EPA

Office of Air Quality

Pla nnin g a nd S ta nd a rds

C Air

lean

United States Office of Air Quality EPA-453/R-98-004b

Environmental Protection Planning and Standards February 1998

Agency Research Triangle Park, NC 27711

Air

Study of Hazardous Air Pollutant

Emissions from Electric Utility Steam

Generating Units -- Final Report to

Congress

Volume 2. Appendices
mm
TABLE OF CONTENTS (Continued)

Volume 2

Section Page

Appendix A Median Emission Factors, Determined from Test

Report Data, and Total 1990, 1994, and 2010

Emissions, Projected with the Emission Factor

Program ..................... . A-1

Appendix B Matrix of Electric Utility Steam-Generating Units

and Emission Test Sites ............. . B-1

Appendix C Listing of Emission Modification Factors for Trace

Elements Used In the Individual Boiler Analysis . . C-1

Appendix D Discussion of the Methodology Used to Develop

Nationwide Emission Totals ............ . D-1

Appendix E Health Effects Summaries: Overview ........ . E-1

Appendix F Documentation of the Inhalation Human Exposure

Modeling for the Utility Study .......... . F-1

Appendix G Data Tables for Dioxin Multipathway Assessment . . . G-1

Appendix H Literature Review of the Potential Impacts of

Hydrogen Chloride and Hydrogen Fluoride ..... . H-1

Appendix I Mercury Control Technologies ........... . I-1
mm
Appendix A - Median Emission Factors, Determined from Test Report

Data, and Total 1990, 1994, and 2010 Emissions, Projected with the

Emission Factor Program
X
Nationwide hydrogen cyanide emissions were determined from stack a

emission factors and not from EMFs.

Nationwide phosphorous emissions were determined from stack emission b

factors and not from EMFs.

A-1

Table A-1. Median Emission Factors, Determined from Test Report

Data, and Total 1990 and Total 2010 HAP Emissions, Projected with

the Emission Factor Program for Inorganic HAPs from Coal-fired

Units

Coal-fired units: inorganic Estimated total 1990 Estimated total 1994 Estimated total 2010

HAPs emissions (tons) emissions (tons) emissions (tons)

Antimony 7.95 7.98 9.93

Arsenic 60.93 55.81 70.61

Beryllium 7.13 7.93 8.20

Hydrogen chloride 143,000 134,000 155,000

Hydrogen cyanide (HCN) 240.66 250.8 318.31 a

Hydrogen fluoride 19,500 23,100 25,700

Cadmium 3.33 3.15 3.82

Chromium 73.27 61.60 87.43

Cobalt 21.21 22.67 27.08

Lead 75.47 61.77 86.89

Manganese 163.97 167.72 219.02

Mercury 45.80 51.34 59.74

Nickel 58.05 52.04 68.65

Phosphorus (P) 270.74 331.41 358.09 b

Selenium 153.83 183.68 213.21
Nationwide phosphorous emissions were determined from stack emission a

factors and not from EMFs.

A-2

Table A-2. Median Emission Factors, Determined from Test Report

Data, and Total 1990 and Total 2010 HAP Emissions, Projected with

the Emission Factor Program for Inorganic HAPs from Oil-fired

Units

Oil-fired units: Estimated total 1990 Estimated total 1994 Estimated total 2010

inorganic HAPs emissions (tons) emissions (tons) emissions (tons)

Arsenic 5.02 3.51 2.54

Beryllium 0.46 0.40 0.23

Cadmium 1.71 1.09 0.86

Chromium 4.74 3.91 2.40

Cobalt 20.41 15.48 10.31

Hydrogen chloride 2860 2100 1450

Hydrogen fluoride 143 284 73

Lead 10.58 8.92 5.35

Manganese 9.28 7.30 4.70

Mercury 0.25 0.19 0.13

Nickel 392.83 322.37 198.17

Phosphorus (P) 67.25 50.89 34.10 a

Selenium 1.65 1.42 0.84
The emission factors are not available for this compound, but the a

compound was detected in one or more tests. Some of these compounds

encompass a group of compounds, and although the total is not available,

some members of the group may be presented elsewhere.

A-3

Table A-3. Median Emission Factors, Determined from Test Report

Data, and Total 1990 and Total 2010 HAP Emissions, Projected with

the Emission Factor Program for Inorganic HAPs from Gas-fired

Units

Gas-fired units: Estimated total 1990 Estimated total 1994 Estimated total 2010

inorganic HAPs emissions (tons) emissions (tons) emissions (tons)

Arsenic 0.15 0.18 0.25

Cadmium a a a

Chromium a a a

Cobalt 0.14 0.15 0.23

Lead 0.43 0.47 0.68

Manganese a a a

Mercury 0.0015 0.0017 0.0243

Nickel 2.19 2.42 3.49

Phosphorus 5.65 6.23 8.98
A-4

Table A-4. Median Emission Factors, Determined from Test Report Data, and Total 1994 and Total 2010 HAP Emissions, Projected with the Emission Factor Organic HAPs from Coal-fired Units

Coal-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of emission Computer of factor: Computer Number of 

emission factor: program: emission lb/trillion program: emission 

factors lb/trillion Btu 1990 total factors Btu 1994 total factors 

Median Number emission 

Median 

1,1,2-trichloroethane 1 4.70 40.39 1 4.70 42.10 1 

2-chloroacetophenone 3 0.29 2.50 3 0.29 2.60 3 2,4-dinitrotoulene 3 0.015 0.130 3 0.015 0.130 3 Acetaldehyde 12 6.75 58.01 12 4.85 43.44 12 

Acetophenone 7 0.68 5.84 8 0.91 8.15 7 

Acrolein 6 3.25 27.93 7 3.3 29.56 6 

Benzene 20 2.50 21.48 22 2.50 22.40 20 

Benzyl chloride 1 0.006 0.050 1 0.006 0.050 1 Bis(2-ethylhexyl) phthalate 9 4.10 35.24 9 4.10 36.73 9 Bromoform 1 6.60 56.73 1 6.60 59.11 1 Carbon disulfide 8 4.30 36.96 9 2.90 25.97 8 Carbon tetrachloride 2 3.25 27.93 2 3.25 29.11 2 Chlorobenzene 2 3.18 27.34 2 3.18 28.48 2 Chloroform 2 3.20 27.50 3 3.00 26.87 2 Cumene 1 0.29 2.50 1 0.29 2.60 1 Dibutyl phthalate 5 2.8 24.07 4 2.55 22.84 5 

Ethyl benzene 5 0.41 3.52 6 0.41 3.63 5 Ethyl chloride 1 2.40 20.63 1 2.40 21.49 1 

Methyl chloroform 4 3.42 29.35 4 2.10 18.81 4 
Table A-4. (Continued)

Coal-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of emission Computer of factor: Computer Number of 

emission factor: program: emission lb/trillion program: emission 

factors lb/trillion Btu 1990 total factors Btu 1994 total factors 

Median Number emission 

Median 

A-5

Ethylene dichloride 3 3.10 26.64 3 3.10 27.76 3 

Formaldehyde 15 4.00 34.38 14 3.25 29.11 15 

Hexane 2 0.83 7.10 3 1.50 13.43 2 

Hexachlorobenzene 1 0.08 0.68 1 0.08 0.70 1 Isophorone 2 24.0 206.28 2 24.0 214.97 2 Methyl bromide 6 0.89 7.65 6 2.245 20.11 6 Methyl chloride 3 5.90 50.71 4 5.5 49.27 3 Methyl ethyl ketone 6 8.00 68.76 7 5.00 44.78 6 Methyl iodine 1 0.40 3.44 2 5.02 44.97 1 Methyl isobutyl ketone 3 4.90 42.12 3 4.90 43.89 3 Methyl methacrylate 1 1.10 9.45 1 1.10 9.86 1 Methyl tert-butyl ether 1 1.40 12.03 1 1.40 12.54 1 Methylene chloride 5 13.0 111.8 7 10.0 89.6 5 n-nitrosodimethylamine 1 0.68 5.84 1 0.68 6.09 1 Naphthalene 11 0.77 6.62 12 0.32 2.87 11 

m,p-cresol 2 0.675 5.80 2 0.675 6.05 2 o-cresol 3 1.7 14.61 3 1.7 15.22 3 

p-cresol 1 0.95 8.16 1 0.95 8.51 1 Perylene 1 0.075 0.65 1 0.075 0.67 1 Pentachlorophenol 1 0.008 0.07 1 Phenol 10 6.1 52.43 10 6.1 54.64 10 
Table A-4. (Continued)

Coal-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of emission Computer of factor: Computer Number of 

emission factor: program: emission lb/trillion program: emission 

factors lb/trillion Btu 1990 total factors Btu 1994 total factors 

Median Number emission 

Median 

A-6

Phthalic anhydride 1 4.9 42.12 1 4.9 43.89 1 

Propionaldehyde 4 10.35 88.96 4 10.35 92.71 4 Quinoline 1 0.053 0.45 1 0.053 0.47 1 Styrene 7 3.1 26.64 8 2.4 21.49 7 

Tetrachloroethylene 5 3.1 26.64 5 3.1 27.76 5 

Toluene 17 3.6 30.94 19 3.3 29.56 17 Trans-1,3-dichloropropene 1 4.7 40.39 1 4.7 42.10 1 

Trichloroethylene 1 3.1 26.64 1 3.1 27.76 1 

Vinyl acetate 1 0.42 3.61 1 0.42 3.76 1 Vinylidene chloride 2 9.7 83.37 1 6.7 59.69 2 

Xylenes 2 4.65 39.96 4 3.05 27.32 2 

o-xylenes 5 0.81 6.96 5 0.81 7.26 5 m,p-xylenes 8 1.45 12.46 9 1.2 10.75 8 

Total TEQ for 2,3,7,8-tetra- 17 9.7 x 10 17 1.2 x 10 17 chlorodibenzo-p-dioxin

-5 -4 

2,3,7,8-tetrachlorodi-benzo-p- 4 1.5 x 10 1.3 x 10 6 2.6 x 10 2.3 x 10 4 

dioxin

-6 -5 -6 -5 

1,2,3,7,8-pentachlorodi-benzo-p- 3 2.8 x 10 2.4 x 10 3 2.8 x 10 2.5 x 10 3 

dioxin

-6 -5 -6 -2 

1,2,3,4,7,8-hexachlorodi-benzo-p- 4 5.9 x 10 5.1 x 10 4 5.9 x 10 5.3 x 10 4 

dioxin

-6 -5 -6 -5 

1,2,3,6,7,8-hexachlorodi-benzo-p- 4 6.6 x 10 5.6 x 10 4 6.6 x 10 5.9 x 10 4 

dioxin

-6 -5 -6 -5 
Table A-4. (Continued)

Coal-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of emission Computer of factor: Computer Number of 

emission factor: program: emission lb/trillion program: emission 

factors lb/trillion Btu 1990 total factors Btu 1994 total factors 

Median Number emission 

Median 

A-7

1,2,3,7,8,9-hexachlorodi-benzo-p- 4 7.9 x 10 6.7 x 10 5 4.2 x 10 3.7 x 10 4 

dioxin

-6 -5 -6 -5 

1,2,3,4,6,7,8-heptachlorodi-benzo- 8 4.2 x 10 3.6 x 10 10 7.4 x 10 6.6 x 10 8 

p-dioxin

-6 -5 -6 -5 

Heptachlorodi-benzo-p-dioxin 6 7.6 x 10 6.5 x 10 6 7.6 x 10 6.8 x 10 6 

-5 -4 -6 -4 

Hexachlorodi-benzo-p-dioxin 7 2.7 x 10 2.3 x 10 7 2.7 x 10 2.4 x 10 7 

-5 -4 -5 -4 

Octachlorodi-benzo-p-dioxin 10 3.6 x 10 3.1 x 10 11 2.6 x 10 2.3 x 10 10 

-5 -4 -5 -4 

Pentachlorodi-benzo-p-dioxin 6 8.0 x 10 6.9 x 10 6 8.0 x 10 7.1 x 10 6 

-6 -5 -6 -5 

Tetrachlorodi-benzo-p-dioxin 9 8.8 x 10 6.8 x 10 9 8.8 x 10 7.1 x 10 9 

-6 -5 -6 -5 

2,3,7,8-tetrachlorodi-benzofuran 1 4.4 x 10 3.8 x 10 9 4.2 x 10 3.7 x 10 1 

-6 -5 -6 -5 

1,2,3,7,8-pentachlorodi- 1 4.6 x 10 3.9 x 10 5 4.6 x 10 4.1 x 10 1 

benzofuran

-6 -5 -6 -5 

2,3,4,7,8-pentachlorodi- 1 4.8 x 10 4.2 x 10 6 7.6 x 10 6.8 x 10 1 

benzofuran

-6 -5 -6 -5 

1,2,3,4,7,8-hexachlorodi- 1 7.9 x 10 6.8 x 10 6 7.3 x 10 6.5 x 10 1 

benzofuran

-6 -5 -6 -5 

1,2,3,6,7,8-hexachlorodi- 1 4.0 x 10 3.4 x 10 8 4.0 x 10 3.6 x 10 1 

benzofuran

-6 -5 -6 -5 

1,2,3,7,8,9-hexachlorodi- 4 6.8 x 10 5.8 x 10 5 6.3 x 10-6 5.6 x 10 4 

benzofuran

-6 -5 -5 

2,3,4,6,7,8-hexachlorodi- 5 1.2 x 10 1.0 x 10 6 1.4 x 10 1.3 x 10 5 

benzofuran

-5 -4 -5 -4 

1,2,3,4,6,7,8-heptachlorodi- 8 5.7 x 10 4.9 x 10 9 7.1 x 10-6 6.4 x 10 8 

benzofuran

-6 -5 -5 
Table A-4. (Continued)

Coal-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of emission Computer of factor: Computer Number of 

emission factor: program: emission lb/trillion program: emission 

factors lb/trillion Btu 1990 total factors Btu 1994 total factors 

Median Number emission 

Median 

A-8

1,2,3,4,7,8,9-heptachlorodi- 4 1.8 x 10 1.6 x 10 5 8.3 x 10-6 7.4 x 10 4 

benzofuran

-5 -4 -5 

Heptachlorodi-benzofuran 9 1.9 x 10 1.6 x 10 9 1.9 x 10 1.7 x 10 9 

-5 -4 -5 -4 

Hexachlorodi-benzofuran 8 2.1 x 10 1.8 x 10 8 2.1 x 10 1.9 x 10 8 

-5 -4 -5 -4 

Octachlorodi-benzofuran 9 1.7 x 10 1.4 x 10 10 1.9 x 10 1.7 x 10 9 

-5 -4 -5 -4 

Pentachlorodi-benzofuran 10 1.2 x 10 1.0 x 10 10 1.2 x 10 1.1 x 10 10 

-5 -4 -5 -4 

Tetrachlorodi-benzofuran 10 1.1 x 10 9.8 x 10 10 1.1 x 10 1.0 x 10 10 

-5 -5 -5 -4 

1-methylnaphthalene 3 0.011 0.098 5 0.011 0.102 3 2-chloronaphthalene 3 0.040 0.352 2 4.6 x 10 0.004 3 

-4

2-methylnaphthalene 6 0.032 0.275 9 0.03 0.275 6 Acenaphthene 8 0.013 0.108 9 0.017 0.154 8 Acenaphthylene 7 0.004 0.036 10 0.0057 0.052 7 Anthracene 7 0.004 0.039 10 0.005 0.047 7 Benz(a)anthracene 6 0.002 0.018 9 0.004 0.040 6 Benzo(a)pyrene 8 0.001 0.009 7 0.0015 0.013 8 Benzo(e)pyrene 4 0.001 0.012 6 0.002 0.022 4 Benzo(b)fluoranthene 1 0.008 0.069 2 0.005 0.048 1 Benzo(b+k)fluoranthene 3 0.004 0.036 4 0.004 0.039 3 Benzo(k)fluoranthene 1 0.004 0.031 3 0.001 0.011 1 Benzo(g,h,i)perylene 4 0.002 0.019 6 0.002 0.017 4 Biphenyl 4 0.18 1.562 6 0.073 0.649 4 
Table A-4. (Continued)

Coal-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of emission Computer of factor: Computer Number of 

emission factor: program: emission lb/trillion program: emission 

factors lb/trillion Btu 1990 total factors Btu 1994 total factors 

Median Number emission 

Median 

A-9

Chrysene 6 0.003 0.022 9 0.006 0.054 6 Dibenzo(a,h)anthracene 2 0.001 0.011 5 7.0 x 10 0.006 2 

-4

Fluoranthene 8 0.016 0.132 10 0.025 0.220 8 Fluorene 7 0.013 0.110 10 0.015 0.132 7 Indeno(1,2,3-c,d)pyrene 4 0.003 0.028 8 0.002 0.021 4 Phenanthrene 7 0.032 0.275 10 0.072 0649 7 Pyrene 7 0.012 0.106 10 0.013 0.121 7 
After a check was made of nondetect procedures used in 1990 and 2010 emissions, this 

a

did not pass the criteria for detection. Therefore, no 1994 emission number is avaicompound. Some of these compounds encompass a group of compounds, and although the available, some members of the group may be presented elsewhere.

