Hiraoka (1976) found that Zn vaporized slightly
between 450 and 500˚ C. The residual ratio decreased to 70 percent at 600˚ C and to
60 percent at 800˚ C. However, only 28 percent of the Zn was associated with particles less
than 1 µm in diameter (Bennett and Knapp, 1982).
Organics and Complete Combustion
Hydrophobic hazardous organic compounds (for example, certain pesticides, polynuclear
aromatic hydrocarbons, and polychlorinated biphenyls) have been shown to partition onto
biosolids during the wastewater treatment process. While control of acid gases (HCl and
SO2), trace metals, and particles depends on pollution abatement equipment, control of CO,
NOx, and organics is achieved through furnace design and operation. For instance, furnaces
designed and operated to achieve nearly complete combustion have low CO emissions
(Dewling, 1980).
Odors are caused by the formation of partially oxidized hydrocarbons (Dewling, 1980).
Wastewater odors, such as those caused by the breakdown of proteins to amines during
anaerobic digestion, are eliminated by incineration (Liao and Pilat, 1972). However, more
stable organic compounds associated with biosolids are more difficult to destroy completely.
Volatile organic emissions are generally much lower from fluidized-bed incinerators than
from multiple-hearth incinerators. Emissions of some volatile compounds appear to be
11
NBMA Summary—Incineration
products of incomplete combustion (Bostian et al., 1988). Takeda and Hiraoka (1976) also
reported that pyrolysis can produce some hydrocarbons, as well as sulfur oxides and nitrogen
oxides. In pyrolysis experiments, the amount of gas formed increased with increasing
temperatures (Stammbach et al., 1988). Kaminsky and Kummer (1989) reported pyrolysis
experiments in which the gaseous portion of the pyrolysis products increased with rising
temperature from 22.7 percent at 620˚ C to 40.8 percent at 750˚ C. The pyrolysis gas was
analyzed and found to contain hydrogen, methane, ethane, ethene, propene, carbon monoxide,
carbon dioxide, and small amounts of other gases including C1–C4 hydrocarbons. In other
studies, the amount of ash and gaseous products derived from pyrolysis depended on the
temperature and heating rate (Kistler et al., 1987).
Principal Organic Hazardous Constituents. Mazer et al. (1987) investigated the potential
emissions of organic compounds during biosolids incineration. Thermal decomposition
experiments were conducted on biosolids spiked with mixtures of hazardous organic
compounds, on the mixtures of pure compounds in the absence of biosolids, and on unspiked
biosolids. The findings of this study suggested that the thermal decomposition behavior of
principal organic hazardous constituents (POHCs) differed when these compounds were
incinerated as mixtures in a biosolids matrix compared to mixtures of pure compounds. The
biosolids did not have an effect on the incinerability of easily destructed POHCs. The
biosolids appeared to decrease the incinerability of POHCs of intermediate thermal stability
(most of the compounds tested), making these POHCs more stable in biosolids. The biosolids
increased the incinerability of very stable POHCs, making them more readily destroyed in the
presence of biosolids. However, the authors also concluded that interactions between the
biosolids and POHCs, such as chlorinated contaminants, may also pose additional health
risks in incineration of biosolids from industrial sites.
Polychlorinated Dibenzo–p–dioxins and Dibenzofurans. Polychlorinated dibenzo–p–dioxins
(PCDDs) and polychlorinated dibenzofurans (PCDFs) are products of combustion. Evidence
from sediment cores suggests that PCDDs and PCDFs have significantly increased since
1940, which may be related to increased organohalide production in the United States.
PCDDs and PCDFs were first found in the fly ash and flue gas of municipal incinerators in
1977 (Fries and Pausenbach, 1990).
The formation of PCDDs and PCDFs from combustion depends on feedstock
composition, combustion facility design, and operational parameters such as temperature and
air supply. PCDDs and PCDFs were commonly found in incineration fly ash and emissions
but only found in trace amounts in a coal-fired power plant. In a study performed by Chiu et
al. (1983), all PCDDs and most PCDFs were associated with particles; PCDFs with low
molecular weight were found in impingers and XAD (Amberlite XAD–2) adsorbers.
(Amberlite XAD–2 is a chemical compound that can be utilized in the measurement of
PCDDs and PCDFs through extraction techniques.)
Clement et al. (1987) investigated the emissions of chlorinated organics from a 12-hearth
incinerator. They found that chlorinated dibenzo–p–dioxins and dibenzofurans were formed
during incineration, although at lower concentrations than in municipal incinerators.
Incinerators tested by Jones et al. (1987 in Fries and Pausenbach, 1990) emitted from 0.12 to
12
NBMA Summary—Incineration
9.23 ng m-3 (mean 1.75 ng m-3) PCDDs and PCDFs under the standard temperature and
pressure .
Testing for PCDDs, PCDFs, and 2,3,7,8,–TCDD was performed on 33 municipal
incinerators and 17 final disposal sites. Variations in the quantities found depended upon the
type of incinerator and operating conditions at the plants (Hiraoka et al., 1987). The results
are summarized in Table 1.
TABLE 1. PCDDs, PCDFs, and 2,3,7,8,–TCDD Test Results of Municipal Incinerators and Final Disposal
Sites (Hiraoka et al., 1987)
Incinerator Refuse
PCDDs
PCDFs
TCDDs
Fly ash
2.7–10,700 ng g-1 (average 1.0–1,510 ng g-1
523)
(average 150)
ND–14.5 ng g-1 (average 523)
(2,3,7,8,–TCDD
ND–4.7 ng g-1)
Bottom ash
5.0–2,470 ng g-1 (average
216)
2.3–864 ng g-1
(average 81.9)
0.3–7.3 ng g-1 (average 1.6)
(2,3,7,8,–TCDD ND–0.050
ng g-1)
Exhaust gas
133–13,600 ng m-3N,
(average 510)
436–10,000 ng m-3N
(average 2,520)
ND–109 ng m-3N
(average 54.9) (2,3,7,8,–
TCDD ND)
Discharged wastewater
ND–46.7 ng L-1
ND–285 ng L-1
ND–9.7 ng L-1
Discharged water
ND–143 ng L-1
ND–384 ng L-1
ND–11.5 ng L-1 (2,3,7,8,–
TCDD ND)
Leachate
ND–34.0 ng L-1
ND–210 ng L-1
ND–13.7 ng L-1
Ground water
ND–8.6 ng L-1
ND–3.9 ng L-1
ND
Surface soil
3.2–84.6 ng g-1 (average
22.4)
1.1–447 ng g-1
(average 76.4)
ND–1.8 ng g-1
(average 1.3)
Final Disposal Site
Edgerton (1989) found that municipal solid waste and biosolids incinerators were major
potential sources of ambient air concentrations of PCDDs and PCDFs. However, given
current knowledge of the health impacts of PCDDs and PCDFs, the low ambient air
concentrations were thought to pose no appreciable risk to public health.
Concentrations of TCDD in soil are determined by accumulation, mixing, persistence of
the compound, and uptake (Fries and Pausenbach, 1990). TCDD remains near the surface on
untilled pasture, but is mixed to 15 cm in cropland. TCDD is least persistent (stays in the soil
for the least amount of time) when applied in a solution or as vapor to the soil surface and is
not incorporated (di Domenico et al., 1982 in Fries and Pausenbach, 1990). TCDD neither
leaches nor becomes volatile (Jackson et al., 1985 in Fries and Pausenbach, 1990).
Plant uptake of TCDD is negligible. Experiments that found concentrations in plants were
probably finding volatilization and redeposition (Fries and Pausenbach, 1990). Little is
13
NBMA Summary—Incineration
known about how TCDD impacts fruits and vegetables. The highest interception factors are
for leafy vegetables (Baes et al., 1984 in Fries and Pausenbach, 1990), with lettuce and
tomatoes having the highest concentration in consumer products. The bulk of TCDD can be
removed by washing. For example, for carrots, 97 percent of TCDD was removed by washing
and peeling—68 percent by washing alone. Since 80 to 90 percent of roots and tubers are
peeled before consumption, the health impact of ingesting these plants was less than expected
from the whole vegetable analysis (Pennington, 1983 in Fries and Pausenbach, 1990).
Exposure of animals to TCDD from incineration varies with the type of feed (Fries and
Pausenbach, 1990). Animals feeding on roughage (for example, hay) can be exposed to
TCDD, while animals feeding on concentrates (high energy feed, typically seed) are less
likely to be exposed. Seeds are generally protected from direct deposition. Poultry is
generally free of TCDD because seeds are protected and the birds are physically protected in
houses (North, 1984 in Fries and Pausenbach 1990).
Polycyclic Aromatic Hydrocarbons. Concentration of polycyclic aromatic hydrocarbons
(PAHs) depends on combustion conditions; more complete combustion forms fewer PAHs.
Halogenated hydrocarbons in incineration emissions were below detection limits (Knudson,
1986).
Air pollution from continuously operating incinerators with complete combustion is far
less than from batch-type incinerators (Kamiya and Ose, 1987a). This is related to poor
control over operating conditions during startup and shutdown. For instance, during a 4-hour
normal operation of an incinerator, PAHs did not become volatile or decompose (Kamiya and
Ose, 1987b). Kamiya and Ose (1987a and 1987b) found that benzo[a]pyrene discharged in
emission gas and fly ash was nine times higher from the spray tower than in wastewater.
Mutagens in emission gas from complete combustion processes were 10 percent of that for
incomplete combustion processes. Ten percent of the mutagens in incinerator’s emission gas
was removed by spray tower water and treated as effluent (Kamiya and Ose, 1987b). Both
studies by Kamiya and Ose confirmed that there were more mutagens in emission gas than in
fly ash and bottom ash.
Saxena and Schwartz (1979 in Kamina and Ose, 1987b) argued that wastewater treatment
methods were unable to adequately remove potentially hazardous substances; and Shinohara
et al. (1980 in Kamina and Ose, 1987b) found that aeration caused PAHs with less than three
fused rings to became volatile after 10 hours while those with more than four fused rings
decomposed very slowly. The work of Kamiya and Ose (1987b) found results similar to that
of Shinohara et al.; specifically, they stated, “...PAHs in wastewater cannot be decomposed
by activated sludge using normal aeration times (4 to 5 h).”
Polychlorinated Biphenyls. Polychlorinated biphenyls (PCBs) in the atmosphere over the
United States are estimated conservatively at 18,000 kg. The source of PCBs to the
atmosphere has not been determined. However, landfills and municipal waste incinerators are
possible, untested sources. Multiple-hearth incinerators are designed to allow the volatile and
evaporative compounds to exit the hearth before combustion. These compounds are then
captured in air pollution abatement equipment or exit through the stack (Murphy et al., 1985).
14
NBMA Summary—Incineration
Incinerator emissions of PCBs are estimated by Murphy et al. (1985) to average 0.23 kg yr1/stack. This amount of PCBs is minor compared to the total amount of PCBs in the
atmosphere above the United States.
Dispersion of Emissions
Studies on the dispersion of emissions were more prevalent for municipal incinerators
than for biosolids incinerators. Thus, a few observations are noted here.