A-10

Table A-5. Median Emission Factors, Determined from Test Report Data, and 

Total 1994 and Total 2010 HAP Emissions, Projected with the Emission Factor Organic HAPs from Oil-fired Units

Oil-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of factor: Computer Number of factor: Computer Number of 

emission lb/trillion program: emission lb/trillion program: emission lb

factors Btu 1990 total factors Btu 1994 total factors 

Median Median 

emission emission e

Acetaldehyde 1 8.2 5.0 1 8.2 3.8 1 

Benzene 6 1.40 0.86 7 1.00 0.46 6 Ethylbenzene 2 0.49 0.30 2 0.49 0.23 2 Formaldehyde 9 30.0 18.4 11 20.0 9.3 9 Methyl chloroform 3 7.6 4.6 3 7.6 3.5 3 

Methylene chloride 2 32.25 17.93 2 32.25 14.92 2 Naphthalene 4 0.34 0.21 10 0.61 0.28 4 Phenol 2 24.30 14.86 2 24.30 11.24 2 Tetrachloroethylene 1 0.55 0.34 1 0.55 0.25 1 Toluene 6 8.00 4.90 7 8.40 3.88 6 Vinyl acetate 2 5.15 3.15 2 5.15 2.39 2 o-xylenes 1 0.84 0.52 1 0.84 0.39 1 m,p-xylenes 2 1.35 0.83 2 1 .35 0.63 2 

Total TEQ for 2,3,7,8-tetra- 17 7.0 x 10 17 8.8 x 10-6 17 chlorodibenzo-p-dioxin

-5 

2,3,7,8-tetrachlorodi-benzo-pdioxin

a a 
Table A-5. (Continued)

Oil-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of factor: Computer Number of factor: Computer Number of 

emission lb/trillion program: emission lb/trillion program: emission lb

factors Btu 1990 total factors Btu 1994 total factors 

Median Median 

emission emission e

A-11

1,2,3,7,8-pentachlorodi-benzo-p- 2 4.0 x 10 2.4 x 10 2 4.0 x 10 1.9 x 10 2 4dioxin

-6 -6 -6 -6 

1,2,3,4,7,8-hexachlorodi-benzo- 2 9.9 x 10 6.1 x 10 2 9.9 x 10 4.6 x 10 2 9p-dioxin

-6 -6 -6 -6 

1,2,3,6,7,8-hexachlorodi-benzo- 1 8.2 x 10 5.0 x 10 1 8.2 x 10 3.8 x 10 1 8p-dioxin

-6 -6 -6 -6 

1,2,3,7,8,9-hexachlorodi-benzo- 1 9.6 x 10 5.9 x 10 1 9.6 x 10 4.5 x 10 1 9p-dioxin

-6 -6 -6 -6 

1,2,3,4,6,7,8-heptachlorodi- 1 5.9 x 10 3.6 x 10 1 5.9 x 10 2.7 x 10 1 5benzo-p-dioxin

-5 -5 -5 -5 

Heptachlorodi-benzo-p-dioxin 1 1.2 x 10 7.3 x 10 1 1.2 x 10 5.5 x 10 1 1

-4 -5 -4 -5 

Hexachlorodi-benzo-p-dioxin 2 8.2 x 10 5.0 x 10 2 8.2 x 10 3.8 x 10 2 8

-5 -5 -5 -5 

Octachlorodi-benzo-p-dioxin 2 1.4 x 10 8.4 x 10 2 1.4 x 10 6.4 x 10 2 1

-4 -5 -4 -5 

Pentachlorodi-benzo-p-dioxin 2 8.0 x 10 4.9 x 10 2 8.0 x 10 3.7 x 10 2 8

-5 -5 -5 -5 

Tetrachlorodi-benzo-p-dioxin 1 1.0 x 10 6.3 x 10 1 1.0 x 10 4.8 x 10 1 1

-4 -5 -4 -5 

2,3,7,8-tetrachlorodi-benzofuran 1 6.7 x 10 4.1 x 10 1 6.7 x 10 3.1 x 10 1 6

-6 -6 -6 -6 

1,2,3,7,8-pentachlorodi- 1 8.2 x 10 5.0 x 10 1 8.2 x 10 3.8 x 10 1 8benzofuran

-6 -6 -6 -6 

2,3,4,7,8-pentachlorodi- 1 5.9 x 10 3.6 x 10 1 5.9 x 10 2.7 x 10 1 5benzofuran

-6 -6 -6 -6 

1,2,3,4,7,8-hexachlorodi- 1 9.6 x 10 5.9 x 10 1 9.6 x 10 4.5 x 10 1 9benzofuran

-6 -6 -6 -6 

1,2,3,6,7,8-hexachlorodi- 1 4.4 x 10 2.8 x 10 1 4.4 x 10 2.1 x 10 1 4benzofuran

-6 -6 -6 -6 
Table A-5. (Continued)

Oil-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of factor: Computer Number of factor: Computer Number of 

emission lb/trillion program: emission lb/trillion program: emission lb

factors Btu 1990 total factors Btu 1994 total factors 

Median Median 

emission emission e

After a check was made of nondetect procedures used in 1990 and 2010 emissions, this 

a

did not pass the criteria for detection. Therefore, no 1994 emission number is avai

compound. Some of these compounds encompass a group of compounds, and although the available, some members of the group may be presented elsewhere.

A-12

1,2,3,7,8,9-hexachlorodibenzofuran

aa 

2,3,4,6,7,8-hexachlorodi- 1 2.8 x 10 1.7 x 10 1 2.8 x 10 1.3 x 10 1 2benzofuran

-6 -6 -6 -6 

1,2,3,4,6,7,8-heptachlorodi- 1 2.0 x 10 1.2 x 10 1 2.0 x 10 9.3 x 10 1 2benzofuran

-5 -5 -5 -6 

1,2,3,4,7,8,9-heptachlorodibenzofuran

aa 

Heptachlorodi-benzofuran 1 2.4 x 10 1.5 x 10 1 2.4 x 10 1.1 x 10 1 2

-5 -5 -5 -5 

Hexachlorodi-benzofuran 1 2.3 x 10 1.4 x 10 1 2.3 x 10 1.1 x 10 1 2

-5 -5 -5 -5 

Octachlorodi-benzofuran 1 2.1 x 10 1.3 x 10 1 2.1 x 10 9.6 x 10 1 2

-5 -5 -5 -6 

Pentachlorodi-benzofuran 1 4.0 x 10 2.4 x 10 1 4.0 x 10 1.9 x 10 1 4

-5 -5 -5 -5 

Tetrachlorodi-benzofuran 1 1.0 x 10 6.3 x 10 1 1.0 x 10 4.8 x 10 1 1

-4 -5 -4 -5 

2-methylnaphthalene 4 0.017 0.010 5 0.027 0.013 4 

Acenaphthene 2 0.358 0.220 4 0.080 0.037 2 Acenaphthylene 1 0.017 0.010 1 0.017 0.008 1 Anthracene 2 0.015 0.009 3 0.013 0.006 2 Benz(a)anthracene 3 0.030 0.018 3 0.030 0.014 3 
Table A-5. (Continued)

Oil-fired units: organic HAPs (1990) (1990) tons (1994) (1994) tons (2010) 

Number of factor: Computer Number of factor: Computer Number of 

emission lb/trillion program: emission lb/trillion program: emission lb

factors Btu 1990 total factors Btu 1994 total factors 

Median Median 

emission emission e

Benzo(b+k)fluoranthene 2 0.033 0.020 2 0.033 0.015 2 Benzo(g,h,i)perylene 2 0.021 0.013 2 0.024 0.011 2 Chrysene 3 0.021 0.013 3 0.021 0.010 3 Dibenzo(a,h)anthracene 2 0.008 0.005 2 0.010 0.005 2 Fluoranthene 6 0.016 0.010 7 0.024 0.011 6 Fluorene 5 0.021 0.013 8 0.018 0.009 5 Indeno(1,2,3-c,d)pyrene 2 0.024 0.014 2 0.024 0.011 2 Nitrobenzofluoranthene 1 0.015 0.009 1 0.015 0.007 1 Nitrochrysene/benzanthracene 1 0.016 0.010 1 0.016 0.007 1 Phenanthrene 9 0.025 0.015 11 0.036 0.017 9 Pyrene 6 0.037 0.022 8 0.010 0.004 6 
A-14

Table A-6. (REVISED) Median Emission Factors, Determined from Test Report 

1990 and Total 2010 HAP Emissions, Projected with the Emission Factor ProgrHAPs from Gas-fired Units

Gas-fired units: organic factors Btu 1990 total factors Btu 1994 total factors Btu 

HAPs (1990) (1990) tons (1994) (1994) tons (2010) (2010) 

Number of factor: Computer of factor: Computer of Median emi

emission lb/trillion program: emission lb/trillion program: emission factor: lb/tr

Median Median

emission Number emission Number

Benzene 1 1.4 1.72 1 1. 1.89 1 1.4 Formaldehyde 8 29.0 35.62 8 29. 39.23 8 29.0 

Naphthalene 2 0.67 0.83 2 0.67 0.90 2 0.67 Toluene 2 10.2 12.53 2 10.2 13.79 2 10.2 2-methylnaphthalene 2 0.026 0.031 2 0.026 0.035 2 0.026 Fluoranthene 1 0.003 0.003 1 0.0028 0.004 1 0.003 Fluorene 1 0.003 0.003 1 0.0026 0.004 1 0.003 Phenanthrene 2 0.013 0.016 2 0.013 0.017 0.013 

2

Pyrene 1 0.005 0.006 1 0.005 0.007 0.005 

1
Appendix B - Matrix of Electric Utility Steam-Generating Units and

Emission Test Sites

Table B-1 is a matrix of utility boiler types and configurations

showing each configuration’s percentage of the total fossil-fuel-fired

electric utility industry and the number of emission test sites

analyzed in this report that fit into that category’s configuration. 

The matrix was then used only as a guide to gather data on the largest

number of unit configurations possible with the available resources by

targeting the most prevalent unit types. It should be noted that the

totals in Table B-1 were taken from the 1991 EEI Power Statistics

Database and do not correlate with the 1994 industry statistics given

in Chapter 2. 

Table B-2 shows the emission test sites whose data were used to

develop this Report to Congress. Some sites are known only by their

provider number because of nondisclosure agreements. 
X
B-1

Table B-1. Utility Boilers in the United States Organized by Unit ConfigurType, and Emission Control Devices

Fuel content % (Coal only) control control type No. units % Units No. MWe 

Fuel sulfur Bottom type Particulate Sulfur dioxide (SO ) 

a b

2

c d 

Bituminous coal <1.5 Dry bottom ESP Compl fuel/SIP or NSPS 271 13.65 73385.57 Natural gas None Compl fuel/SIP or NSPS 388 19.54 67538.40 Bituminous coal 1.5 and 3.5 Dry bottom ESP Compl fuel/SIP or NSPS 274 13.80 64250.00 

Subbituminous coal <1.5 Dry bottom ESP Compl fuel/SIP or NSPS 91 4.58 40668.70 Fuel oil <1.5 None Compl fuel/SIP or NSPS 99 4.98 19948.40 Natural gas None N/A 110 5.54 18072.40 Bituminous coal 1.5 and 3.5 Dry bottom ESP FGD 39 1.96 14699.90 

Subbituminous coal <1.5 Dry bottom ESP FGD 27 1.36 13536.60 Bituminous coal 1.5 and 3.5 Wet bottom ESP Compl fuel/SIP or NSPS 45 2.27 11078.70 

Fuel oil <1.5 ESP Compl fuel/SIP or NSPS 45 2.27 10451.20 Fuel oil <1.5 Multiclone Compl fuel/SIP or NSPS 38 1.91 10403.55 Lignite <1.5 Dry bottom ESP FGD 13 0.65 8744.70 Bituminous coal <1.5 Dry bottom ESP Coal washing 24 1.21 8144.40 Bituminous coal <1.5 Dry bottom ESP FGD 18 0.91 7137.10 Bituminous coal 1.5 and 3.5 Dry bottom ESP Coal washing 11 0.55 6040.10 

Fuel oil 1.5 and 3.5 None Compl fuel/SIP or NSPS 31 1.56 5767.40 

Subbituminous coal <1.5 Wet bottom ESP Compl fuel/SIP or NSPS 18 0.91 5731.70 Subbituminous coal <1.5 Dry bottom Fabric filter FGD 10 0.50 5500.70 Bituminous coal 3.5 Dry bottom ESP FGD 11 0.55 5275.14 

Other Types of Units TestedBituminous coal <1.5 Dry bottom Fabric filter < 0.55 

Subbituminous coal <1.5 AFBC ESP < 0.55 Subbituminous coal <1.5 Dry bottom Fabric filter SDA < 0.55 
Table B-1. (continued)

Fuel content % (Coal only) control control type No. units % Units No. MWe 

Fuel sulfur Bottom type Particulate Sulfur dioxide (SO ) 

a b

2

c d 

B-2

Bituminous coal <1.5 AFBC Fabric filter < 0.55 

Bituminous /Subbituminous coal <1.5 Dry bottom Fabric filter SDA < 0.55 Fuel oil <1.5 SCR < 0.55 Fuel oil <1.5 PJFF < 0.55 Bituminous coal 1.5 and 3.5 Dry bottom PJFF SDA < 0.55 

Bituminous coal <1.5 Dry bottom ESP/COHPAC < 0.55 Natural gas Water injection < 0.55 

Combined cycle gas turbine

Lignite coal <1.5 AFBC Fabric filter < 0.55 Bituminous coal 1.5 and 3.5 Wet bottom Fabric filter WSA/SCR < 0.55 

Bottom type:dry bottom = bottom ash is disposed of in a dry state / wet bottom = bottom ash is kept molten by sweeping the furnace bottom with the fu

a

the bottom ash is disposed of by a molten liquid removal system.

AFBC - atmospheric Fluidized bed combustorESP = electrostatic Precipitator b

Multicyclone = a series of cyclones (impact particle collectors)PJFF = pulse jet fabric COHPAC = compact hybrid particulate collectorFGD: fluidized gas desulfurization (wet scrubber with some type of reagent) cSDA = spray drier absorber [SO control] 2WSA = wet sulfuric acid (system)Coal washing: coal is washed before burning to remove both ash and sulfur.Compl fuel/SIP or NSPS = fuel burned complies with applicable State Implementation Plan (SIP) or other local regulation or New Source Performance requirements.SCR = selective catalytic reduction [NO control] xNote 1: These units are essentially the same as the natural gas/compliance fuel/SIP or NSPS except these units do not have SIP or NSPS requiring emissions.Note 2: From the research performed, almost all bituminous coal is washed. These utilities choose to report coal washing as a SO control procedure 

2

other unit types.
B-3

Table B-2. Utility Boiler Emission Tests

Provider Site Fuel Fuel sulfur Bottom type Control Test contractor Re

a b 

DOE Baldwin Bituminous coal Medium Wet ESP Roy F. Weston 

DOE Boswell Subbituminous coal Low Dry FF Roy F. Weston DOE Cardinal Bituminous coal Medium Dry ESP 

Energy and Environmental Research Corp.

DOE Coal Creek Lignite coal Low Dry ESP/FGD Battelle DOE Niles Bituminous coal Medium Wet ESP Battelle DOE Niles/NO Bituminous coal Medium Wet FF/SCR/WSA Battelle 

x

DOE Springerville Subbituminous coal Low Dry SDA/FF Southern Research Institute DOE Yates Bituminous coal Medium Dry ESP/JBR Radian DOE Paradise Costain coal Medium Wet ESP Southern Research Institute NSPC A.S. King Subbituminous coal Low Wet ESP Interpoll NSPC Black Dog 1,3,4 Subbituminous coal Low Dry ESP Interpoll NSPC Black Dog 2 Subbituminous coal Low AFBC ESP Interpoll NSPC High Bridge 3,4,5,6 Subbituminous coal Low Dry ESP Interpoll NSPC Riverside 6,7 Subbituminous coal Low Dry FF Interpoll NSPC Riverside 8 Subbituminous coal Low Wet ESP Interpoll NSPC Sherburne 1,2 Subbituminous coal Low Dry FGD Interpoll 1NSPC Sherburne 3 Subbituminous coal Low Dry SDA/FF Interpoll EPRI Site 10 Bituminous coal Low AFBC FF Radian EPRI Site 101 Subbituminous coal Low Dry FF/FGD Radian EPRI Site 102 Subbituminous coal Low Wet ESP Radian EPRI Site 11 Subbituminous coal Low Dry ESP/FGD Radian EPRI/SRI Site 110 Bituminous coal Medium Dry SRI 

H-ESP/ C-ESP
Table B-2. (continued)

Provider Site Fuel Fuel sulfur Bottom type Control Test contractor Re

a b 

B-4

EPRI/SRI Site 110A/Low NO Bituminous coal Medium Dry SRI 

x H-ESP/ C-ESP

EPRI Site 111 Low Dry SDA/FF Radian 

Bituminous/ Subbituminous coal

EPRI Site 112 Oil Low Dry ESP Carnot EPRI Site 113 Oil Medium Dry Multicyclone Carnot EPRI Site 114 Bituminous coal Medium Wet ESP NA EPRI Site 115 Bituminous coal Low Dry FF Carnot EPRI Site 116 Bituminous coal Medium Dry ESP Radian EPRI Site 117 Oil Low Dry SCR Carnot EPRI Site 118 Oil Low Dry ESP Carnot EPRI Site 119 Oil Low Dry ESP Carnot EPRI Site 12 Bituminous coal Medium Dry ESP/FGD Radian EPRI Site 120 Gas Dry None NA EPRI Site 121 Gas Dry None NA EPRI Site 125 Bituminous coal Medium Wet ESP SRI EPRI Site 13 Oil Dry PJFF Radian EPRI Site 14 Bituminous coal Medium Dry SDA/PJFF Radian EPRI Site 15 Bituminous coal Medium Dry ESP Radian EPRI Site 16/OFA Bituminous coal Medium Dry ESP Radian EPRI Bituminous coal Medium Dry ESP Radian 

Site 16/OFA/Low NOx

EPRI Site 18 Bituminous coal Low Dry Radian 

ESP/ COHPAC

EPRI Site 19 Bituminous coal Low Dry ESP Radian EPRI Site 20 Lignite coal Medium Dry ESP/FGD Radian 
Table B-2. (continued)

Provider Site Fuel Fuel sulfur Bottom type Control Test contractor Re

a b 

B-5

EPRI Site 21 Bituminous coal Medium Dry ESP/FGD Radian 

EPRI Site 22 Bituminous coal Low Dry ESP Radian/Carnot EPRI Site 103 Oil Low Dry None Carnot EPRI Site 104 Oil Low Dry None Carnot EPRI Site 105 Oil Low Dry None Carnot EPRI Site 106 Oil Low Dry None Carnot EPRI Site 107 Oil Low Dry None Carnot EPRI Site 108 Oil Medium Dry None Carnot EPRI Site 109 Oil Low Dry None KVB - Analect EPA (Combined Cycle Gas Water injection Entropy 

T. H. WhartonGas Turbine)

EPA Northport Oil Low Dry ESP Entropy EPA Lignite coal Low AFBC FF Entropy 

Hammond (TNP One)

EPA Greens Bayou Gas Dry None Entropy EPA Kintigh Bituminous coal Medium Dry ESP/FGD Entropy 

DOE = Department of Energy aNSPC = Northern States Power CompanyEPRI = Electric Power Research InstituteSRI = Southern Research InstituteESP = Electrostatic precipitator bFF = Fabric filterFGD = Flue gas desulfurizationSCR = Selective catalytic reduction [NO control] xWSA = Wet sulfuric acid (system)SDA = Spray drier absorber [SO control] 2JBR = Jet bubbling reactor [SO control] 2H-ESP = Hot-side electrostatic precipitatorC-ESP = Cold-side electrostatic precipitatorPJFF = Pulse jet fabric filterCOHPAC = Compact hybrid particulate collector(P) - Preliminary Test Report c(F) - Final Test Report
B-6

1. Preliminary draft emissions report for Baldwin Power Station -

Unit 2 (Illinois Power Company) for the Comprehensive Assessment

of Toxic Emissions from Coal-fired Power Plants, prepared by Roy

F. Weston, Inc., for the Department of Energy/Pittsburgh Energy

Technology Center (DOE/PETC), DOE contract # DE-AC22-93PC93255,

Weston project # 10016-011, Weston report # DOE018G.RP1. 