Greenberg et al. (1981) compared the relative emissions from a biosolids incinerator with
those of other large sources, such as coal-fired plants, municipal incinerators, and Pb from
motor vehicles. The authors concluded that, on a per capita basis, the emissions from the
incinerator were very small compared to other sources. However, the authors noted that,
because the incinerator stack was short, the temperature of the stack emissions was not far
above the ambient temperature, and since there were homes within a few hundred meters of
the incinerator, the impact of the emissions may have been much greater on the nearby area
than in the cases of other sources with greater total emissions but taller stacks. Thus,
residents living near a biosolids incinerator could receive considerably greater exposure than
in areas impacted by emissions from a coal-fired plant (Greenberg et al., 1981).
Berlincioni and di Domenico (1987) tested the emissions of tetra- to octachlorinated
PCDDs and PCDFs and resulting soil contamination from a municipal solid waste incinerator
in Florence, Italy. Most topsoil sampling sites in the study area were within 1 km of the
incinerator. The maximum PCDD soil content found was 7 x 104 ng m-2 of the soil surface.
The more chlorinated congeners generally exhibited levels remarkably higher than the less
chlorinated ones.
Accumulation on plants is related to plant biomass, density, leaf area, roughness, and
other factors (Baes et al., 1984), with decreased retention on the plants as the particle size
increases (Fries and Pausenbach, 1990). Incinerators with high exhaust temperatures have
more volatile 2,3,7,8,–TCDD in vapor form; however, the TCDD is less likely to deposit on
soil and plants because it is degraded by ultraviolet radiation (Fries and Pausenbach, 1990).
Kukkonen and Raunemaa (1984) investigated the dispersion of emissions from a
municipal solid waste incinerator. Out of 17 elements studied, they found the concentrations
of 11 elements on birch leaves varied strongly as a function of distance from the incinerator.
However, elemental concentrations on grass samples contained more fluctuations and the
distance dependency was less clear.
In another study (Stephens et al., 1972), foliage of trees in the vicinity of two biosolids
incinerators was examined. In the trees with the higher emissions, Pb was always present,
although its concentration varied with distance and direction from the incinerator and the
highway. Although loss of other metals through emissions exceeded the loss of Pb, only
traces of these metals could be found on the foliage; the cause of this phenomenon was
unknown. Pollution abatement equipment was not described in this study. However, it can be
15
NBMA Summary—Incineration
assumed by the age of the study that the equipment was less effective at the time of the study
than it is currently.
INCINERATOR ASH
Incineration yields odorless, sterile bottom ash and fly ash, which have a combined
weight of 10 percent to 60 percent of the original dewatered biosolids cake (Fraser and Lum,
1983; Furr et al., 1979; Denison et al., 1987). Ash typically contains concentrations of trace
metals that are four times those in dried biosolids; some of the metals may be at
concentrations above 1 percent of the mass of ash (Dewling, 1980). The composition of ash
is not well documented. Variations in design, pollution abatement equipment, and operating
conditions affect ash characteristics.
The cost of disposing incinerator ash is high. Landfilling in sanitary landfills cost from
$1.00 to $38.00 per 0.907 mg (with an average cost of $12.79) in 1987 (Repa, 1987). With
the closing of many landfills, these costs are increasing dramatically. Other ash disposal
practices are more expensive but are often required by state and possibly federal regulations.
Nineteen states regulate ash disposal and three require special handling (Governmental
Advisory Associates Inc., 1986, in Repa, 1987).
Trace Metals in Ash
Levels of trace metals in ash can be highly variable. Combustion systems and air
pollution control equipment affect the chemical form of metals in bottom ash and fly ash. No
correlation was found between extractable and total metals in ash (Wiles, 1988). A study by
Sawell et al., (1988) of two Canadian energy-from-waste facilities, which differed in design
and pollution abatement equipment, found that concentrations of Cd, Cu, Pb, and Zn were
low to moderate in bottom ash. They found that boiler and fly ash samples contained high
trace metal concentrations. Bierman et al. (1994) found that concentrations of Cu and Zn in
plant tissue were consistently higher and Mn was consistently lower when incinerator ash
was used as a fertilizer than when phosphate (P) fertilizer was used. Ash also significantly
increased tissue Cd and molybdenum (Mo) in sweet corn. However, trace metal
concentrations in tissue were not in the range that would suggest phytotoxicity or animal
health problems.
Some research has tested the feasibility of metal recovery. Two methods of minimizing
trace metals in biosolids and ash are to remove metals from the source or to recover them
from the biosolids or ash. Oliver and Carey (1976) investigated recovering metals from ash
by using H2SO4 or HCl as an acid extraction procedure to lower the pH to 1.5 and then using
vacuum filtration to separate the ash from the leachate. The ash was then disposed of on
agricultural land. Results showed that, on average, 50 to 80 percent of the Cd, Cr, Ni, and Zn
leached, while the Cu and Pb did not. Over 75 percent of the P, 30 percent of the N, and 10 to
15 percent of the organic carbon also leached. Oliver and Carey attempted a second approach,
using H2SO4 or HCl as an acid extraction procedure and then using a two-stage neutralization
to precipitate the metals. Results showed that P, Cd, Cr, Ni, and Zn were recoverable, but
66 percent of the Fe leached, which made recovery of P and trace metals more difficult. In an
16
NBMA Summary—Incineration
attempt to alleviate this problem, a NH4OH–(NH4)2SO4 method used in ore processing was
used to recover the metals; this proved unsuccessful. Oliver and Carey concluded that the
acid extraction procedure was prohibitively expensive and separation of metals from ash was
more economically feasible for large operations.
Organics in Ash
Dioxins and other organics associate with the small respirable particles in gas emissions.
However, levels of all organics tested were very low in the biosolids incinerator bottom ash
examined by Clement et al. (1987).
Polychlorinated Dibenzo–p–dioxins and Dibenzofurans. Organic compounds created during
combustion include dioxins and furans. Total dioxins in municipal solid waste incineration
ash ranged from 20 to 20,000 ppb, with 0.1 to 8 ppb of 2,3,7,8,–TCDD (Denison et al.,
1987).
TABLE 2. Concentrations of Total and Leachable Toxic Metals in Municipal Solid Waste Incinerator Ash
in the United States (Wiles, 1988)
Compound
Bottom Ash (mg
kg-1)
Bottom Ash
Leachate (mg L-1)
Fly Ash
(mg kg-1)
Fly Ash
Leachate (mg L-1)
Pb
31 to 36,600
0.02 to 34
2 to 26,600
0.019 to 53.35
Cd
0.18 to 100
0.018 to 3.94
5 to 2,210
0.025 to 100
As
0.8 to 50
ND(0.001) to .0122
4.8 to 750
ND(0.001 to 0.858)
Cr
13 to 1,500
ND(0.007) to 0.46
21 to 1,900
0.006 to 0.135
Barium (Ba)
47 to 2000
0.27 to 6.3
88 to 9000
0.67 to 22.8
Ni
ND(1.5) to 12,91
0.241 to 2.03
ND(1.5) to 3,600
0.09 to 2.90
Cu
40 to 10,700
0.039 to 1.19
187 to 2,300
0.033 to 10.6
( ) = below detection limits
The bioavailability of dioxins in fly ash is inversely related to the organic carbon content
of the ash. Therefore, better combustion and stack controls may yield ash with a greater
dioxin exposure. Bioaccumulation of PCDDs and PCDFs from municipal solid waste
incinerator fly ash in Carp (Cyprinus carpio) was tested by Kuehl et al. (1987). Previous
studies showed a strong preference for 2,3,7,8,–TCDD bioaccumulation, but the
concentration was not directly related to the concentration in the original incinerator fly ash.
Ingestion of fly ash from sediments was thought to be the primary route of carp exposure.
Preferential accumulation occurred for compounds with chlorine in the 2,3,7, and 8 positions.
The Bioavailability Index (contaminant in fish to contaminant in fly ash) decreased with
increasing chlorination (Kuehl et al., 1987).
Polycyclic Aromatic Hydrocarbons. Wszolek (1982) investigated the presence of PAHs in
incinerated biosolids ash. PAHs are known to be a product of combustion, and some PAHs
are carcinogenic and mutagenic. PAHs and other trace organics have been found in the fly
17
NBMA Summary—Incineration
ash from incinerated municipal solid waste. Wszolek found a wide range in the amount (0.1
to 1.0 µg g-1) and variety of PAHs. No clear correlation was found between the amount of
PAHs and the industrial input into the wastewater of the cities studied.
Polychlorinated Biphenyls. Few studies have investigated the presence of PCBs in
incinerated biosolids ash. However, Furr et al. (1979) reported finding no PCBs in
incinerated biosolids ash of 10 cities in the United States. This is possibly due to thermal
decomposition or condensates that have become volatile on escaping particulates.
Comparison of the Impact of Pyrolysis and Incineration on Ash Content
The type of plant and pollution abatement equipment affects the characteristics of the ash.
This review focuses on the differences between pyrolysis and incineration.
Pyrolysis. In pyrolysis experiments, Kaminsky and Kummer (1989) found that trace
metals were more strongly incorporated into the matrix of pyrolysis ash than in incinerated
ash or biosolids. They found that 70 percent of the Cd, 100 percent of the Cr, 95 percent of
the Cu, 87 percent of the Ni, 92 percent of the Pb, and 89 percent of the Zn that were present
in the feed biosolids remained in pyrolysis ash. The authors concluded that, with the
exception of the volatile metals (Hg and Cd), trace metals originating from the biosolids were
enriched in pyrolysis ash. For Cd, the pyrolysis temperature and residence time were
significant, while Hg was already volatile at temperatures below 500˚ C.
Kistler et al. (1987) studied both raw and digested biosolids and found that Cr, Ni, Cu,
Zn, and Pb were completely retained in pyrolysis ash up to the maximum investigated
temperature of 750˚ C. The studies classified the metals as follows:
•
Cr, Ni, Cu, Zn, and Pb are nonvolatile under pyrolysis conditions up to 705˚ C
•
Cd is volatile under pyrolysis conditions at temperatures exceeding 600˚ C
•
Hg is volatile under pyrolysis conditions at temperatures exceeding 350˚ C
The studies also found that the amount of metals leached from pyrolysis ash was very
small, with concentrations low enough to be better than drinking water standards. The
authors concluded that the high H+ buffering capacity and the alkaline nature of the slaked
chars rendered the metals in the char very immobile. Thus, the pyrolysis char appeared to be
suitable for disposal in a sanitary landfill. However, since these metals might be mobilized by
a combination of acids and organic liquids, they should be disposed of in landfills where
organic material is excluded (“monofill”).
The results of pyrolysis experiments in a multiple-hearth incinerator indicate that
lowering furnace temperature from 900 to 700˚ C improves metal retention in the ash from
50 percent to 90 percent for Be, from 10 to 30 percent to 50 percent for Cd, and from 30 to
50 percent to 60 percent for Pb (Nichols Engineering and Research Corporation, 1981).
18
NBMA Summary—Incineration
Incineration. It is important to note that incineration and pyrolysis can yield different
emissions and ash quality. Few studies have examined this issue, although one study found
that incineration retained 60 percent of the feed Ag in the ash, whereas pyrolysis retained
only 20 percent. Cr, Cu, Fe, Mn, Ni, and Zn were entirely retained in the ash regardless of the
mode of operation (Nichols Engineering and Research Corporation, 1981).