December 1993.

2. Preliminary draft emissions report for Boswell Energy Center -

Unit 2 (Minnesota Power Company) for the Comprehensive Assessment

of Toxic Emissions from Coal-fired Power Plants, prepared by Roy

F. Weston, Inc., for the Department of Energy/Pittsburgh Energy

Technology Center (DOE/PETC), DOE contract # DE-AC22-93PC93255,

Weston project # 10016-011, Weston report # DOE017G.RP1. 

December 1993.

3. Preliminary draft emissions report for Cardinal Station - Unit 1

(American Electric Power) for the Comprehensive Assessment of

Toxic Emissions from Coal-fired Power Plants, prepared by Energy

and Environmental Research Corp. for the Department of

Energy/Pittsburgh Energy Technology Center (DOE/PETC), DOE

contract # DE-AC22-93PC93252. December 1993.

4. Preliminary draft emissions report for Coal Creek Station - Unit

2 (Cooperative Power Association) for the Comprehensive

Assessment of Toxic Emissions from Coal-fired Power Plants,

prepared by Battelle for the Department of Energy/Pittsburgh

Energy Technology Center (DOE/PETC), DOE contract # DE-AC22-

93PC93251. December 1993.

5. Preliminary draft emissions report for Niles Station Boiler No. 2

(Ohio Edison) for the Comprehensive Assessment of Toxic Emissions

from Coal-fired Power Plants, prepared by Battelle for the

Department of Energy/Pittsburgh Energy Technology Center

(DOE/PETC), DOE contract # DE-AC22-93PC93251. December 1993.

6. Preliminary draft emissions report for Niles Station Boiler No. 2

with NO control (Ohio Edison) for the Comprehensive Assessment x

of Toxic Emissions from Coal-fired Power Plants, prepared by

Battelle for the Department of Energy/Pittsburgh Energy

Technology Center (DOE/PETC), DOE contract # DE-AC22-93PC93251. 

December 1993.

7. Preliminary draft emissions report for Springerville Generating

Station Unit No. 2 (Tucson Electric Power Company) for the

Comprehensive Assessment of Toxic Emissions from Coal-fired Power

Plants, prepared by Southern Research Institute for the

Department of Energy/Pittsburgh Energy Technology Center

(DOE/PETC), DOE contract # DE-AC22-93PC93254, SRI report No. SRIENV-93-1049-7960. December 1993.

Appendix B: References
B-7

8. Preliminary draft emissions report for Plant Yates Unit No. 1

(Georgia Power Company) for the Comprehensive Assessment of Toxic

Emissions from Coal-fired Power Plants, prepared by Electric

Power Research Institute for the Department of Energy/Pittsburgh

Energy Technology Center (DOE/PETC), EPRI report No. DCN 93-643-

004-03. December 1993.

9. Draft final report for Paradise Fossil Plant for the

Comprehensive Assessment of Air Toxic Emissions, prepared by

Southern Research Institute for the Department of

Energy/Pittsburgh Energy Technology Center (DOE/PETC), SRI report

No. SRI-ENV-95-338-7960. May 1995.

10. Results of the Air Toxic Emission Study on the No. 1 Boiler at

the NSPC A.S. King Plant, prepared by Interpoll Laboratories,

Inc., for NSPC, report No. 1-3304. November 1991. 

11. Results of the Air Toxic Emission Study on the Nos. 1, 3, & 4

Boilers at the NSC Black Dog Plant, prepared by Interpoll

Laboratories, Inc., for NSPC, report No. 1-3451. January 1992.

12. Results of the Air Toxic Emission Study on the No. 2 Boiler at

the NSPC Black Dog Plant, prepared by Interpoll Laboratories,

Inc., for NSPC, report No. 2-3496. May 1992.

13. Results of the Air Toxic Emission Study on the Nos. 3, 4, 5, & 6

Boilers at the NSPC High Bridge Plant, prepared by Interpoll

Laboratoried, Inc., for NSPC, report No. 1-3453. January 1992.

14. Results of the Air Toxic Emission Study on the Nos. 6 & 7 Boilers

at the NSPC Riverside Plant, prepared by Interpoll Laboratorieds,

Inc., for NSPC, report No. 1-3468A. February 1992.

15. Results of the Air Toxic Emission Study on the No. 8 Boiler at

the NSPC Riverside Plant, prepared by Interpoll Laboratories,

Inc., for NSPC, report No. 2-3590. September 1992. 

16. Results of the Air Toxic Emission Study on the No. 1 & 2 Boilers

at the NSPC Sherburn Plant, prepared by Interpoll Laboratories,

Inc., for NSPC, report No. 0-3053. July 1990/ October 1991. 

17. Results of the Mercury Removal Tests on Units 1 & 2, and the Unit

3 Scrubber System at the NSPC Sherburne Plant, prepared by

Interpoll Laboratories, Inc., for NSPC, report No. 1-3409. 

October 1991. 

18. Results of the May 1, 1990, Trace Metal Characterization Study on

Units 1 & 2 at the NSPC Sherburne Plant, prepared by Interpoll

Laboratories, Inc., for NSPC, report No. 0-3033E. July 1990. 
B-8

19. Results of the Air Toxic Emission Study on the No. 3 Boiler at

the NSPC Sherburne Plant, prepared by Interpoll Laboratories,

Inc., for NSPC, report No. 0-3005. June 1990/October 1991. 

20. Preliminary draft emissions report for EPRI Site 10, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 92-213-152-35. 

October 1992.

21. Preliminary draft emissions report for EPRI Site 101, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 94-643-015-02. 

October 1994.

22. Preliminary draft emission report for EPRI Site 102, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 92-213-152-35. 

February 1993. 

23. Preliminary draft emissions report (and mercury retest) for EPRI

Site 11, Field Chemical Emissions Monitoring Project, prepared by

Radian Corporation for EPRI. EPRI report Nos. DCN 92-213-152-24

and DCN 92-213-152-48. November 1992/October 1993.

24. Preliminary draft emissions report for EPRI Site 110 (baseline

and with NOx control) for the EPRI PISCES Study, prepared by

Southern Research Institute, SRI report No. SRI-ENV-92-796-7496. 

October 1993.

25. Preliminary draft emissions report for EPRI Site 111, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRi. EPRI report No. DCN 93-213-152-42. 

January 1994. 

26. Preliminary draft emissions report for EPRI Site 112, Field

Chemical Emissions Monitoring Project, prepared by Carnot for

EPRI. Carnot report No. EPRIE-10106/R016C374.T. March 1994.

27. Preliminary draft emission report for EPRI Site 113, Field

Chemical Emission Monitoring Project, prepared by Carnot for

EPRI. EPRI report No. EPRIE-10106/R140C808.T. March 1994.

28. Preliminary draft emissions report for EPRI Site 114, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 92-213-152-51. May

1994.

29. Preliminary draft emissions report for EPRI Site 115, Field

Chemical Emissions Monitoring Project, prepared by Carnot for

EPRI. Carnot report No. EPRIE-10106-R022C855.T. 
B-9

30. Preliminary draft emissions report for EPRI Site 116, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 94-213-152-55. 

October 1994.

31. Preliminary draft emissions report for EPRI Site 117, Field

Chemical Emissions Monitoring Project, prepared by Carnot for

EPRI. Carnot report No. EPRIE-10106-R120C844.T. January 1994.

32. Preliminary draft emissions report for EPRI Site 118, Field

Chemical Emissions Monitoring Project, prepared by Carnot for

EPRI. Carnot report No. EPRIE-10106/R140C928.T. January 1994. 

33. Preliminary draft emissions report for EPRI Site 119, Field

Chemical Emissions Monitoring Project, prepared by Carnot for

EPRI. Carnot report No. EPRIE-10106/R027C882.T. January 1994. 

34. Preliminary draft emissions report (and mercury retest) for EPRI

Site 12, Field Chemical Emissions Monitoring Project, prepared by

Radian Corporation for EPRI. EPRI report Nos. DCN 92-213-152-27

and DCN 93-213-152-49. November 1992/ October 1993.

35. Preliminary draft emissions supplement for EPRI Site 120, Field

Chemical Emission Monitoring Project.

36. Preliminary draft emissions report for EPRI Site 121, Field

Chemical Emissions Monitoring Project, prepared by Carnot for

EPRI. Carnot report No. EPRIE-12102/R120E916.T. December 1994.

37. Preliminary draft emissions report for EPRI Site 125, Field

Chemical Emissions Monitoring Project, prepared by Southern

Research Institute for EPRI. EPRI report No. RP9028-10. August

1995.

38. Preliminary draft emissions report for EPRI Site 13, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 93-213-152-36. 

February 1993. 

39. Preliminary draft emissions report for EPRI Site 14, Field

Chemical Emissions Monitoring Project, prepared by Radian

Coporation for EPRI. EPRI report No. DCN 93-213-152-28. 

November 1992.

40. Preliminary draft emissions report for EPRI Site 15, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 93-213-152-26. 

October 1992. 
B-10

41. Preliminary draft emissions report for EPRI Site 16 (OFA and

OFA/Low NOx) for the Clean Coal Technology Project (CCT),

prepared by Electric Power Research Institute, for the Department

of Energy/Pittsburgh Energy Technology Center (DOE/PETC), EPRI

report No. DCN 93-209-061-01. November 1993. 

42. Preliminary draft emission report for EPRI Site 18, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 93-213-152-43. April

1993.

43. Preliminary draft emissions report for EPRI Site 19, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 93-213-152-41. April

1993. 

44. Preliminary draft emissions report for EPRI Site 20, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 93-213-152-54. March

1994. 

45. Preliminary draft emissions report for EPRI Site 21, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation for EPRI. EPRI report No. DCN 93-213-152-39. May

1993. 

46. Preliminary draft emissions report for EPRI Site 22, Field

Chemical Emissions Monitoring Project, prepared by Radian

Corporation and Carnot for EPRI. EPRI report No. DCN 93-213-152-

53. February 1994.

47. Preliminary draft emissions report for EPRI Sites 103-109, Field

Chemical Emissions Monitoring Project: Emissions Report for Sites

103-109, prepared by Radian Corporation for EPRI. March 1993. 

48. Final electric utility combined cycle gas-fired gas turbine

emission test report for T.H. Wharton Electric Generating Station

(Houston Lighting and Power Company), prepared by Entropy, Inc.,

for the U.S. Environmental Protection Agency, Emissions

Measurement Branch (EPA/EMB), EMB report No. 93-UTL-2. May 1994. 

49. Final electric utility fuel oil-fired electric utility boiler

emission test report for Northport One powerplant - Unit 1 (Long

Island Lighting Corporation), prepard by Entropy, Inc., for the

U.S. Environmental Protection Agency, Emissions Measurement

Branch (EPA/EMB), EMB report No. 93-UTL-4. April 1994.
B-11

50. Final electric utility coal-fired fluidized bed boiler emission

test report for TNP One - Unit 2 (Texas - New Mexico Power

Company), prepared by Entropy, Inc., for the U.S. Environmental

Protection Agency, Emissions Measurement Branch (EPA/EMB), EMB

report No. 93-UTL-1. June 1994.

51. Final electric utility gas-fired boiler emission test report for

Greens Bayou Electric Generating Station - Unit 5 (Houston

Lighting and Power Company), prepared by Entropy, Inc., for the

U.S. Environmental Protection Agency, Emissions Measurement

Branch (EPA/EMB), EMB report No. 93-UTL-3. May 1994.

52. Final electric utility coal-fired boiler emission test report for

Kintigh - Unit 1 (New York State Electric and Gas Company),

prepared by Entropy, Inc., for the U.S. Environmental Protection

Agency, Emissions Measurement Branch (EPA/EMB), EMB report No.

93-UTL-5. June 1994.
mm
Appendix C - Listing of Emission Modification Factors for Trace

Elements Used in the Individual Boiler Analysis

Note: The following test reports were not used to develop

emission modification factors (EMFs) for the reasons listed below. 

Northern States Power’s (NSP) A.S. King unit is the same test site as

the Electric Power Research Institute’s (EPRI’s) Site 102, and the EPA

chose to use the EPRI test report. Northern States Power’s Sherco

unit 1 and 2 were not used to develop boiler EMFs because no coal

composition data were provided. Northern States Power’s Black Dog

unit 1 was not used to develop boiler EMFs because tangentially-fired

emissions were combined with emissions from two front-fired boilers. 

Finally, NSPC’s High Bridge was not used to develop boiler EMFs

because the test report was missing the coal feed rate during testing. 
X
C-1

Table C-1. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Circulating Fluidized Bed Furnaces

(Coal-Fired)

Unit Name EPRI Site 10 NSP - Black Dog #2 mean) deviation

EMF (Geometric Geometric standard

Arsenic 1.00 0.59 0.77 1.44

Beryllium 0.77 0.41 0.56 1.56

Chromium 0.40 0.54 0.46 1.25

Cobalt 1.00 1.00 N/A

Lead 0.49 0.36 0.42 1.24

Manganese 0.59 0.68 0.63 1.11

Mercury 1.00 1.00 1.00 1.00

Nickel 1.00 0.45 0.67 1.76

Selenium 1.00 0.71 0.84 1.27

Table C-2. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Tangentially-fired, Dry-bottom Furnace with

NO Control (Coal-Fired) x

Unit Name EPRI Site 11 w/NO control Creek Springerville mean) deviation

EPRI Site 110 DOE - Coal DOE - (Geometric standard

x

EMF Geometric

Antimony 0.66 0.01 0.03 0.06 4.74

Arsenic 0.92 0.39 0.69 0.29 0.52 1.11

Beryllium 0.79 0.43 0.35 0.87 0.57 1.08

Cadmium 0.35 0.70 0.11 1.00 0.41 1.42

Chromium 0.72 1.00 1.00 0.68 0.84 1.02

Cobalt 0.92 1.00 0.61 0.73 0.80 1.02

Lead 1.00 0.36 0.29 0.19 0.37 1.21

Manganese 0.98 0.76 0.59 0.72 0.75 1.02

Mercury 0.85 1.00 0.92 1.00

Nickel 0.25 0.97 1.00 0.70 0.64 1.18

Selenium 1.00 0.82 0.38 0.93 0.74 1.08
C-2

Table C-3. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Tangentially-fired, Dry-bottom Furnace Without

NO Control (Coal-fired) x

Unit Name EPRI Site 15 EPRI Site 110 DOE - Yates mean) deviation

EMF (Geometric standard

Geometric

Antimony 0.55 0.67 0.61 1.01

Arsenic 0.60 0.89 1.00 0.81 1.02

Beryllium 0.54 0.93 1.00 0.80 1.04

Cadmium 0.01 1.00 1.00 0.23 8.72

Chromium 0.58 1.00 1.00 0.84 1.03

Cobalt 0.94 0.93 0.97 0.95 1.00

Lead 1.00 1.00 1.00 1.00 1.00

Manganese 0.81 0.71 1.00 0.83 1.01

Mercury 0.66 1.00 0.81 1.02

Nickel 0.43 0.91 0.84 0.69 1.06

Selenium 0.70 0.58 0.70 0.66 1.00

Table C-4. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Opposed-fired, Dry-bottom Furnace with NOx

Control (Coal-fired)