Dewling et al. (1980) described the distribution of trace metals in a fluidized-bed
biosolids incinerator. The percent of each metal studied found in the ash, scrubber water, and
stack emissions are shown in Table 3.
TABLE 3. Normalized Distribution of Trace Metals at a FluidizedBed Incinerator (Percent of Weight)
Compound
Ash
Scrubber
Stack Emissions
Zn
79.0
20.0
1.0
Cu
78.0
21.0
1.0
Pb
87.0
12.0
1.0
Cr
95.0
4.0
1.0
Ni
80.0
20.0
ND
Hg
0.4
2.0
97.6
Cd
80.0
20.0
ND
ND = below detection limits
Leachate
Liquids (water or other weak solvents) passed through ash will extract soluble and weakly
held metal ions and other dissolvable compounds. This is often used as a method to assess the
availability of a particular contaminant to plants or animals, or the potential of the
contaminant to pass through the soil into the groundwater. Availability of metals in leachate
decreases with increasing incinerator temperature (Oliver and Carey, 1976), which makes it
difficult to compare different studies.
A study of ash from incinerated New Jersey municipal solid waste found that Cd and Pb
leached from the fly ash in excessive concentrations (15 to 18 ppm for Cd, 27 to 44 ppm for
Pb), creating a serious disposal problem. Furthermore, this problem was magnified by
increases in acid concentrations of the leaching medium (Clapp et al., 1988). di Casa et al.
(1982) also found excessive leaching of Cd, Cu, Ni, and Pb from ash from a municipal
incinerator in northern Italy and suggested this ash was the source of groundwater
contamination.
Behel et al. (1986) investigated various amendments to reduce the solubility of Cd and Pb
in incinerator ash. Pb solubility was reduced by amendments of CaO, CaCO3, acidic
NH4H2PO4, basic K2HPO4, and elemental Se. Cd solubility was reduced by CaO and the P
19
NBMA Summary—Incineration
amendments. Malone and Jones (1985) also investigated three stabilization treatments and
found that the addition of lime to fly ash to form a cement-like solid proved to be the most
durable treatment and reduced leaching to the greatest extent. However, none of the treated
products were very durable.
In leaching studies with soil amended with fly and bottom ash from municipal
incinerators, it was found that Cd, and, to a lesser extent Pb, were much more mobile than if
similar concentrations of these elements had been applied as inorganic salts or in biosolids
(Giordano, 1983). Repa (1987), using the EP toxicity test, also found high concentrations of
Pb and Cd in leachate. All other metals were below regulatory limits, and all organics tested
were below regulatory limits. Fraser and Lum (1983) investigated the availability of elements
in incinerated biosolids ash and found that above 90 percent of the Cd, Cr, Cu, Ni, Pb, and Zn
were in unavailable forms. The measurable, but small, amounts of aluminum (Al), Cr, Mn,
Ni, Pb, and P present in the ash could be released, however, if the ash were to come into
contact with ice or snow melt contaminated with road salt or acid rain.
Only a small fraction of metals were recovered in leachate by Klei et al. (1981); cobalt
(Co), Pb, and Fe were undetectable. It was suggested that acid conditions of at a pH of 4.0
would have increased leaching two- to threefold. It is believed that calcium (Ca) and Cr could
be recovered from incinerator ash.
A study of incinerated biosolids ash found that, once initial leaching of trace metals
occurred, leaching rates were very slow. Fe was found to leach in the highest concentration of
the metals studied, with leaching loss of more than 4 percent (by weight) of the initial Fe in
the ash. Other metals leached the following percentages of their initial concentrations in the
ash: Ca (2.30 percent), Mg (1.10 percent), and K (0.95 percent). Zn, sodium (Na), Mn, Cr,
Co, Pb, and Ni leached less than 0.20 percent (Sweeney, 1978).
Theis and Padgett (1983) looked at extractable trace metals from biosolids incinerator ash
and dewatered biosolids using the EP toxicity test. The test was based on a 5 percent ash in a
95 percent municipal solid waste landfill in which decomposition of municipal solid waste
produces acids. Repa (1987) found that ash used as a daily cover or mixed with municipal
solid waste is most likely to encounter acid leaching, while monofills are least likely. Theis
and Padgett (1983) found that elevated concentrations of Fe, Al, and Ca were common for
locations using FeCl2, AlSO4, or lime to condition biosolids, which affected EP toxicity test
results. Incineration of biosolids increased the leachability of all metals except Cd, and As if
Fe(III) or Al(III) salts were added. Sequential extraction (a series of leachings on the same
sample with different strength solvents) indicated that Cd and Pb were associated with the
amorphous Al oxides, and Fe, Cr, and Ni were associated with Fe oxyhydroxides. There was
little correlation between the amount of element released and the amount of the element in
bulk ash or biosolids.
Maskarinec et al. (1987) investigated laboratory extraction procedures for testing
leachate. In the study, lysimeters packed with municipal solid waste were treated with
distilled water to generate a municipal waste leachate. This leachate was then used to leach a
variety of industrial wastes, including resource recovery ash (the result of processes
20
NBMA Summary—Incineration
producing “energy-from-waste”), a biosolids incinerator ash mixture, and a coking plant
wastewater treatment sludge (the result of coal combustion processes). The resulting
leachates were analyzed for both inorganic and organic constituents. Inorganic constituents
showed two distinct types of leaching behavior: freely soluble elements (for example, B, Na,
and Ca) had leaching curves that showed exponential decline; acid-soluble elements (Ni and
Zn) eluded as “peaks” as the pH of the leachate decreased. Organic constituents gave
leaching curves that were dependent on solubility: either an exponential decline (freely
soluble compounds) or a relatively flat curve at or near the solubility limit.
Because of the possibility of acid leachate in municipal landfills, monofills are considered
to be a more appropriate means of disposal. Two studies considered the reactions that may
occur in monofills. Hjelmar (1988) tested bottom and fly ash from two Danish municipal
solid waste incinerators not equipped with scrubbers with a column leaching test. Solid
residue from municipal solid waste incinerators is typically 80 to 90 percent bottom ash and
10 to 20 percent fly ash. For most metals, increased leaching corresponded to increased fly
ash content of the mixture tested. Concentrations of metals leached in the column leaching
test were higher than concentrations in leachate from the incinerator ash monofill that had
been monitored for 15 years by landfill operators. The only exceptions were for SO42- and
Ca2+; these were highly saturated in the accelerated column experiments. Similarly, Krebs et
al. (1988) explored the main reaction of bottom ash with water, the long-term mobilization
potential of indicator elements from ash, and the feasibility of bottom ash monofills as a
means of long-term storage. Solutions were analyzed for Ca, Na, potassium (K), Al, OH-, Cl-,
SO42-, P, NO3-, NH4+, COD, Zn, Pb, Cd, and Cu. Ca2+, Na+, K+, and Cl- increased in solution
with increasing reaction time with bottom ash. OH- decreased in concentration after 50 hours
of reaction time, while other elements stayed constant or increased slowly. OH-, Cl-, and
Al(OH)4- dissolution occurred before dissolution of SO42-. During the first decade after
disposal, leaching should occur under alkaline conditions, with total organic carbon as a main
variable in the long-term behavior of bottom ash.
Health Effects
Exposure to dust, surface soils, reentrained particles, contaminated food, and drinking
water all constitute postdepositional exposure. The importance of postdepositional exposure
can be seen in major studies of Pb exposure, which indicate postdepositional exposure to Pb
is 5 to 50 times more intense than ambient air exposure.
Other exposure pathways involve ash handling and disposal, including: 1) handling,
storage, transport, and disposal of ash and other waste products of incineration; 2) exposure
to airborne dust, suspended particles, and contaminated soil from landfills; 3) direct ground
and surface water contamination by leachate; 4) erosion of landfill cap; and 5) transport of
metals out of the landfill by plant uptake and burrowing animals (Denison and Silbergeld,
1988; Denison et al., 1987). In addition, workers in biosolids incinerators can be exposed to
airborne irritants. Nethercott (1981) investigated an outbreak of dermatitis among workers at
a biosolids incinerator. The outbreak was caused when the workers replaced a malfunctioning
incinerator exhaust fan and were exposed to airborne irritants. The problems ceased when
skin contact with the irritants was minimized.
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NBMA Summary—Incineration
Stern et al. (1989) assessed the potential exposure levels and pursuant public health
implications of neighborhood exposure to a landfill which, from 1954 to 1973, received
bottom ash from a municipal incinerator. The incinerator processed residential and
commercial waste, but not industrial waste. Soil was sampled for 10 trace metals, PCDDs,
PCDFs, 2–3–7–8–TCDD and TCDF congeners, PCBs, and PAHs. The results of soil analysis
and modeling indicated that the Pb level at the landfill was considerably above the level
recommended by the Center for Disease Control and could lead to Pb poisoning in exposed
children. The health impact of exposure to other substances measured in the soil was
considered to be small, and no significant increased cancer risk was expected. It is important
to note that the incinerator that supplied the ash used no air pollution control devices to
collect fly ash, whereas most modern incinerators do. Fly ash and bottom ash are now
routinely co-disposed. Because the landfill received no fly ash, results from this study are not
necessarily applicable to disposal sites containing both fly and bottom ash.
Trace Metals. Trace metals exert a broad range of toxic effects, including carcinogenic,
neurological, hepatic, renal, hematopoietic, and other adverse effects (Denison and
Silbergeld, 1988). Metal emissions can have direct inhalation effects or secondary and
tertiary effects after fallout (Dewling, 1980).
Trace metals accumulate in the environment and in the human body. Pb is neurotoxic at
very low doses, and prenatal neurological effects occur at any level of Pb exposure. With Pb
exposure already a problem from automobile emissions, paint, and other sources, even small
increases in Pb exposure can drastically increase the incidence of Pb toxicity. Cd is also
neurotoxic and can damage the lung and kidneys. It is carcinogenic and may be toxic to
fetuses. Pb and Cd may act synergistically as toxins to bone. Hg is neurotoxic in organic
forms, and microorganisms in the environment can alkylate inorganic Hg into hazardous
organic forms. As is neurotoxic and carcinogenic. Hg, As, Cd, Zn, and Cu are toxic to aquatic
organisms. For an extensive review of the toxicology of trace metals see Friberg et al. (1986).
In addition, a comprehensive list of metallic carcinogens, designated by the International
Agency for Research on Cancer, appears as Appendix A in Knudson (1986).
Polychlorinated Dibenzo–p–dioxins and Dibenzofurans. Potential sources for human
exposure to TCDD in food is now believed to stem more from combustion and waste
disposal practices than phenoxy herbicide production (Fries and Pausenbach, 1990). TCDD is
not genotoxic (EPA, 1988a in Fries and Pausenbach, 1990); its half-life in human adipose
tissue is 7 years (Poiger and Schlatter, 1986 in Fries and Pausenbach, 1990; Pirkle et al., 1989
in Fries and Pausenbach, 1990).
Incinerators located in rural areas can potentially contaminate milk, meat, and vegetables
with TCDD (Fries and Pausenbach, 1990). A population near the incinerator source can
potentially receive 500 to 1,000 times greater exposure to TCDD from food than from
inhalation.