Unit Name Site 12 Site 14 #3 Site 111 LNO Burners 16 w/OFA mean) deviation

EPRI EPRI Sherburne EPRI w/OFA and EPRI Site (Geometric standard

NSP - EPRI Site 16 EMF Geometric

X

Antimony 0.80 1.00 0.90 1.17

Arsenic 1.00 0.50 0.79 0.11 1.00 1.00 0.59 2.41

Beryllium 1.00 0.92 0.58 0.82 1.00 0.85 1.25

Cadmium 0.14 0.02 0.99 0.05 0.11 1.00 0.16 4.68

Chromium 1.00 0.67 0.49 0.20 0.69 0.58 0.55 1.72

Cobalt 1.00 1.00 0.66 0.61 0.80 1.31

Lead 1.00 0.79 0.49 0.66 1.00 0.76 1.35

Manganese 1.00 0.93 0.88 0.60 0.84 1.26

Mercury (Note 1) 0.74 1.00 1.00 0.64 0.81 1.22

0.74

Nickel 0.29 0.37 0.67 0.02 0.54 0.33 0.24 2.25

Selenium 1.00 0.05 0.21 0.37 1.00 0.33 3.51

Note 1 - This EMF was obtained from the mercury retest.
C-3

Table C-5. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Front-fired, Dry-bottom Furnace Without NOx

Control (Coal-fired)

Unit Name Riverside DOE - (Geometric standard EPRI Site Geometric Standard

NSP - 1990 EMF Geometric 1994 Geometric

#6-7 Boswell mean) deviation 116 Mean Deviation

1990 1994

Antimony 0.20 0.59 0.34 1.16 0.12 0.24 1.93

Arsenic 0.99 0.23 0.48 1.30 0.7 0.54 1.76

Beryllium 0.40 0.60 0.49 1.02 0.35 0.44 1.08

Cadmium 0.25 1.00 0.50 1.27 0.12 0.31 3.18

Chromium 1.00 1.00 1.00 1.00 0.27 0.65 1.77

Cobalt 0.98 0.98 1.00 0.3 0.54 1.42

Lead 0.19 0.42 0.28 1.08 0.26 0.28 1.16

Manganese 0.77 0.57 0.66 1.01 0.21 0.45 1.58

Mercury 1.00 0.87 0.93 1.00 0.97 0.94 1.01

Nickel 0.78 1.00 0.88 1.00 0.23 0.56 1.86

Selenium 1.00 0.14 0.37 1.63 0.47 0.40 2.71

Table C-6. Tested EMFs and Geometric Means used in the Emission

Factor Program for Cyclone-fired, Wet-bottom Furnace Without NOx

Control (Coal-fired)

Unit Name 102 #8 114 Niles #2 Control Baldwin mean) deviation

EPRI Site Riverside EPRI Site DOE - w/NOx DOE- (Geometric standard

NSP - Niles #2 EMF Geometric

DOE -

Antimony 0.61 1.00 0.92 0.32 0.65 1.02

Arsenic 0.48 0.51 0.15 0.58 0.85 1.00 0.51 1.20

Beryllium 0.04 0.08 0.15 0.26 0.30 0.44 0.16 1.38

Cadmium 0.02 0.16 0.01 0.11 0.20 1.00 0.10 3.79

Chromium 0.25 0.22 0.30 0.28 0.35 0.42 0.29 1.02

Cobalt 0.21 0.20 0.33 0.34 0.26 1.03

Lead 0.61 0.38 0.50 0.56 0.60 0.53 0.52 1.01

Manganese 0.33 0.13 0.20 0.15 0.19 0.20 0.19 1.04

Mercury 1.00 1.00 0.73 1.00 1.00 0.92 0.93 1.01

Nickel 0.30 0.12 0.72 0.29 0.39 0.43 0.33 1.15

Selenium 0.65 1.00 0.26 1.00 0.92 0.04 0.43 1.96
C-4

Table C-7. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Vertically-fired, Dry-bottom Furnace With NOx

Control (Coal-fired)

Unit Name EPRI Site 115 EMF (Geometric mean) deviation

Geometric standard

Arsenic 0.61 0.61 N/A

Beryllium 0.52 0.52 N/A

Cadmium 0.58 0.58 N/A

Chromium 0.57 0.57 N/A

Cobalt 1.00 1.00 N/A

Lead 0.38 0.38 N/A

Manganese 0.58 0.58 N/A

Mercury 0.78 0.78 N/A

Nickel 0.64 0.64 N/A

Selenium 0.34 0.34 N/A

Table C-8. Tested EMFs and Geometric Means used in the Emission

Factor Program for Cyclone-fired, Wet-bottom Furnace With NOx

Control (Coal-fired)

Unit Name EPRI Site 114, NO EMF (Geometric mean) deviation x

Geometric standard

Antimony N/A

Arsenic 0.25 0.25 N/A

Beryllium 0.15 0.15 N/A

Cadmium 0.01 0.01 N/A

Chromium 0.23 0.23 N/A

Cobalt N/A

Lead 0.84 0.84 N/A

Manganese 0.18 0.18 N/A

Mercury 0.54 0.54 N/A

Nickel 0.31 0.31 N/A

Selenium 1.00 1.00 N/A
C-5

Table C-9. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Opposed-fired, Dry-bottom Furnace Without Nox

Control (Coal-fired)

Unit Name DOE - Cardinal EMF (Geometric mean) deviation

Geometric standard

Antimony 0.08 0.08 N/A

Arsenic 0.91 0.91 N/A

Beryllium 0.96 0.96 N/A

Cadmium 1.00 1.00 N/A

Chromium 0.61 0.61 N/A

Cobalt 0.96 0.96 N/A

Lead 1.00 1.00 N/A

Manganese 0.27 0.27 N/A

Mercury 0.41 0.41 N/A

Nickel 0.76 0.76 N/A

Selenium 0.07 0.07 N/A
C-6

Table C-10. Tested EMFs and Geometric Means Used in the Emission Factor PrFront-fired, Dry-bottom Furnace Without No Control (Oil-fired) x

Unit Name EPRI Site EPRI Site EPRI Site EPRI Site EPRI Site EPRI Site EPRI Site EMF (Geometric 

13, no NOx 103 104 105 107 108 117 Mean) 

Arsenic 1.00 0.06 0.10 0.07 0.23 0.12 0.40 0.17 Beryllium 0.02 0.02 1.00 0.07 Cadmium 0.50 0.19 0.06 0.06 0.14 0.36 0.03 0.13 Chromium 0.25 0.29 0.22 0.02 0.66 0.05 0.56 0.18 Cobalt 0.82 0.32 0.42 0.48 Lead 0.03 0.06 0.04 0.14 0.16 0.32 0.09 Manganese 1.00 1.00 0.57 0.16 0.45 1.00 0.21 0.51 Mercury 1.00 1.00 1.00 

Nickel 0.89 0.69 0.28 0.40 0.74 0.90 0.76 0.62 Selenium 0.005 0.04 0.07 0.23 0.37 0.06 
C-7

Table C-11. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Opposed-fired, Dry-bottom Furnace Without Nox

Control (Oil-fired)

Unit Name EPRI Site 106 EPRI Site 109 mean) deviation

EMF (Geometric Geometric standard

Arsenic 0.45 0.01 0.08 11.80

Beryllium 0.02 0.02 N/A

Cadmium 0.10 0.39 0.20 2.61

Chromium 0.32 1.00 0.57 2.23

Cobalt 1.00 1.00 1.00 1.00

Lead 0.46 0.26 0.35 1.50

Manganese 1.00 0.80 0.89 1.17

Mercury 0.04 0.04 N/A

Nickel 1.00 0.79 0.89 1.18

Selenium 0.10 0.02 0.04 3.41

Note: This EMF suggests that this boiler type does a good job of controlling/reducing mercury from the fuel oil being burned. 

Nothing in the test report suggests that there were any problems encountered with this test sample so EPA chose to leave it in the

EFP. It should be noted that all other tested oil-fired boilers had mercury EMFs of 1.0.

Table C-12. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Front-fired, Dry-bottom Furnace with Nox

Control (Oil-fired)

Unit Name Site 13 Site 118 Site 117 mean) deviation Site 113 mean) deviation

EPRI EPRI EPRI (Geometric standard EPRI (Geometric standard

1990 EMF Geometric 1994 EMF Geometric

1990 1994

Arsenic 0.64 0.14 1.00 0.44 2.83 0.21 0.37 1.95

Beryllium 1.00 1.00 1.00 1.00 1.00 1.00

Cadmium 0.47 1.00 0.69 1.71 1.00 0.78 1.27

Chromium 0.19 0.78 1.00 0.53 2.42 1.00 0.62 1.68

Cobalt 0.62 0.29 0.98 0.56 1.83 0.40 0.52 1.58

Lead 0.08 0.57 0.97 0.35 3.73 0.37 0.36 2.18

Manganese 1.00 1.00 1.00 1.00 1.00 0.93 0.98 1.02

Mercury 1.00 1.00 1.00 1.00 1.00 1.00 1.00 1.00

Nickel 0.71 0.64 1.00 0.77 1.26 0.40 0.65 1.32

Selenium 0.58 0.46 1.00 0.64 1.50 1.00 0.72 1.30
C-8

Table C-13. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Tangentially-fired, Dry-bottom Furnace Without

No Control (Oil-fired) x

Unit Name EPRI Site 112 EMF (Geometric mean) Geometric standard deviation

Arsenic 1.00 1.00 N/A

Beryllium 0.79 0.79 N/A

Cadmium 0.67 0.67 N/A

Chromium 0.66 0.66 N/A

Cobalt 0.38 0.38 N/A

Lead 0.26 0.26 N/A

Manganese 0.80 0.80 N/A

Mercury 1.00 1.00 N/A

Nickel 0.53 0.53 N/A

Selenium 1.00 1.00 N/A

Table C-14. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Tangentially-fired, Dry-bottom Furnace with

No Control (Oil-fired) x

Unit Name EPRI Site 119 See Note 1 EMF (Geometric mean) Geometric standard deviation

Arsenic 0.44 0.44 N/A

Beryllium 1.00 1.00 N/A

Cadmium 0.69 0.69 N/A

Chromium 1.00 0.62 0.79 1.40

Cobalt 0.56 0.56 N/A

Lead 0.35 0.35 N/A

Manganese 1.00 1.00 N/A

Mercury 1.00 1.00 N/A

Nickel 0.57 0.72 0.64 1.17

Selenium 0.64 0.64 N/A

Note 1. Since the only source of data for this type of unit was limited to data on only two metals, it was decided to take the data

from another similar unit (a front-fired, dry-bottom furnace with NOx control) along with the 2-data-point set to develop a set of

geometric means. This set of geometric means is the data set in the "See Note 1" column. The geometric means of the "See Note

1" set and the 2-data-point set were derived. These means were used to represent a tangential-fired, dry-bottom furnace with NO x

control burning oil.
C-9

Table C-15. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Fabric Filters (baghouses)

Unit Name 10 #6-7 115 Control Boswell mean) deviation

EPRI Site Riverside EPRI Site #2 w/NO DOE - (Geometric standard

NSP - DOE - Niles EMF Geometric

x

Antimony 0.03 0.005 0.06 0.02 1.66

Arsenic 0.001 0.03 0.03 0.004 0.009 0.01 2.20

Beryllium 0.004 0.06 0.003 0.001 0.01 0.01 2.51

Cadmium 1.00 0.05 0.01 0.06 0.08 3.47

Chromium 0.05 0.013 0.003 0.004 0.01 1.84

Cobalt 0.004 0.007 0.001 0.006 0.004 1.36

Lead 0.002 0.03 0.007 0.001 0.01 0.01 2.70

Manganese 0.018 0.05 0.005 0.005 0.003 0.01 1.66

Mercury 1.00 0.27 0.92 0.39 0.56 1.17

Nickel 0.002 0.20 0.05 0.001 0.007 0.01 11.26

Selenium 0.06 0.02 0.79 0.31 0.12 3.09

Hydrogen Chloride Note 1 0.56

Hydrogen Fluoride Note 1 1.00

Note 1 - These EMFs were developed from emission tests that examined HCl and HF emissions through a baghouse.

Table C-16. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Electrostatic Precipitators - Hot Side

(Located Before the Air Preheater, Controlling an Coal-fired Unit)

Unit Name EPRI Site 110 Control mean) deviation

EPRI Site 110 w/NO EMF (Geometric Geometric standard x

Antimony 0.11 0.02 3.87 0.04

Arsenic 0.01 0.15 7.13 0.04

Beryllium 0.01 0.01 1.08 0.01

Cadmium 0.004 0.01 2.39 0.01

Chromium 0.02 0.04 1.84 0.03

Cobalt 0.04 0.02 1.55 0.03

Lead 0.02 0.03 1.59 0.02

Manganese 0.04 0.02 1.86 0.02

Mercury 1.00 1.00 1.00 1.00

Nickel 0.002 0.01 3.24 0.004

Selenium 1.00 0.87 1.10 0.93

Hydrogen Chloride Note 2 1.00

Hydrogen Fluoride Note 2 1.00

Note 2 - Because there were no data on HCl and HF emissions through an ESP attached to an oil-fired unit or a hot-side ESP

attached to a coal-fired unit, the EMF was left as "1" so that all HCl and HF emissions passed through the ESP.
C-10

Table C-17. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Electrostatic Precipitators - Cold Side 

(Located after the Air Preheater, Controlling an Oil-fired Unit)

Unit Name EPRI Site 112 EPRI Site 118 mean) deviation

EMF (Geometric Geometric standard

Arsenic 0.49 0.55 1.09 0.52

Beryllium 0.23 0.10 1.76 0.16

Cadmium 0.69 N/A 0.69

Chromium 0.44 0.44 1.00 0.44

Cobalt 0.25 0.08 2.27 0.14

Lead 0.47 0.43 1.07 0.45

Manganese 1.00 0.83 1.14 0.91

Mercury 0.17 0.58 2.39 0.31

Nickel 0.27 0.07 2.50 0.14

Selenium 0.648 N/A 0.65

Hydrogen Chloride Note 2 1.00

Hydrogen Fluoride Note 2 1.00

Note 2 - Because there were no data on HCl and HF emissions through an ESP attached to an oil-fired unit or a hot-side ESP

attached to a coal-fired unit, the EMF was left as "1" so that all HCl and HF emissions passed through the ESP.

Table C-18. Tested EMFs and Geometric Means Used in the Emission

Factor Program for Particulate Matter Scrubber Unit -

(controlling a coal-fired unit)

Unit Name EPRI Site 125 EMF (Geometric mean) Geometric standard deviation

Antimony 0.10 0.10 N/A

Arsenic 0.10 0.10 N/A

Beryllium 0.02 0.02 N/A

Cadmium 0.09 0.09 N/A

Chromium 0.03 0.03 N/A

Cobalt 0.02 0.02 N/A

Lead 0.03 0.03 N/A

Manganese 0.01 0.01 N/A

Mercury 0.96 0.96 N/A

Nickel 0.01 0.01 N/A

Selenium 1.00 1.00 N/A

Hydrogen Chloride 0.06 0.06 N/A

Hydrogen Fluoride 0.09 0.09 N/A
C-11

Table C-19. Tested EMFs and Geometric Means Used in the Emission Factor Pr

Gas Desulfurization Unit - (Controlling a Coal-fired Unit)

Unit Name Site 11 Site 12 #1-2 2nd test) Yates Creek mean) deviation Site 20 101 Paradis

EPRI EPRI Sherburne #1-2 (mercury DOE - Coal (Geometric standard EPRI Site DOE - 

NSP - Sherburne DOE - 1990 EMF 1990 Geometric EPRI 

NSP - 

Antimony 0.21 0.16 0.18 1.23 0.28 Arsenic 0.49 0.23 0.04 0.08 0.76 0.20 3.29 0.19 0.57 0.14 Beryllium 0.02 0.11 0.97 0.13 6.96 0.15 0.33 0.03 Cadmium 0.85 0.62 0.16 0.75 0.50 2.16 1.00 0.83 0.10 Chromium 0.51 0.40 0.03 0.18 1.00 0.26 3.72 0.47 0.90 0.07 Cobalt 1.00 0.23 0.23 0.91 0.46 2.30 0.24 0.41 0.10 Lead 1.00 0.59 0.02 0.04 0.93 0.22 6.36 0.45 0.33 0.09 Manganese 1.00 0.03 0.17 1.00 0.26 5.65 0.32 1.00 0.01 Mercury 0.89 1.00 0.77 0.41 0.54 0.88 0.71 1.42 0.80 0.38 0.55 

(Note 1) (Note 2)

Nickel 0.88 1.00 0.05 1.00 1.00 0.53 3.79 0.78 1.00 0.01 Selenium 0.17 0.27 0.17 0.26 0.60 0.26 1.66 0.21 0.47 0.71 Hydrogen Note 3 0.21 Note 3 0.20ChlorideHydrogen Note 3 0.73 Note 3 0.71Fluoride

Note 1 - This EMF was obtained from the mercury retest.Note 2 - This EMF was obtained from the mercury retest. Because emissions testing was done at the inlet to the ESP and the outlet of the FGD only, EPmercury removal potential to the ESP (see Table C-21).Note 3 - See section 3.4.7. for explanation of HCl and HF EMF Development.
C-12

Table C-20. Tested EMFs and Geometric Means Used in the Emission Factor Pr

Dryer Adsorber / Fabric Filter Units - (Controlling a Coal-fired Unit)

Unit Name #3 test) EPRI Site 111 EPRI Site 14 Springerville me

NSP - Sherburne #3 (mercury 2nd DOE - EMF (Ge

NSP - Sherburne 

Antimony 0.01 0.001 0.0Arsenic 0.001 0.04 0.002 0.0003 0.0Beryllium 0.0003 0.01 0.0001 0.0Cadmium 0.08 0.13 0.19 0.001 0.0Chromium 0.03 0.03 0.005 0.0Cobalt 0.01 0.0004 0.0Lead 0.01 0.002 0.001 0.0Manganese 0.003 0.003 0.001 0.0Mercury 0.54 0.46 1.00 0.98 0.7Nickel 0.04 0.04 0.01 0.0003 0.0Selenium 0.05 0.13 0.0004 0.0Hydrogen Chloride Note 1 0.1Hydrogen Fluoride Note 1 0.1