Using models of radioactive deposition, dispersion, and bioaccumulation, Fries and
Pausenbach (1990) reviewed TCDD bioaccumulation and routes of exposure, then estimated
TCDD deposition and transfer to the food chain from incinerators. The review found that soil
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NBMA Summary—Incineration
accumulation was less important than direct plant deposition and the forage–cattle–beef
pathway was more important than the forage–dairy cow–milk pathway in human exposure.
Animal food sources contained an estimated 10 to 40 fg kg-1 d-1 and plant food sources
contained an estimated 1 to 5 fg kg-1 d-1. Fries and Pausenbach concluded that exposure from
incinerators was insignificant compared to background sources.
The concentration of PCDDs and PCDFs in human milk samples is a possible route of
infant exposure. The possible risk from municipal solid waste incinerator emissions was
analyzed based on formulas presented by Fries and Pausenbach (1990). The maximum levels
of PCDDs and PCDFs from incinerators would account for 1 to 10 percent of the PCDD and
PCDF levels in milk; other sources of PCDDs and PCDFs must account for the additional
PCDD and PCDF found in milk (Smith, 1987).
The toxicity and toxic potential of fly ash can affect aquatic life. Helder et al. (1982)
assessed the impact of fly ash on rainbow trout yolk sac fry. Fry exposed to fly ash showed no
toxic response during the 60-day test, but fry exposed to raw extract (using toluene) and
cleaned extract (passed through a column of silica gel and acid-, base-, and silver nitratemodified silica gel) became inactive, stopped feeding, hemorrhaged, and finally died from
severe edema. Toxic response was similar to toxic response for 2,3,7,8,–TCDD, but only
minor quantities of 2,3,7,8,–TCDD were found in extracts. Response was attributed to a
combined effect of TCDDs and TCDFs.
Incineration of biosolids produces dioxins and is thought to make the highest contribution
to the general background. However, Fiedler and Hutzinger (1992) estimated that, besides
combustion processes, nonthermal waste treatment methods, such as composting, represent a
reservoir to the release of PCDDs and PCDFs in agriculture.
Polycyclic Aromatic Hydrocarbons. Gustavsson (1989) investigated mortality for workers
at a municipal incinerator in Stockholm, Sweden. Workers were exposed to PAHs, other
polycyclic compounds, dust, nitrous oxides, sulfur dioxide, carbon monoxide, and several
trace metals. The incinerator used in this study was an older incinerator that caused more
environmental contamination and occupational exposure than modern plants. The study
found that lung cancer risk was 3.5 times higher than the national rate, and 2 times higher
than the local rate. Although eight out of nine lung cancer deaths were attributed to smokers,
the rate increase could not be attributed solely to smoking. Metal exposure was thought to be
too low to cause lung cancer. Additionally, ischemic heart disease was higher than expected
by national and local standards. The association between PAH and lung cancer is well
documented, but the association of ischemic heart disease is not. Gustavsson (1989) further
suggested that exposure to PAHs, TCDDs, and TCDFs may relate to increased
atherosclerosis, but precise evidence is lacking.
Uses of Incinerated Ash
The expense of landfilling incinerator ash has led to research into alternate uses of
incinerated biosolids ash. Ash has been studied as a soil amendment, although there may be
problems with trace metal concentrations. Another option for highly contaminated ash has
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NBMA Summary—Incineration
been to use it as a cement filler, which would stabilize the hazardous constituents. Finally,
some form of solidification into a usable product may be possible.
Soil Amendment. Biosolids ash may be used as a soil amendment, particularly as it often
contains several essential macro- and micro-nutrients. Fitzpatrick (1985) investigated the use
of biosolids ash as a growing media for containerized tropical trees. He found that four of the
five species studied grew equally well in this media compared to a traditional growing media.
Rosen (1989) studied the use of biosolids incinerator ash as a source of P for the production
of corn. For early growth, at equivalent P rates, corn response was greater for the fertilizer
treatment than for the ash treatment. However, final yield was unaffected by treatment.
In comparing the fertilizer properties of biosolids and its incinerated ash, Jakobsen and
Willett (1986) found that biosolids were a better fertilizer. Although biosolids and ash were
both good sources of lime and, in the short-term, ash was more effective in raising soil pH,
the biosolids were more effective in raising extractable P. The biosolids were also a good
source of nitrogen (N) and P, and increased foliar concentrations of K, Se, Mg, Cu, and Zn.
The ash was not a source of N or P, and increased only Se and Mg. The ash also appeared to
decrease the availability of K, Cu, and Zn.
Some characteristics of incinerated ash can be detrimental to plant growth. Eis (1978)
studied revegetation problems at a landfill that contained ash from incinerated municipal
solid waste. Revegetation had low success because of the ash characteristics, which included
high porosity, low water-holding capacity, very low N and P content, and toxic
concentrations of Se, Zn, and Pb.
Although incinerated biosolids ash may sometimes be used as a soil amendment, Mellbye
(1982) noted that its use on agricultural land may be limited by the concentration and
availability of trace metals in the ash. Plant uptake and phytotoxicity vary considerably
between inorganic and organic sources of elements; thus, information from studies on metal
availability from soils amended with biosolids may not be wholly applicable to soil amended
with incinerated biosolids ash.
In experiments by Giordano et al. (1983), corn and Swiss chard were grown in
greenhouse pots of Decatur silt loam amended with fertilizer as well as 300 g of fly ash or
600 g of bottom ash. The chard displayed a positive growth response to both ash treatments.
The plants were tested for concentrations of Cd, Pb, Zn, Ni, Na, and Mn. Na, Cd, Pb, and
possibly Mn were found at levels considered to be excessive.
In experiments with corn, Rosen (1989) found ash to be a good source of Zn, which can
be limited at high rates of P. However, Zn was found to move through the soil profile. Other
trace metals such as Cd, Pb, Ni, and Cr did not move out of the top 15.24 cm, nor did they
accumulate in the corn grain, stover, or cob.