Note 1 - See section 3.4.7. for explanation of HCl and HF EMF Development.
C-13

Table C-21. Tested Emission Modification Factors (EMFs) and Geometric MeanEmission Factor Program for Electrostatic Precipitators - Cold Side (LocatePreheater, Controlling Coal-fired Unit)

Unit Name Site 11 Site 12 Site 15 Site 102 3, 4, 5, and 6 3, and 4 #2 #8 114 

EPRI EPRI EPRI EPRI NSP - High Bridge NSP - Black Dog 1, NSP - Black Dog NSP - Riverside EPRI Site 

Antimony 0.25 0.10 0.32 

Arsenic 0.01 0.004 0.02 0.06 0.01 0.01 0.01 0.01 0.05 Beryllium 0.10 0.002 0.01 0.001 0.01 0.02 0.003 0.02 0.06 

Cadmium 0.03 0.06 0.41 0.11 0.11 0.04 0.99 0.13 0.10 Chromium 0.03 0.01 0.01 0.06 0.07 0.01 0.04 0.07 0.05 Cobalt 0.01 0.01 0.01 0.03 Lead 0.04 0.03 0.01 0.02 0.02 0.05 0.66 0.04 0.06 Manganese 0.01 0.04 0.005 0.03 0.03 0.02 0.02 0.05 0.05 

Mercury 1.00 1.00 0.93 0.92 0.78 1.00 0.70 

0.18 (Note 1)

Nickel 0.01 0.003 0.01 0.37 0.03 0.04 0.15 0.03 0.05 Selenium 0.04 0.24 0.65 0.004 0.02 0.03 0.02 0.01 1.00 Hydrogen Chloride

Hydrogen Fluoride

Unit Name Yates Creek and LNOX Burners 16 w/OFA Cardinal Baldwin (Geometric Mean) Standard Deviation 116 

DOE - DOE - Coal EPRI Site 16 w/OFA EPRI Site DOE - DOE - 1990 EMF 1990 Geometric EPRI Site (G

1

Antimony 0.01 0.02 0.03 0.04 0.02 0.10 0.04 4.21 0.12 

Arsenic 0.04 0.001 0.06 0.05 0.003 0.01 0.02 3.07 0.01 Beryllium 0.02 0.05 0.02 0.03 0.001 0.03 0.01 4.13 0.73 

Cadmium 0.04 0.18 0.17 0.05 0.03 0.05 0.08 2.93 1 Chromium 0.01 0.005 0.02 0.04 0.01 0.05 0.02 2.50 0.003 Cobalt 0.02 0.01 0.02 0.03 0.002 0.07 0.01 3.52 0.29 Lead 0.02 0.002 0.03 0.08 0.01 0.06 0.03 4.03 0.002 Manganese 0.02 0.003 0.02 0.03 0.01 0.03 0.02 2.30 0.03 

Mercury 0.45 0.87 0.45 0.91 0.26 0.74 0.67 1.13 0.92 

Nickel 0.01 0.005 0.02 0.04 0.01 0.03 0.02 4.03 0.11 Selenium 0.62 0.13 1.00 0.27 0.71 1.00 0.17 6.36 0.32 Hydrogen Chloride Note 1 0.93 Note 1 

Hydrogen Fluoride Note 1 1.00 Note 1 

Note 1 - See Note 2 in Table C-19.

Note 1 - These EMFs were developed from emission tests that examined HCl and HF emissions through an ESP.
mm
Appendix D - Discussion of the Methodology Used to Develop 

Nationwide Emission Totals
xX
D-1

D.1 INTRODUCTION

To estimate emissions of hazardous air pollutants (HAPs) from

fossil-fuel–fired electric utility units (>25 MWe), the EPA developed

the emission factor program (EFP). This program incorporates unit

configuration data from individual units as well as emission testing

data to compute estimated emissions. An explanation of the program

and several assumptions about the data and how they were used are

described here.

D.2 PROGRAM OPERATION

Emissions of HAPs considered in this study consist of two types:

trace elements and organic compounds. Trace elements exist in the

fuel when fired, while the organic HAPs are formed during combustion

and postcombustion processes. Different programing methods are

required for handling the two types of HAPs. Program diagrams for

modeling trace element emissions are shown in Figure D-1 for coal and

Figure D-2 for oil and gas. The two figures differ only in treatment

of the fuel before the trace elements reach the boiler. Figure D-3

shows the program diagram for modeling organic HAP emissions.

D.3 DATA SOURCES

The EFP was built to accept data from two sources. The first is

a data input file containing plant configurations, unit fuel usage,

and stack parameters. This input file was based on the Utility Data

Institute/Edison Electric Institute (UDI/EEI) Power Statistics

database (1991 and 1994 editions) and an extract from Production

Costs, U.S. Gas Turbine and Combined-Cycle Power Plants (for 1994

estimates). The UDI/EEI database is composed of responses from

electric utilities to the yearly updated U.S. Department of Energy

(DOE) Energy Information Administration (EIA) Form EIA-767.

The second data file is the emission modification factor (EMF)

database. This database contains information from emissions tests

conducted by the Electric Power Research Institute (EPRI), DOE, and

the electric utility industry. The program first searches the input

file for the type of fuel burned and the amount of fuel consumed per

year in an individual unit. If the fuel type is coal, the EFP then

looks for the coal’s State of origin. Origin is important because the

trace elements in coal are addressed by coal type (bituminous,

subbituminous, and lignite) and State of origin in the U.S. 

Geological Survey (USGS) database, which analyzed core and channel

samples (3,331 samples) of coal from the top 50 (1990 or later)

economically feasible coal seams in the United States. 

D.4 OPERATIONAL STATUS OF BOILERS

The operational status of units was taken from the UNIT_90.dbf

file of the EEI/UDI Power Statistics database (1991 edition addressing

1990 data and 1996 edition addressing 1994 data). Only units that 
*Taken from UDI/EEI data.

UDI/EEI plant configuration

information

Bituminous Subbituminous Lignite

USGS coal data

(by State and coal type)

USGS coal data

(by State and coal type)

Apply coal cleaning

factor

Trace elements (TE) to boiler

No cleaning factor

What is

boiler type?*

Apply boiler TE emission

modification factor

What is

particulate

matter (PM) control

type?*

Apply the PM control TE

emission modification factor

What is

the SO control

type?*

kg/yr of specific trace element

exiting unit stack

Figure D-1. Trace elements in coal.

2

Apply SO control TE emission

modification factor

2

D-2
*Taken from UDI/EEI data.

Trace elements (TE) to boiler

What is

boiler type?*

Apply boiler TE emission

modification factor

What is

particulate

matter (PM) control

type?*

Apply the PM control TE

emission modification factor

What is

the SO control

type?*

kg/yr of specific trace element

exiting unit stack

Apply SO control TE emission

modification factor

Figure D-2. Trace elements in oil and natural gas.

2

2

Oil

Used fuel oil No. 6 (residual) for

all oil types

Trace elements in oil taken

from plant testing

Trace elements in natural gas

taken from plant testing

(only two sets of data)

Natural gas

Used a denisty of 8.2 lb/gal for

feed rate calculation

UDI/EEI plant configuration

information

D-3
Type of fuel

burning?

Oil Coal Gas

Obtain unit's fuel

consumption

Obtain unit's fuel

consumption

Obtain unit's fuel

consumption

For individual HAPs,

find the median

lb/trillion Btu emission

factor for a specific

HAP1

For individual HAPs,

find the median

lb/trillion Btu emission

factor for a specific

HAP2

For individual HAPs, find

the geometric mean

kg/10 9 cu ft emission

factor for a specific

HAP3

Individual fuel

consumption x

emission factor x heat

content of 150,000

Btu/gal

Coal type?

Butuminous Subbituminous Lignite

Individual fuel

consumption x

emission factor x

higher heating value

for bituminous coal

(12,688 Btu/lb coal)

Individual fuel

consumption x

emission factor x

higher heating value

for subbituminous

coal (9,967 Btu/lb

coal)

Individual fuel

consumption x

emission factor x

higher heating value

for lignite coal (6,800

Btu/lb coal)

Figure D-3. Organic emissions.

Convert into kg/yr

stack emission for

HAP

1Only oil-fired units were used to obtain these emission factors.

2Only coal-fired units were used to obtain these emission factors.

3Only gas-fired units were used to obtain these emission factors.

Individual fuel

consumption x

emission factor

D-4
D-5

were listed as either operational or on standby were used in the EFP. 

One hundred fifty-one units were listed as being on standby in the

1990 EEI/UDI Power Statistics database but were actually on indefinite

standby and thus did not emit any HAPs. These units were excluded

from the nationwide emissions totals in Appendix A. Other units

listed on indefinite standby (i.e., no fuel burned) were excluded from

1994 emission estimates.

Only coal-fired, oil-fired, and natural gas-fired units were

included in the EFP. This decision was made because units using these

fuels make up an overwhelming majority of the fossil-fuel–fired

electric utility units with a capacity >25 MWe.

Anthracite was disregarded as a fuel because of the limited

number of units burning this type of coal. Four units burning 1

anthracite coal (in 1990 and 1994) were assigned to burn bituminous

coal for program computations. 

The EEI/UDI database had a number of gaps in the fuel consumption

data. Some of these gaps were filled by data supplied voluntarily by

the industry. To address the remaining gaps, EPA plotted the

available data and fitted point-slope equations to estimate fuel

consumption. These equations involved plotting nameplate megawatts 2

(modified to take into account the unit’s capacity factor) against

fuel usage. If the fuel usage and the unit capacity factor in 1990

were not given, 1989 fuel consumption data were used. If 1989 data

were not available, the geometric mean of the 1980–1988 EEI fuel

consumption data was used. When all other options had been tried

unsuccessfully, an average fuel consumption of units rated within 

±5 MW of the unit with unknown fuel usage was used. Similar problems

in the 1994 UDI/EEI database were solved by using 1990 data where

possible and by similar methods to those stated above when not

possible.

Capacity factors were taken from the UDI/EEI database for as many

units as possible. If the above linear equation or (±5 MWe)

estimating procedure were used, then the capacity factor for the unit

(with unknown fuel consumption) would fit an industry norm for that

size unit and fuel type.

Limestone is used in circulating fluidized bed (CFB) combustors

to control sulfur dioxide (SO ). Early in the program’s development, 2

the EPA sought to address limestone’s contribution to trace metal

emissions. Based on the fact that limited trace metal data were

available and that there were only 19 listed CFB units in the country

in 1990, limestone’s effect was disregarded for 1990 and 1994.3

Utility units may burn coal that originated from several States;

however, in the EFP each coal-fired unit was assigned a single State

of coal origin. The State of origin used in the EFP was the State 4

that contributed the highest percentage of the unit’s coal. Coal
D-6

consumption by State for each utility was found in volumes of Cost and

Quality of Fuels for Electric Utility Plants for 1990 and 1994. 5 6

D.5 BOILER CONFIGURATION

The EPA received 51 emissions tests conducted by EPRI, DOE, and

industry in time for inclusion in the EFP for 1990 emission estimates. 

A further seven reports were available for the 1994 emission

estimates. Because of this limited sample, not all boiler

configurations, particulate, and SO control types could be sampled. 2

To estimate the emissions from all units in the United States, the

substitution of unknown units into units with known EMFs was

necessary. After studying the tested EMFs, the following patterns

were observed. Coal-fired unit emissions seemed to be affected by

whether the unit had a dry- or wet-bottom furnace. Oil-fired unit

emissions seemed to be affected by whether or not the unit had

nitrogen oxides (NO ) control. Since only one type of gas-fired boiler x

was tested, all gas-fired units obtained their EMFs from this type of

unit.7

One of the emission test reports that analyzed an oil-burning,

tangentially fired (with NO control) unit contained information on two x

trace metals. Because this was the only unit of its kind to be

tested, it was necessary to substitute the trace metal data of another

similar unit (one having NO control) for which more data were x

collected. The EMFs of the oil burning, front-fired unit (with NO x

control) were averaged (by geometric mean) into the unit along with

the two trace metal concentrations found in the tangentially-fired

boiler. Because there were organic HAP concentration numbers

available for the tangentially-fired boiler, these numbers were

maintained without modification.

No conventional emission testing (multimetals, volatile organic

sampling train [VOST], semi-VOST) was done on combined-cycle gas

turbines. The Fourier Transform Infrared (FTIR) system was used to

test a combined-cycle gas turbine unit for organic HAPs, but few HAPs

were found. Combined-cycle gas turbines were categorized as

conventional gas-fired units to address their emissions.

Testing by FTIR was also done on one example each of pulverized

coal-, circulating fluidized bed-, oil-, and conventional gas-fired

boilers and a combined-cycle gas turbine. However, the EPA decided

not to use the data in developing estimated emissions.

Of the test reports received, four contained data that were not

feasible for use in the EFP because the test contractors did not or

could not test between the boiler and the particulate control device. 

The result was a test containing only a fuel analysis and stack

emission numbers.

One EPRI emission test report (identified as EPRI Site 10)

contained only one sample run instead of the normal three runs. 
D-7

Because only two emission test reports on CFBs (including Site 10)

were available, the EPA decided to use these data.

For 1990 emission estimates, units were deemed dual-fuel-firing

units if they fired more than 10 percent of at least one other fuel. 

Dual-fuel firing emissions were modeled by splitting the dual-firing

units (only oil- and gas-fired units) into two separate units with

emissions exiting from the same stack. If the unit were listed as an

oil-fired unit, its oil consumption rate and configuration were used

to obtain its HAP emission rates for oil. The unit in question was

then split into a gas-fired portion by using its gas consumption rate

and changing its boiler type to the equivalent gas-fired type. This

method was considered the most equitable way at the time to represent

dual-fuel-fired emissions, for both trace metals and organic HAPs

created by either oil-fired or gas-fired boilers, respectively.

For 1994 estimates, where units fired more than one type of fuel,

emissions were modeled for each fuel. If the unit was listed as a

coal-fired unit, its coal consumption rate and configuration were used

to obtain its HAP emission rate for coal. Similarly, if the unit also

fired oil, its oil consumption rate and equivalent boiler

configuration were used to obtain its HAP emission rate for oil. If

gas were also fired, the unit’s gas consumption rate and equivalent

gas-fired boiler type were used obtain its HAP emission rate for gas.

Substitution was also performed on particulate control and SO2

control devices. Particulate control was addressed in one of six

ways: electrostatic precipitator, cold-side (ESP,CS); ESP, hot-side

(ESP,HS); ESP, cold-side, controlling an oil-fired unit (O-ESP,CS);

fabric filter (FF); particulate scrubber; or no control.

Cold-side ESPs are placed after the air preheaters, while

hot-side ESPs are placed before the air preheaters. The UDI/EEI

database reported several units with combination HS/CS ESPs. These

were units with separate ESPs before and after their air preheaters. 

Although one such unit was tested for HAP emissions, during the

majority of its testing the cold-side ESP was turned off. Therefore,

the data for this unit were used to develop hot-side ESP EMFs for the

EFP. Because more data were available on ESP,CS devices, and because

units controlled by HS/CS ESPs had a cold-side ESP as their last

particulate matter (PM) control device, HS/CS ESPs were projected to

behave like cold-side ESPs in terms of trace metal emissions. In

assigning the boiler type for coal-fired units, when there was no

information on whether the unit had NO control, it was assumed that x

the unit had no NO control and the unit was assigned TANGDRYNONOX x

boiler factors. The boiler and PM control device data were assigned

in this manner for units that had hot-side ESPs since the temperature

at the inlet to the hot-side ESP was approximately 700 F, whereas the

temperature at the inlet to cold-side ESPs were typically around 
D-8

300 F. The assignment was made to account for any effect that the

approximately 700 F temperature might have on air toxic emissions. 

Table D-1 shows the boiler substitutions and associated PM control

devices.

Emission modification factors for particulate control by

scrubbers were derived from data on controlling trace elements by one

venturi scrubber used for combined SO and PM control. Particulate 2

matter scrubbers use water only, while flue gas desulfirization units

(FGDs) use water and a reagent (lime, limestone, etc). Although the

presence of this reagent could cause the FGD to affect HAPs

differently from the PM scrubbers, the EPA believes that the small

number of PM scrubbers (<5) should not cause U.S. aggregate emissions

to be adversely effected.

Mechanical collectors (multicyclones) are used either as

precollection devices, before FFs or ESPs, or as primary collection

devices for some oil-fired plants. No HAP emissions testing was done

exclusively on mechanical collectors. Since mechanical collectors

were projected to have little or no effect on reducing HAPs because of

their ineffectiveness at removing small particles, units with only

multicyclones were determined to have no control effect on HAPs in the

program.

In the EFP, devices for controlling SO emissions were classified 2

as either WETSCRUB (containing all types of wet FGDs) or DRYSCRUB

(containing all types of spray dryers/dry scrubbers). This

substitution was necessary due to the lack of test data on a variety

of wet FGD and dry scrubber types. Also, the EMFs include data from

units tested with bypasses operating when using a bypass is normal

operation, i.e., emissions from bypassed gas are included in the EFP

results.

D.6 STACK CHARACTERISTICS

Stack data for 1990 in the UDI/EEI from some electric utility

units were incomplete. Some of these gaps were due to the database

reporting stack parameters from a shared stack on only one of the

plant’s units instead of reporting on both. The shared stack

parameters were completed for these sister units. Next, an industry

contractor made contact with a number of utility plants to retrieve

missing stack data. This information was useful but still incomplete. 

The remaining gaps in stack parameter data were filled by either

(1) finding a sister unit of the same configuration (and site, if

possible) in order to duplicate its stack data, or (2) using the

original EEI/UDI stack data to create a set of equations to estimate

the relationships between stack height and gas flow, stack exit

temperature, and exit velocity from stack diameter, respectively. 

These linear equations (point-slope) were specific to coal-, oil-,

gas-, and combined cycle gas turbine-fired units. A spreadsheet 
D-9

Table D-1. Boiler Substitutions Used in the Emission Factor

Program (EFP)

Facility Boiler Boiler Used in EFP Associated Control Device

CFBDRYNONOX CFBDRYNOX BAGHOUSE

CFBDRYNOX CFBDRYNOX BAGHOUSE

COMCYCLNONOX G-FRONTDRYNONOX NO CONTROL

COMCYCLNOX G-FRONTDRYNONOX NO CONTROL

COMCYCLNOX O-FRONTDRYNONOX NO CONTROL

CYCLWETNONOX CYCLWETNONOX ESP, CS

CYCLWETNONOX TANGDRYNONOX ESP, HS

CYCLWETNOX CYCLWETNONOX BAGHOUSE

FRONTDRYNONOX FRONTDRYNONOX BAGHOUSE

FRONTDRYNONOX TANGDRYNONOX ESP, HS

FRONTDRYNOX FRONTDRYNONOX BAGHOUSE

FRONTDRYNOX TANGDRYNONOX ESP, HS

FRONTWETNONOX CYCLWETNONOX BAGHOUSE

FRONTWETNONOX TANGDRYNONOX ESP, HS

OPPODRYNONOX TANGDRYNOX ESP, HS

OPPODRYNONOX OPPODRYNOX BAGHOUSE

OPPODRYNOX TANGDRYNOX ESP, HS

OPPODRYNOX OPPODRYNOX BAGHOUSE

OPPOWETNONOX TANGDRYNONOX ESP, HS

OPPOWETNONOX OPPOWETNONOX ESP, CS

OPPOWETNOX TANGDRYNONOX ESP, HS

OPPOWETNOX OPPOWETNONOX BAGHOUSE

REARDRYNONOX FRONTDRYNONOX ESP, CS

REARDRYNONOX TANGDRYNONOX ESP, HS

STOKDRYNONOX CYCLWETNONOX ESP, CS

TANGDRYNONOX TANGDRYNONOX ESP, HS

TANGDRYNONOX TANGDRYNONOX BAGHOUSE

TANGDRYNOX TANGDRYNOX ESP, HS

TANGDRYNOX TANGDRYNOX BAGHOUSE

TANGWETNONOX CYCLWETNONOX ESP, CS

TANGWETNOX TANGDRYNONOX ESP, HS

TANGWETNOX OPPOWETNONOX BAGHOUSE

UNKNOWN TANGDRYNONOX ESP, CS

VERTDRYNONOX VERTDRYNOX BAGHOUSE

VERTWETNONOX CYCLWETNONOX ESP, CS

G-CYCLWETNONOX G-FRONTDRYNONOX NO CONTROL

G-CYCLWETNOX G-FRONTDRYNONOX NO CONTROL

G-FRONTDRYNONOX G-FRONTDRYNONOX NO CONTROL

G-FRONTDRYNOX G-FRONTDRYNONOX NO CONTROL

G-FRONTWETNONOX G-FRONTDRYNONOX NO CONTROL

G-HORZDRYNONOX G-FRONTDRYNONOX NO CONTROL

G-OPPODRYNONOX G-FRONTDRYNONOX NO CONTROL

(continued)
Table D-1. (Continued)

Facility Boiler Boiler Used in EFP Associated Control Device

D-10

G-OPPODRYNOX G-FRONTDRYNONOX NO CONTROL

G-REARDRYNONOX G-FRONTDRYNONOX NO CONTROL

G-REARDRYNOX G-FRONTDRYNONOX NO CONTROL

G-REARDRYNOX G-FRONTDRYNONOX NO CONTROL

G-TANGDRYNONOX G-FRONTDRYNONOX NO CONTROL

G-TANGDRYNOX G-FRONTDRYNONOX NO CONTROL

G-TANGWETNONOX G-FRONTDRYNONOX NO CONTROL

G-TANGWETNONOX G-FRONTDRYNONOX NO CONTROL

G-UNKNOWN G-FRONTDRYNONOX NO CONTROL

G-UNKNOWNDRYNONOX G-FRONTDRYNONOX NO CONTROL

G-VERTDRYNONOX G-FRONTDRYNONOX NO CONTROL

G-VERTWETNONOX G-FRONTDRYNONOX NO CONTROL

O-CYCLDRYNONOX O-FRONTDRYNONOX NO CONTROL

O-CYCLEDRYNOX O-FRONTDRYNONOX NO CONTROL

O-CYCLWETNONOX O-FRONTDRYNONOX O-ESP, CS

O-CYCLWETNOX O-FRONTDRYNONOX NO CONTROL

O-FRONTDRYNONOX O-FRONTDRYNONOX NO CONTROL

O-FRONTDRYNOX O-FRONTDRYNOX NO CONTROL

O-FRONTWETNONOX O-FRONTDRYNONOX NO CONTROL

O-FRONTWETNOX O-FRONTDRYNOX O-ESP, CS

O-OPPODRYNONOX O-OPPODRYNONOX NO CONTROL

O-OPPODRYNOX O-FRONTDRYNOX NO CONTROL

O-OPPOWETNONOX O-FRONTDRYNONOX NO CONTROL

O-REARDRYNONOX O-FRONTDRYNONOX NO CONTROL

O-REARDRYNOX O-FRONTDRYNOX O-ESP, CS

O-TANGDRYNONOX O-TANGDRYNONOX NO CONTROL

O-TANGDRYNOX O-TANGDRYNOX NO CONTROL

O-TANGWETNONOX O-FRONTDRYNONOX NO CONTROL

O-UNKNOWNDRYNONOX O-FRONTDRYNONOX NO CONTROL

O-VERTDRYNONOX O-FRONTDRYNONOX O-ESP, CS

Notes to Table D-1: The following conventions are used for naming boilers in this table. The name describes primary fuel used,

firing type, bottom type, and presence or absence of nitrogen oxides control.

Fuel: No prefix = coal Boiler bottom type:

 G- = gas DRY Dry ash

 O- = oil WET Molten ash

 CFB boilers are coal fired

Boiler firing type: NONOX No nitrogen oxides control

 CFB Circulating fluidized bed NOX Nitrogen oxides control by any of several

COMCYC Combined cycle means as specified in the UDI database

CYC Cyclone

FRONT Front Control device:

HORZ Horizontal BAGHOUSE Fabric filter

OPPO Opposed ESP, CS Cold-side electrostatic precipitator (after the air

REAR Rear preheater)

TANG Tangential ESP, HS Hot-side electrostatic precipitator (before the air

UNKNOWN Firing type not specified in Utility Data Institute preheater)

database (UDI) O-ESP, CS Cold-side electrstatic precipitator applied to an

VERT Vertical oil-fired boiler

Nitrogen oxides control:

NO CONTROL Particulate controls not applied to this boiler

(usually gas or oil fired)
D-11

procedure was developed to enter a stack height for a unit and use

four separate equations to estimate the other parameters.

A few stack latitudes or longitudes not addressed in either the

original EEI database or the contractor’s research were found by

calling the operators of the utility plants in question. 

Because only 1990 estimated emissions were used for risk

analysis, missing stack data or latitude/longitude data for 1994 were

not addressed.

D.7 TRACE ELEMENT CONCENTRATION IN FUEL

The USGS database contains concentrations of trace elements that

were extracted from coal in the ground but does not include analyses

of coal shipments. The concentrations of trace elements in coal in

the ground and in coal shipments to utilities may differ because, in

the process of preparing a coal shipment, some of the mineral matter

in coal may be removed. Since approximately 77 percent of the Eastern

and Midwestern bituminous coal shipments are cleaned to meet customer 8

specifications on heat, ash, and sulfur content, a coal cleaning

factor was applied to most bituminous coals in the EFP. Two 9

exceptions were bituminous coals from Illinois and Colorado, for which

analyses were on an as-shipped basis representative of the coal to be

fired. Tables at the end of this appendix (D-8 and D-9) list trace

element concentrations in fuel and coal cleaning factors,

respectively, as used in the EFP.

Arithmetic averages of the concentrations of trace elements were

determined from the USGS database by State of coal origin, and the 10

average concentrations were then used in the EFP. (Note: statewide

data were not separated by coal region, and statewide averages were

not weighted by coal production within the State.) Two sets of

concentration data exist for coal that originated from Arizona and one

set for coal that originated from Washington. The two sets of 11

Arizona data were averaged with data for Colorado, Utah, and New

Mexico coal. The trace element concentrations for coals from Arizona,

Louisiana, and Washington were needed for five, one, and two utility

units, respectively. Because no data were available for coal from

Louisiana, data from Texas lignite coal were used to represent the

concentration of trace elements in Louisiana coal.12

Additional data on the concentrations of the trace elements in

utility coal shipments were received from ARCO Coal Company on 145

samples of Wyoming coal and on 30 samples of bituminous Colorado

coal, and from the Illinois State Geological Survey (ISGS) on 34 13

samples of Illinois cleaned coal. Arithmetic averages of the trace 14

element concentrations provided by ARCO Coal Company and ISGS were

converted to an as-received basis and used directly, without

application of cleaning factors, in the EFP. In summary, USGS data 15

were used for all States with the following exceptions: two sets of

USGS data for Arizona coals were averaged with USGS data for Colorado,
D-12

Utah, and New Mexico coals; Texas lignite data were substituted for

Louisiana coal; Arco data were used for Wyoming coals and for Colorado

bituminous coals; and ISGS data were used for Illinois coals.

For a unit that burned bituminous coal, the kilogram/year (kg/yr)

feed rate of trace elements to the boiler was determined from the

average trace element concentration in the coal, a coal cleaning

factor, and the annual fuel consumption rate. No coal cleaning

factors were applied to lignite and subbituminous coals (see Equations

1 and 2 in Table D-2).

If the fuel type was oil, the program accessed a database

containing the arithmetic average of trace element concentrations in

residual oil (see Figure D-2). Each concentration data point was the

arithmetic average of repeated measurements, and at least one of the

repeated measurements had to be a detected concentration (see

discussion of nondetected data in section D.12). Because trace

element data were available only on residual oil-fired units, and

since 95 percent of the oil-fired units burn residual oil, all units

were assumed to burn residual oil. Although densities of residual

oils vary, an average density of 8.2 lb/gal was chosen for the feed

rate calculation for oil. The concentration data and density were

used, as shown in Equation 3 in Table D-2, to calculate a kg/yr rate

of each trace element entering the unit’s oil-fired boiler. Oil-fired

organic HAP exit concentration calculations included a 

150,000-Btu/gallon heating value for oil.

An emission rate for each organic HAP emitted from gas-fired

units was extracted from the test reports. Only two test reports for

gas-fired units analyzed organic HAPs, and a geometric mean emission

rate of each observed organic HAP was used. This rate in kilogram

HAP/10 cubic feet was then multiplied by the unit’s gas consumption to 9

obtain a kilogram HAP/year stack emission rate of each specific HAP

(see Equation 4 in Table D-1). This result was equivalent to a stack

emission because there were no PM control or SO control devices on 2

gas-fired units. The geometric mean of the concentrations were

averaged and used in the gas-fired boiler calculations (see 

Figure D-2). The few trace elements found in the gas database were

estimated by this procedure. Fuel gas density was assumed to follow

the ideal gas law. 

Total quantities burned for each type of fuel (coal, gas, or oil)

and each type of boiler (as shown in Table D-1) are shown in

Table D-3. Coal consumption is further quantified by coal rank.

D.8 HYDROGEN CHLORIDE AND HYDROGEN FLUORIDE CONCENTRATION IN FUEL

To obtain hydrogen chloride (HCl) or hydrogen fluoride (HF)

emissions from the boiler, emission factors were derived by performing

mass balances for chloride and fluoride, then converting these

balances to the equivalent levels of HCl or HF throughout the boiler

system. For example, for each lb/hr of chloride in the feed coal at 16
Unit coal consumption tons coal

yr

x 2,000 lb

ton

x

0.454 kg

lb

x trace element ppmw (by State of coal origin)

1,000,000

x coal cleaning factor  kg trace element

yr

Unit coal consumption tons coal

yr

x 2,000 lb

ton

x

0.454 kg

lb

x trace element ppmw (by State of coal origin)

1,000,000

 kg trace elem

yr

Unit oil consumption bbls oil

yr

x

42 gal

bbl

x

8.2 lb

gal

x

0.454 Kg

lb

x trace element ppmw

1,000,000

 kg trace element

yr

Unit gas consumption cf

yr

x trace element ppmw

1,000,000

x

1 atm x lb mole x R

0.7302 cf x atm

x

1

(460 68)R

x trace element molecular we

lb mole

x

0.454 kg

lb

 kg trace element

yr

D-13

Table D-2. Computer Calculations (Inorganic HAPs)

Feed rate of trace elements to the boiler

For bituminous coal: Equation 1For lignite and subbituminous coal:Equation 2For oil:Equation 3For natural gas:Equation 4
D-14 (continued)

Table D-3. Fuel Consumption by Type of Boiler

Facility Boiler Fuel Consumed Primary Fuel a

CFBDRYNONOX 1,157 Coal

CFBDRYNOX 1885 Coal

COMCYCLNONOX 166,767 Gas

CYCLWETNONOX 56,723 Coal

CYCLWETNOX 3,098 Coal

FRONTDRYNONOX 5,713 Coal

FRONTDRYNOX 59,686 Coal

FRONTWETNONOX 10,045 Coal

OPPODRYNONOX 121,708 Coal

OPPODRYNOX 138,237 Coal

OPPOWETNONOX 6,217 Coal

OPPOWETNOX 29,590 Coal

REARDRYNONOX 733 Coal

STOKDRYNONOX 731 Coal

TANGDRYNONOX 217,266 Coal

TANGDRYNOX 138,085 Coal

TANGWETNONOX 1,546 Coal

TANGWETNOX 16,606 Coal

UNKNOWN 93 Coal

VERTDRYNONOX 5,008 Coal

VERTWETNONOX 792 Coal

G-CYCLWETNONOX 10,245 Gas

G-CYCLWETNOX 789 Gas

G-FRONTDRYNONOX 278,662 Gas

G-FRONTDRYNOX 269,047 Gas

G-FRONTWETNONOX 4,262 Gas

G-HORZDRYNONOX 190 Gas

G-OPPODRYNONOX 409,885 Gas

G-OPPODRYNOX 662,119 Gas

G-REARDRYNONOX 758 Gas

G-REARDRYNOX 29,747 Gas

G-TANGDRYNONOX 416,979 Gas

G-TANGDRYNOX 221,779 Gas

G-TANGWETNONOX 6,287 Gas

G-UNKNOWNDRYNONOX 1,337 Gas

G-VERTDRYNONOX 139 Gas

O-CYCLDRYNONOX 898 Oil

O-CYCLWETNONOX 228 Oil

O-CYCLWETNOX 792 Oil

O-FRONTDRYNONOX 2,772 Oil

O-FRONTDRYNOX 35,929 Oil

O-FRONTWETNONOX 182 Oil
Table D-3. (Continued)

Facility Boiler Fuel Consumed Primary Fuel a

D-15

O-FRONTWETNOX 134 Oil

O-OPPODRYNONOX 7,967 Oil

O-OPPODRYNOX 6,589 Oil

O-OPPOWETNONOX 766 Oil

O-REARDRYNONOX 1,407 Oil

O-REARDRYNOX 1,287 Oil

O-TANGDRYNONOX 32,654 Oil

O-TANGDRYNOX 7,304 Oil

O-TANGWETNONOX 9,226 Oil

Coal in thousands of tons per year, gas in millions of cubic feet per year, oil in thousands of barrels per year. Quantities do not a

include gas and oil used as starting or temporary fuels in boilers that normally burn other fuels as the primary fuel.

Note: Nationwide total fuel consumption from these boilers is, in the units given above:

Coal 815,135 (18.0 x 10 Btu) 15

Bituminous 405,013 (10.3 x 10 Btu) 15

Subbituminous 330,978 ( 6.6 x 10 Btu) 15

Lignite 79,128 ( 1.1 x 10 Btu) 15

Gas 2,708,342 (2.84 x 10 Btu) 15

Oil 146,148 (0.92 x 10 Btu) 15

one of the test sites, 0.63 lb/hr of HCl was found in the gas stream

leaving the boiler. Similarly for HF, the boiler emissions were 0.64

lb/hr for each lb/hr of fluoride in the coal. For ease of

programming, the HCl and HF emissions were addressed starting in the

fuel. This programming was done by multiplying the chloride and

fluoride concentrations in the fuel constituents by 0.63 or 0.64,

respectively. The resulting numbers allowed direct conversion into

boiler emissions that could be further modified for systems with

PM control or SO control. For the 1990 emission estimates, before 2

obtaining further test reports, the factors were 0.61 for HCl and 0.56

for HF.

The chloride concentrations were not available for coals from the

following States: Alaska, Illinois, Indiana, Iowa, Missouri, Utah, and

Washington. Chloride concentrations were assigned, as shown in

Table D-4, for coals originating from these States.17

D.9 EMISSION MODIFICATION FACTORS FOR INORGANIC HAPS

The HCl and HF emission factors were addressed in the fuel;

therefore, all HCl and HF boiler EMFs for all fuel types, were made

equal to 1 in the EFP.