Bierman and Rosen (1994) evaluated biosolids incinerator ash as a P fertilizer on corn
grown in an Estherville sandy loam (sandy, mixed, mesic Typic Hapludol). Limited effects
on growth occurred the first 2 years, and significant increases in yield occurred in the third
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NBMA Summary—Incineration
and fourth years. The ash increased Cu, Cd, Mo, P, and Zn concentrations in plant tissue.
However, the increase did not reach phytotoxicity levels.
Furr et al. (1979) investigated the absorption of elements by cabbage growing on soil
amended with incinerated biosolids ash. One objective of the study was to determine if the
cabbage would absorb elements in proportion to their concentrations in the ash. Although Ni
and Zn were found in high concentrations in the ash, absorption by the cabbage was low,
possibly because these elements were present in the ash in insoluble compounds formed
during incineration. Fe, Mn, and Pb were higher in the control cabbage than in any of the
cabbage grown in the ash. These elements may have been present naturally in the soil but
were unavailable in the ash.
Cement Filler. Tay (1987) found that that biosolids ash could be used as a partial
replacement for cement. Lisk (1989) also investigated this use of biosolids ash and found that
some mixtures created a cement that met the specifications for compressive strength for
masonry cement. Lisk also found that the strength of the cement generally decreased with
increasing percentages of bottom ash; fly ash (which is more homogeneous and finer than
bottom ash) became more intimately mixed with cement particles, thus creating a stronger
product.
Solidification. In Japan, the use of incinerated biosolids ash as brick has been practiced as
an alternative to disposal (NEDO, 1994). Solidification of ash is possible, but results in a 5 to
50 percent volume increase, and the reagent materials may be prohibitively expensive.
However, these costs may be acceptable if the ash is classified as a hazardous waste. Another
option is to screen ash and dispose of larger, nontoxic particles as nonhazardous material
while solidifying the smaller particles; this might reduce costs as well as prevent leaching. If
solidified material is used for building materials, exposure to biota may represent a hazard.
Studies of breakwater concrete and attached marine life indicated no metal uptake, although
these results were preliminary (Wiles, 1988).
SUMMARY
Biosolids incineration is practiced in the United States and remains a viable biosolids
management alternative. However, incineration yields both emissions and ash. Potential
environmental impacts of and concerns regarding incineration include: 1) air pollution from
emissions; 2) water pollution from boiler blowdown and scrubber water discharge; 3)
decreased effluent quality from recirculating loads; and 4) ash handling and disposal.
Additionally, deposition onto the soil from emissions can gradually degrade soil quality.
Thus, regulations have been developed to minimize the risk to humans and the environment.
Modern incinerators are increasingly equipped with pollution abatement equipment, but such
devices increase the amount of ash and change its nature.
Emissions
Emissions from incinerators include particulates, water vapor, heat, CO, CO2, SO2, NOx,
hydrocarbons, volatile organics, and metals. Of the metals, Hg is the most volatile; As, Cd,
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NBMA Summary—Incineration
and Zn are considered to be at least partially volatile; Pb and Ni are not volatile except in salt
form, and Cu and Cr are generally retained in the ash. Toxic organics can also be found in
emissions, either originating from the feedstock (for example, PCBs and PAHs), or as a result
of the incineration process (for example, dioxins and dibenzofurans). Emission or production
of these organics is, again, dependent upon feedstock and operation.
Ash
Biosolids incineration yields an odorless, sterile ash, which typically contains trace metals
concentrations that are four times those in dried biosolids. Levels of trace metals in ash can
be highly variable. The combustion system and air pollution control equipment affects the
chemical form of metals in bottom ash and fly ash. In contrast, of the organics tested,
concentrations are usually very low in biosolids incinerator bottom ash.
Leachate from ash can be contaminated with soluble metals. Generally, availability of
metals in the leachate decreases with increasing incinerator temperature. However, acidic
conditions in leachate will increase concentrations of the metals. Thus, because of the
possibility of acid leachate in municipal landfills, monofills are considered to be more
appropriate means of disposal.
The expense of landfilling incineration ash has led to research into alternate uses of
incinerated biosolids ash. Ash contains both macro- and trace nutrients, can be used as a
liming material to raise soil pH, and studies have investigated its use as a soil amendment.
However, plant uptake and phytotoxicity vary considerably between ash types. Studies have
also investigated the use of ash as a cement filler and as a building material (when solidified).
Health Effects
Exposure to dust, surface soils, reentrained particles, contaminated food, and drinking
water all constitute postdepositional exposure; postdepositional exposure may be 5 to 50
times more intense than ambient air exposure. Other pathways of exposure involve ash
handling and disposal, including: 1) handling, storage, transport, and disposal of ash and
other waste products of incineration, 2) exposure to airborne dust, suspended particles, and
contaminated soil from landfills, 3) direct ground and surface water contamination by
leachate, 4) erosion of landfill cap, and 5) transport of metals out of the landfill by plant
uptake and burrowing animals. The incineration of biosolids are known to produce dioxins
and are thought to make the highest contribution to the general background.
Reference: Henry, C., and R. Harrison. 1998. Environmental Effects of Biosolids
Management. Trace Metals: Potential for Movement and Toxicity from Biosolids
Application, Effects on Wildlife and Domestic Animals from Biosolids Application, Air
Emissions and Ash Resulting from Incineration of Biosolids, Nitrogen Cycle and Nitrate
Leaching from Biosolids Application, Microbial Activity, Survival and Transport in Soils
Amended with Biosolids, The Fate of Trace Synthetic Organics in Biosolids Applied to
Soil, Runoff Water Quality from Biosolids Application, Effects of Organic Residuals on
Poplars. Northwest Biosolids Management Association.
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