To address the partitioning of the HAP stream through the

combustion and pollution control process, partitioning factors 
D-16

Table D-4. Assigned Chloride ppmw and HCl ppmw Concentrations in

Coal, by State of Coal Origin 17

State assigned HCl ppmw coal

Conversion of assigned ppmw chloride to Assigned ppmw HCl in

Alaska 54 x 0.63 = 34.0

Illinois 1,136 x 0.63 = 715.7

Indiana 1,033 x 0.63 = 650.8

Iowa 1,498 x 0.63 = 943.7

Missouri 1,701 x 0.63 = 1,071.6

Utah 220 x 0.63 = 138.6

Washington 104 x 0.63 = 65.5

(EMFs) were developed from inorganic HAP testing data. The EMFs are

fractions of the amount of a HAP compound exiting a device (boiler or

air pollution control device [APCD]) divided by the amount of the same 

HAP compound entering that device. These EMFs were averaged by 18

taking the geometric mean of similar devices (e.g., all oil-fired

tangential boilers, all cold-side ESPs). Geometric means were used

because of the presence of outlying data points, the small amount of

data, and the general fit of the data to a log-normal curve. These

geometric means were then applied to the kg/yr feed rates entering the

boiler, the effect of which either reduced or left unchanged the

emissions that passed through them. Those EMFs calculated as being

greater than 1.0 (i.e., more material exiting a device than entering

it) are set to equal 1.0. The EMFs are based on emission test report

data collected and analyzed after 1990.

Nearly all EMFs were computed from three data samples before and

three data samples after the particular device. When all six data

samples for a particular EMF computation were nondetects, the EPA

decided to disregard the EMF. As such, EMFs were computed when there

was at least one detected sample among the six measured samples.

The EMFs were computed with data from different test reports but

for similar devices (i.e., cold-side ESPs, front-fired boilers in oilfired units). The data from coal-fired units were not segregated by

State of coal origin. The EMFs from devices are generally segregated

into only coal-, oil-, or gas-fired bins.

The EFP itself uses EMFs to partition the emissions as they

proceed from the fuel through the unit to the stack exit as follows. 

The average concentrations of metallic HAPs in an individual fuel by

State (based on USGS data) were multiplied by the amount of fuel that

the unit burned in 1990 and 1994. After accounting for variables such

as coal cleaning (bituminous coal only) and coal type (higher heating
D-17

value), the emission concentration of an inorganic HAP was converted

into an emission rate in kg/yr entering the boiler. The emission rate

entering the boiler was then modified by EMFs for the boiler, the

particulate control device (when applicable), and the SO control 2

device (when applicable).

As stated above, these geometric mean EMFs were then applied to

the fuel HAP concentration estimates and the kg/yr fuel feed rates

entering the boiler, which either reduced or left unchanged the

emissions that passed through it, depending on the value of the EMF.

Table C-1 (Appendix C) shows two sets of EMF data for the DOE

Niles test site. One unit with NO control is in a section designated x

without NO control. This apparent anomaly occurs because the NO x X

control method used, SCR, is a postcombustion NO control and does not x

effect the boiler EMFs. The data are labeled this way to identify the

data obtained from a separate test report.

Appendix C contains all of the EMFs used to develop the unit 

emission estimates for inorganic HAPs. 

D.10 ACID GAS HAPS

The method used with HCl or HF emissions allowed direct

conversion from coal chlorine or fluorine content into boiler

emissions, as described in section D.8, that could be further modified

for systems with PM control or SO control. 2

Hydrochloric acid and HF EMFs for PM and SO control devices were 2

developed with data from test reports in which contractors conducted

tests individually for HCl, chlorine, HF, and fluorine before and

after each control device. These tests were in contrast to the

remaining tests for which HCl and HF values were estimated or omitted

rather than measured. 

The next steps after obtaining amounts of HCl or HF leaving the

boiler were to construct EMFs for the PM control device, then for the

SO control device. Using chlorine as an example, the measured amount 2

of HCl entering the PM control device (in kg/yr with suitable

conversion factors) was compared with the measured amount of HCl

leaving the PM control device. Using these two quantities, an EMF was

formed as described in section D.9. 

In the final step, EMFs were formed for HCl and HF through the

SO control device based on the measured mass of HCl or HF entering 2

that device (leaving the PM control device) and the mass measured at

the exit of the SO control device. However, a modification was 2

required to account for flue-gas bypass around the SO control device. 2

A portion of the flue gas is bypassed to maintain SO removal at the 2

minimum permitted amount. This action is used as a means of reducing

energy required to reheat the flue gas for effective plume rise from

the stack. In developing the HCl and HF EMFs for wet FGDs and dry
D-18

scrubbers, the effect of flue gas bypass was treated by analyzing

utility test data from the four plants (of eight tested) that used

bypasses, reviewing municipal waste incinerator results that showed a

typical HCl or HF removal efficiency of 95 percent, and having

discussions with industry representatives. Based on the 95 percent

removal efficiency coupled with the measured values for quantity of

flue gas bypassed, an industry average effective value for flue gas

bypass in 1994 was estimated. The value was assumed to be 15 percent

(17 percent for 1990 data) for wet FGDs and 14 percent (for 1990 and

1994 data) for dry scrubber systems. These assumptions were used only

in the development of HCl and HF EMFs. Future wet FGDs are not 19

expected to use flue-gas bypass in normal operation.

D.11 ORGANIC HAPS

Because organic HAPs were not always tested at the entrance and

exit of each control device in the emissions testing, all organic HAP

emissions were addressed by examining the test data and determining

the concentration of a particular HAP exiting the stack. Organic HAP

concentrations were obtained from emission test reports. Table D-5

gives the equations used to estimate organic HAP emissions from coal-,

oil-, and gas-fired boilers.

Organic stack emissions from coal-fired boilers were first

determined on an emission factor basis (lb/trillion Btu) to account

for different coal heating values, then converted to a rate basis

(kg/yr of individual HAP). This procedure was necessary because

different coal ranks had different heating values. For example, it

would require burning more lignite to achieve the same heat input to

the boiler as burning bituminous coal. These values were determined

as averages for each type of coal (see Table D-6).20

If stack emission or APCD exit emission data were unavailable or

reported as nondetected, and, if at least one-third of the data

samples at the inlet of the APCD were detected concentrations, EPA

used organic emissions at the inlet of the APCD and accounted for the

effect of the APCD with EMFs. Where nondetected data were used (in

about 40 percent of the individual congener test series), the same

procedure as for EMFs (described below) was followed to establish a

calculated mean for the (usually) three test values. For each

individual organic HAP observed in testing, a median concentration was

obtained to represent the average value of the usually small and

scattered data set. For example, of the coal-fired boilers tested for

dioxins/furans, 12 had detected values for one or more of the

congeners. This fuel-specific median concentration was then

individually multiplied by each utility unit’s fuel consumption. The

result was a fuel-specific emission rate for all organic HAPs that

were observed at least once during testing.
Unit coal consumption tons coal

yr

x 2,000 lb coal

ton coal

x Coal HHV Btu

lb coal (see Table D4) x 11012 trillion Btu Btu

x (Median emission factor) lb HAP

trillion Btu

x

0.454 kg HAP

lb HAP

kg HAP

yr

Unit oil consumption bbls oil

yr

x

42 gal oil

bbl oil

x (HHV for residual oil) 150,000 Btu gal oil x 11012 trillion Btu Btu

x (Median emission factor) lb HAP

trillion Btu

x

0.454 kg HAP

lb HAP

kg HAP

yr

Unit gas consumption cf gas

yr

x

1

109 billion cf x (Geometric mean emission factor) kg HAP billion cf kg HAP yr

D-19

Table D-5. Computer Calculations (Organic HAPs)

All non-trace-element HAPs emitted from the stack

Coal-fired boiler:Equation 1Oil-fired boiler:Equation 2Gas-fired boiler:Equation 3
D-20

Table D-6. Average Higher Heating Values of Coal21

Class and group character than Less than than Less than than Less than Average a Agglomerating greater greater greater

Fixed carbon limits, % Volatile matter limits, Calorific value limits,

(dry, mineral-matter- % (dry, mineral-matterfree basis) free basis)

Btu/lb (moist,

b

mineral-matter-free

basis)

Equal or Equal or Equal or

I. Bituminous

1. Low-volatile bituminous coal commonly 78 86 14 22 --- ---

agglomeratingc

2. Medium-volatile bituminous  69 78 22 31 --- ---

coal

3. High-volatile A bituminous coal  --- 69 31 --- 14,000 --- 14,000 d

4. High-volatile B bituminous coal  --- --- --- --- 13,000 14,000 13,500 d

5. High-volatile C bituminous coal  --- --- --- --- 11,500 13,000 12,250

 High-volatile C bituminous coal agglomerating --- --- --- --- 10,500 11,500 11,000

Average of Averages (Value used in EFP for bituminous coal) 12,688

II. Subbituminous

1. Subbituminous A Coal nonagglomerating --- --- --- --- 10,500 11,500 11,000

2. Subbituminous B Coal  --- --- --- 9,500 10,500 10,000

3. Subbituminous C Coal  --- --- --- --- 8,300 9,500 8,900

Average of Averages (Value used in EFP for subbituminous coal) 9,967

III. Lignitic

1. Lignite A nonagglomerating --- --- --- --- 6,300 8,300 7,300

2. Lignite B  --- --- --- --- --- 6,300 6,300

Average of Averages (Value used in EFP for lignite coal) 6,800

This classification does not include a few coals, principally nonbanded varieties, which have unusual physical and chemical properties and which a

come within the limits of fixed carbon or calorific value for high-volatile and subbituminous ranks. These excluded coals either contain less than

48 percent dry, mineral-matter-free fixed carbon or have more than 15,500 moist, mineral-matter-free Btu per pound.

Moist refers to coal containing its natural inherent moisture but not including visible water on the surface of the coal. b

It is recognized that there may be nonagglomerating varieties in these groups of the bituminous class, and there are notable exceptions in high- c

volatile C bituminous group.

Coals having 69 percent or more fixed carbon on the dry, mineral-matter-free basis shall be classified by fixed carbon, regardless of calorific d

value.
D-21

D.12 TREATMENT OF NONDETECTED DATA IN THE DEVELOPMENT OF EMFS

In the raw data taken from the test reports, the EPA used a

protocol to analyze detected and nondetected compounds in the test

samples. The protocol is as follows:

 When all values for a specific compound are above the

detection limit, the mean arithmetic concentration is

calculated using the reported quantities.

 For results that include values both above and below the

detection limit (with the detection limit shown in

parentheses), one half of the detection limit is used for

values below the detection limit to calculate the mean. For

example:

Analytical values Calculation Mean value 

10,12,ND(8) (10+12+[8/2])/3 8.7 ND

The calculated mean cannot be smaller than the largest

detection limit value. In the following example, the

calculated mean is 2.8. This quantity is less than the

largest detection limit, so the reported mean becomes ND(4).

Analytical values Calculation Mean value 

5,ND(4),ND(3) (5+[4/2]+[3/2])/3 ND(4)

 When all sample results are less than the detection limit,

the data are not used.

D.13 MODEL CHANGES FOR ESTIMATES IN THE YEAR 2010

Emission estimates for 2010 were derived from the same basic 1990

model described above. However, changes to input files were made to

accommodate expected changes in fuel usage, generating capacity, and

responses to Phases I and II of the 1990 amendments under Title IV. 

The details of these expected changes, except for coal usage, are

described in section 2.7 of this report. Details of coal usage are

described below.

To approximate the projected increase in the use of coal, and

particularly lower sulfur coals, the 2010 coal consumption was determined

as follows. First the estimated overall increase in electric utility coal

consumption was determined (37 percent). Then, instead of using an 22

overall percentage increase for each coal-fired unit, a factor was derived

for each coal State of origin to represent the expected increase or

decrease in consumption for that State’s coal in 2010. The 1990 coal

consumption was then multiplied by the 2010 factor, listed in Table D-7,

that corresponded to the State of coal origin assigned to each unit. 23

Tables D-8a,b,c and D-9 list trace element concentrations in fuel

and coal cleaning factors, respectively, as used in the EFP.
For each coal-fired unit, the 2010 coal consumption was determined as follows: The 1990 coal a

consumption was multiplied by the 2010 factor that corresponded to the State of coal origin assigned to the

unit.

D-22

Table D-7. Coal Consumption Scaling Factors for 2010

State of coal origin 2010 factora

Kentucky 1.27

Pennsylvania 1.23

West Virginia 1.24

Maryland 0.872

Ohio 0.872

Alabama 1.41

Louisiana 1.41

Texas 1.41

Virginia 1.41

Illinois 1

Indiana 1

Iowa 1

Kansas 1

Missouri 1

Oklahoma 1

Alaska 1.599

Arizona 1.599

Colorado 1.599

Montana 1.599

New Mexico 1.599

North Dakota 1.599

Utah 1.599

Washington 1.599

Wyoming 1.599
D-23

Table D-8a. Trace Element Concentrations in Coal

State Coal type Compound Concentration, ppmw

AK Subbituminous ANTIMONY 1.90

ARSENIC 3.00

BERYLLIUM 0.50

CADMIUM 0.15

CHROMIUM 20.00

COBALT 5.00

CHLORINE 53.93

FLUORINE 95.00

LEAD 5.40

MANGANESE 88.00

MERCURY 0.07

NICKEL 10.00

SELENIUM 1.60

AL Bituminous ANTIMONY 1.82

ARSENIC 53.00

BERYLLIUM 1.88

CADMIUM 0.06

CHROMIUM 22.80

COBALT 8.20

CHLORINE 380.00

FLUORINE 127.00

LEAD 7.00

MANGANESE 41.00

MERCURY 0.19

NICKEL 17.50

SELENIUM 1.88

AR Lignite ANTIMONY 1.17

ARSENIC 4.30

BERYLLIUM 2.40

CADMIUM 0.29

CHROMIUM 16.90

COBALT 6.00

CHLORINE 142.00

FLUORINE 63.00

LEAD 9.80

MANGANESE 119.00

MERCURY 0.25

NICKEL 11.80

SELENIUM 5.00

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-24

AZ Subbituminous ANTIMONY 0.47

ARSENIC 2.10

BERYLLIUM 1.10

CADMIUM 0.10

CHROMIUM 4.60

COBALT 2.10

CHLORINE 200.00

FLUORINE 79.00

LEAD 9.00

MANGANESE 27.00

MERCURY 0.07

NICKEL 4.80

SELENIUM 1.50

CO Bituminous ANTIMONY 0.91

ARSENIC 1.34

BERYLLIUM 0.36

CADMIUM 0.18

CHROMIUM 1.89

COBALT 1.03

CHLORINE 92.97

FLUORINE 98.78

LEAD 5.44

MANGANESE 10.83

MERCURY 0.07

NICKEL 1.25

SELENIUM 0.87

CO Subbituminous ANTIMONY 0.35

ARSENIC 1.03

BERYLLIUM 0.84

CADMIUM 0.08

CHROMIUM 4.10

COBALT 1.60

CHLORINE 118.00

FLUORINE 99.00

LEAD 3.50

MANGANESE 32.00

MERCURY 0.14

NICKEL 7.90

SELENIUM 0.89

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-25

IA Bituminous ANTIMONY 2.30

ARSENIC 12.00

BERYLLIUM 1.88

CADMIUM 14.00

CHROMIUM 12.10

COBALT 10.00

CHLORINE 1498.36

FLUORINE 77.00

LEAD 68.00

MANGANESE 259.00

MERCURY 0.19

NICKEL 31.00

SELENIUM 3.60

IL Bituminous ANTIMONY 0.82

ARSENIC 6.78

BERYLLIUM 1.31

CADMIUM 0.98

CHROMIUM 12.66

COBALT 3.19

CHLORINE 1136.07

FLUORINE 84.14

LEAD 24.51

MANGANESE 33.74

MERCURY 0.08

NICKEL 12.74

SELENIUM 1.72

IN Bituminous ANTIMONY 1.40

ARSENIC 10.10

BERYLLIUM 2.82

CADMIUM 0.49

CHROMIUM 15.40

COBALT 5.20

CHLORINE 1032.79

FLUORINE 65.00

LEAD 10.90

MANGANESE 38.00

MERCURY 0.11

NICKEL 17.90

SELENIUM 2.17

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-26

KS Bituminous ANTIMONY 0.85

ARSENIC 25.00

BERYLLIUM 1.47

CADMIUM 10.00

CHROMIUM 10.10

COBALT 15.00

CHLORINE 2500.00

FLUORINE 64.00

LEAD 111.00

MANGANESE 160.00

MERCURY 0.19

NICKEL 41.00

SELENIUM 2.70

KY Bituminous ANTIMONY 1.13

ARSENIC 19.10

BERYLLIUM 3.17

CADMIUM 0.16

CHROMIUM 16.30

COBALT 6.60

CHLORINE 1139.00

FLUORINE 86.00

LEAD 10.60

MANGANESE 32.00

MERCURY 0.15

NICKEL 17.50

SELENIUM 3.83

LA Lignite ANTIMONY 0.82

ARSENIC 3.70

BERYLLIUM 1.90

CADMIUM 0.15

CHROMIUM 11.40

COBALT 3.30

CHLORINE 115.00

FLUORINE 83.00

LEAD 5.50

MANGANESE 141.00

MERCURY 0.19

NICKEL 7.80

SELENIUM 6.00

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-27

MD Bituminous ANTIMONY 0.81

ARSENIC 26.00

BERYLLIUM 2.01

CADMIUM 0.14

CHROMIUM 26.70

COBALT 11.00

CHLORINE 914.00

FLUORINE 107.00

LEAD 10.00

MANGANESE 13.00

MERCURY 0.42

NICKEL 22.00

SELENIUM 3.80

MO Bituminous ANTIMONY 1.60

ARSENIC 10.00

BERYLLIUM 2.01

CADMIUM 0.80

CHROMIUM 12.20

COBALT 6.70

CHLORINE 1701.64

FLUORINE 60.00

LEAD 67.00

MANGANESE 99.00

MERCURY 0.17

NICKEL 23.00

SELENIUM 4.20

MT Bituminous ANTIMONY 0.69

ARSENIC 7.00

BERYLLIUM 0.52

CADMIUM 0.08

CHROMIUM 3.10

COBALT 1.50

CHLORINE 80.00

FLUORINE 104.00

LEAD 3.00

MANGANESE 37.00

MERCURY 0.09

NICKEL 3.90

SELENIUM 0.70

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-28

MT Lignite ANTIMONY 0.92

ARSENIC 18.00

BERYLLIUM 1.04

CADMIUM 0.11

CHROMIUM 0.94

COBALT 0.80

CHLORINE 67.00

FLUORINE 159.00

LEAD 4.80

MANGANESE 68.00

MERCURY 0.12

NICKEL 4.00

SELENIUM 0.72

MT Subbituminous ANTIMONY 0.69

ARSENIC 7.00

BERYLLIUM 0.52

CADMIUM 0.08

CHROMIUM 3.10

COBALT 1.50

CHLORINE 80.00

FLUORINE 104.00

LEAD 3.00

MANGANESE 37.00

MERCURY 0.09

NICKEL 3.90

SELENIUM 0.70

ND Lignite ANTIMONY 0.58

ARSENIC 8.40

BERYLLIUM 0.82

CADMIUM 0.11

CHROMIUM 7.00

COBALT 2.70

CHLORINE 110.00

FLUORINE 34.00

LEAD 3.73

MANGANESE 86.00

MERCURY 0.13

NICKEL 4.10

SELENIUM 0.79

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-29

NM Subbituminous ANTIMONY 1.07

ARSENIC 1.80

BERYLLIUM 2.70

CADMIUM 0.16

CHROMIUM 6.00

COBALT 2.65

CHLORINE 95.00

FLUORINE 87.00

LEAD 31.00

MANGANESE 45.00

MERCURY 0.06

NICKEL 4.60

SELENIUM 1.94

OH Bituminous ANTIMONY 0.81

ARSENIC 23.20

BERYLLIUM 2.39

CADMIUM 0.12

CHROMIUM 14.30

COBALT 0.90

CHLORINE 719.00

FLUORINE 92.00

LEAD 7.30

MANGANESE 28.30

MERCURY 0.22

NICKEL 14.90

SELENIUM 3.80

OK Bituminous ANTIMONY 0.69

ARSENIC 24.00

BERYLLIUM 0.86

CADMIUM 0.10

CHROMIUM 15.00

COBALT 6.20

CHLORINE 267.00

FLUORINE 77.00

LEAD 10.00

MANGANESE 74.00

MERCURY 0.17

NICKEL 17.00

SELENIUM 1.80

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-30

PA Bituminous ANTIMONY 1.23

ARSENIC 32.10

BERYLLIUM 2.45

CADMIUM 0.10

CHROMIUM 20.10

COBALT 7.90

CHLORINE 1096.00

FLUORINE 78.00

LEAD 10.80

MANGANESE 23.50

MERCURY 0.29

NICKEL 20.40

SELENIUM 3.55

TX Lignite ANTIMONY 0.82

ARSENIC 3.70

BERYLLIUM 1.90

CADMIUM 0.15

CHROMIUM 11.40

COBALT 3.30

CHLORINE 115.00

FLUORINE 83.00

LEAD 5.50

MANGANESE 141.00

MERCURY 0.19

NICKEL 7.80

SELENIUM 6.00

UT Bituminous ANTIMONY 0.23

ARSENIC 0.89

BERYLLIUM 0.61

CADMIUM 0.08

CHROMIUM 7.70

COBALT 2.70

CHLORINE 219.67

FLUORINE 57.00

LEAD 3.90

MANGANESE 8.00

MERCURY 0.04

NICKEL 4.10

SELENIUM 2.00

(continued)
Table D-8a. (Continued)

State Coal type Compound Concentration, ppmw

D-31

VA Bituminous ANTIMONY 0.93

ARSENIC 11.00

BERYLLIUM 1.66

CADMIUM 0.05

CHROMIUM 12.50

COBALT 6.30

CHLORINE 930.00

FLUORINE 74.00

LEAD 5.80

MANGANESE 19.00

MERCURY 0.14

NICKEL 11.20

SELENIUM 2.70

WA Subbituminous ANTIMONY 0.30

ARSENIC 1.50

BERYLLIUM 1.10

CADMIUM 0.11

CHROMIUM 0.70

COBALT 4.70

CHLORINE 103.28

FLUORINE 14.00

LEAD 2.80

MANGANESE 41.00

MERCURY 0.06

NICKEL 7.90

SELENIUM 0.40

WV Bituminous ANTIMONY 0.93

ARSENIC 10.60

BERYLLIUM 2.78

CADMIUM 0.10

CHROMIUM 15.30

COBALT 7.20

CHLORINE 1216.00

FLUORINE 58.00

LEAD 7.20

MANGANESE 19.10

MERCURY 0.16

NICKEL 14.20

SELENIUM 3.97

(continued)
State Coal type Compound Concentration, ppmw

D-32

WY Subbituminous ANTIMONY 0.73

ARSENIC 0.69

BERYLLIUM 0.18

CADMIUM 0.13

CHROMIUM 2.82

COBALT 0.87

CHLORINE 118.30

FLUORINE 43.70

LEAD 2.07

MANGANESE 5.65

MERCURY 0.08

NICKEL 2.17

SELENIUM 0.51

Table D-8b. Trace Element Concentrations in Fuel Oil (for 1994

estimates)

Trace Element Concentration in Oil, ppmw

Arsenic 0.306

Beryllium 0.027

Cadmium 0.020

Chromium 0.31

Cobalt 1.63

Chlorine 131

Fluorine 17.5

Lead 1.41

Manganese 0.35

Mercury 0.0092

Nickel 26

Selenium 0.095

Table D-8c. Trace Element Concentrations in Gas

Trace Element Concentration in gas mg/m3

Arsenic 0.000963

Cobalt 0.100

Lead 0.100

Mercury 0.0000024

Nickel 0.0500
D-33

Table D-9. Coal Cleaning Factors for Bituminous Coals Used in

the Emission Factors Programa

Constituent Cleaning factor

Antimony 0.715

Arsenic 0.554

Beryllium 0.711

Cadmium 0.624

Chromium 0.512

Cobalt 0.537

Chlorine 0.496

Fluorine 0.496

Lead 0.449

Manganese 0.382

Mercury 0.790

Nickel 0.568

Selenium 0.745

Applying the cleaning factors to United States Geographical Survey (USGS) constituent concentrations for a

bituminous coals from the States named below results in new, lower constituent concentrations (modified USGS

concentrations), which are used in the emission factors program.

Note: States to which applied: AL, IA, IN, KS, KY, MD, MO, OH, OK, PA, UT, VA, WV
D-34

1. Utility Data Institute. EEI Power Statistics Data Base. 

Washington, D.C. 1992. (excluding nonutility units)

2. Memorandum from Cole, J. D., Research Triangle Institute to

Maxwell, W.H., Environmental Protection Agency. February 3,

1993. Addressing fuel consumption gaps in the EEI power

statistics data base data.

3. Ref. 1.

4. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. July 14,

1993. State of coal origin used in the computer emission

program.

5. Energy Information Administration. Cost and Quality of Fuels for

Electric Utility Plants 1991. U.S. Department of Energy,

Washington, D.C. August 1992. 

6. Energy Information Administration. Cost and Quality of Fuels for

Electric Utility Plants 1994. U.S. Department of Energy,

Washington, D.C. July 1995.

7. Memorandum from Turner, J. H., RTI to Cole, J. D., RTI. April

18, 1994. Collapsing utility boiler types for emissions

modeling.

8. Akers, D., C. Raleigh, G. Shirley, and R. Dospoy, The Effect of

Coal Cleaning on Trace Elements, Draft Report, Application of

Algorithms. Prepared for EPRI by CQ Inc., February 11, 1994.

9. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. April 5,

1994. Proposed coal cleaning factors.

10. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. March 19,

1993. Review of the U.S. Geological Survey data.

11. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. July 8,

1993. Assignment of concentration of trace elements in coals

from Arizona, Louisiana, and Washington.

12. Ref. 11.

13. Letter from Bowling, C.M., ARCO Coal Company to Maxwell, W. H.,

EPA. June 9, 1993. Concerning the use of USGS data to represent

the concentration of trace elements in coal shipments.

D.14 REFERENCES
D-35

14. Demir, I., R. D. Harvey, R. R. Rush, H. H. Damberger, C. Chaven,

J. D. Steele, W. T. Frankie, and K. K. Ho, Characterization of

Available Coal from Illinois Mines, draft report. Illinois

State Geological Survey file number to be assigned. December 28,

1993. 

15. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. April 29,

1994. As-shipped coal data and how gaps in the USGS and UDI

database were filled for the computer emission program.

16. Memorandum from Turner, J. H., RTI, to Cole, J. D., RTI. 

December 7, 1997. Methodology for determining 1994 HCl and HF

concentrations from utility boilers.

17. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. May 27,

1994. USGS data gaps in chloride concentrations for seven

states.

18. Memorandum from Cole, J. D., RTI, to Maxwell, W. H., EPA. March

31, 1994. Emission factor memorandum.

19. Memorandum from Cole, J.D., RTI to Maxwell, W. H., EPA. May 9,

1994. Emission modification factors for HCl and HF including FGD

system bypass.

20. Singer, J. G. ed. Combustion Fossil Power, 4th ed. Combustion

Engineering, Incorporated, Windsor, CT. 1991. p 2-3, modified

table.

21. Ref. 19, 20.

22. Economic Analysis of the Title IV Requirements of the 1990 Clean

Air Act Amendments. Prepared for the U.S. EPA, Office of Air and

Radiation, Acid Rain Division. Prepared by ICF Resources

Incorporated. February 1994.

23. Memorandum from Heath, E., RTI to Maxwell, W. H., EPA. May 9,

1994. Proposed method to account for utilities switching to

lower sulfur coals in 2010.
mm
Appendix E - Health Effects Summaries: Overview
x
E-1

Appendix E contains summaries of health effects data for seven

hazardous air pollutant (HAPs) emitted from utilities (i.e., arsenic,

chromium, nickel, mercury, hydrogen chloride, hydrogen fluoride, and

dioxins). Radionuclides are discussed in Chapter 9 of the interim

report. All of the numbers presented in these summaries are subject

to change, if EPA obtains new data in the future indicating that the

risk is higher or lower than that currently being considered. For

more information on health effects, readers can refer to the

referenced sources at the end of Appendix E. Also, health effects

information for these HAPs and other HAPs can be obtained from the

EPA’s Integrated Risk Information System or from an EPA document 1

titled Health Effects Notebook for Hazardous Air Pollutants. Each 2

summary, except the one for mercury, contains the following sections:

E.1 INTRODUCTION

E.2 CANCER EFFECTS

E.3 NONCANCER EFFECTS

E.3.1 Acute (Short-Term)

E.3.2 Chronic (Long-Term)

E.3.3 Reproductive and Developmental

The following is a discussion of the information contained in each of

these sections:

E.1 INTRODUCTION 

This section presents a brief overview of the chemical, with

information on its chemistry, physical properties, and major uses. If

available, EPA’s National Ambient Air Quality Standard (NAAQS) and/or

Maximum Contaminant Level Goal (MCLG) or Maximum Contaminant Level

(MCL) are also presented in this section. EPA’s NAAQS are legally

enforceable air standards set under the Clean Air Act Amendments of

1990; these are health-based standards with considerations such as

economics and technical feasibility factored in. EPA’s MCLGs are

nonenforceable health goals that are set at levels at which no known

or anticipated adverse health effects occur and that allow an adequate

margin of safety. Maximum contaminant levels are legally enforceable

drinking water standards which are set as close to the MCLGs as

feasible. 

E.2 CANCER EFFECTS 

The results of available cancer studies in animals and/or humans

are presented in this section. In addition, the EPA’s cancer weightof-evidence classification system is included. EPA uses a weight-ofevidence, three-step procedure to classify the likelihood that the

chemical causes cancer in humans. In the first step, the evidence is

characterized separately for human studies and for animal studies. 

The human studies are examined considering the validity and

representativeness of the populations studied, any possible

confounding factors, and the statistical significance of the results

of the studies. The animal studies are evaluated to decide whether

biologically significant responses have occurred and whether the
E-2

responses are statistically significant increases in treated versus

control animals. Secondly, the human and animal evidence is combined

into an overall classification. This classification is based on an

analysis of both the human and animal evidence, considering the number

and quality of both types of studies. In the third step, the

classification is adjusted upward or downward, based on an analysis of

other supporting evidence. Supporting evidence includes structureactivity relationships (i.e., the structural similarity of a chemical

to another chemical with known carcinogenic potential), studies on the

metabolism and pharmacokinetics of a chemical, and short-term genetic

toxicity tests. The result is that each chemical is placed into one

of the following six categories:

Group Description

A Known human carcinogen

B1 Probable human carcinogen, limited human data are available

B2 Probable Human carcinogen, sufficient evidence in animals and

inadequate or no evidence in humans

C Possible human carcinogen

D Not classifiable as to human carcinogenicity

E Evidence of noncarcinogenicity for humans

This section also includes information on the inhalation cancer

risk and oral unit cancer risk. If EPA has calculated both inhalation

and oral unit cancer risk values, then this section is divided into

two subsections. 

The inhalation unit risk estimate (IURE) for the chemical is the

estimated increased probability of a person’s developing cancer from

breathing air containing a concentration of 1 microgram pollutant per

cubic meter (g/m ) of air for 70 years. The IURE is derived using 3

mathematical models that assume a nonthreshold approach: i.e., there

is some risk of cancer occurring at any level of exposure. The

methods used to derive these values typically result in an "upper

bound" estimate; i.e., the true risk is unlikely to exceed this value

and may be lower. However, some unit risk estimates are not "upper

bound" estimates but rather are based on a "maximum likelihood"

estimate (e.g., arsenic).

The risk-specific dose, which is an estimate of the dose

corresponding to a specified level of cancer risk, is also included. 

This section presents risk-specific doses corresponding to a one-in-amillion and one-in-a-hundred-thousand excess risk attributed to

exposure to the chemical. This means that EPA has estimated that if

an individual were to breathe air containing these concentrations of

the chemical, over his or her lifetime, that person would

theoretically have no more than a one-in-a-million or one-in-a-
E-3

hundred-thousand increased chance of developing cancer as a direct

result of breathing air containing the chemical. 

If available, the oral unit cancer risk is also presented. Both

the oral cancer risk and the corresponding risk-specific dose are

developed for an exposure of 70 years to the chemical through the

drinking water. The oral unit risk estimate (OURE) is the estimated

increased risk of cancer for drinking for 70 years 2 liters/day of

water that contains a concentration of 1 g of pollutant per liter. 

It is expressed in units of g/L.

E.3 NONCANCER EFFECTS

E.3.1 Acute (Short-Term) 

Results from acute animal tests or acute human studies are

presented in this section. Acute animal studies usually report an

estimated median lethal dose (LD ) or median lethal concentration 50

(LC ). This is the dose (or concentration) estimated to kill 50 50

percent of the experimental animals. Results from these tests are

divided into the following toxicity categories:

Lethality Extreme High Moderate Low

Oral LD50  50 mg/kg 50 to 500 mg/kg 500 to 5,000 mg/kg 5,000 mg/kg

Dermal LD50 200 mg/kg 200 to 2,000 mg/kg 2,000 to 20,000 mg/kg 20,000 mg/kg

Inhalation LC50 200 mg/m 200 to 2,000 mg/m 2,000 to 20,000 mg/m 20,000 mg/m 3 3 33

Source: U.S. EPA. Office of Pesticide Programs, Registration and Classification Procedures, Part II. Federal Register. 40:28279.

Acute human studies usually consist of case reports from

accidental poisonings. These case reports often help to define the

levels at which acute toxic effects are seen in humans.

E.3.2 Chronic (Long-Term)

This section summarizes the major chronic noncarcinogenic effects

seen from exposure to the chemical. Chronic animal studies usually

range from 90 days to 2 years. Human studies investigating effects

ranging from exposure of a few years to a lifetime are also included. 

In addition, subchronic studies may be included in this section. 

Subchronic studies are usually animal studies of several weeks to 90

days.

The Inhalation Reference Concentration (RfC) is presented in this

section. The RfC is an estimate (with uncertainty spanning perhaps an

order of magnitude) of the daily exposure of a chemical to the human

population by inhalation (including sensitive subpopulations) that is

likely to be without deleterious effects during a lifetime of

exposure. The RfC is derived based on the assumption that thresholds

exist for noncancer effects; i.e., there is a level below which no

toxic effects would occur. The RfC is calculated as follows: EPA

reviews many human and/or animal studies to determine the highest dose
E-4

level tested at which the critical adverse effect does not occur—i.e.,

the no observed adverse effect level (NOAEL)—or the lowest dose level

at which the critical adverse effect is observed, the lowest observed

adverse effect level (LOAEL). The NOAEL from an animal study is

adjusted for exposure duration and respiratory tract differences

between animals and humans. EPA then applies uncertainty factors to

adjust for the uncertainties in extrapolating from animal data to

humans (10), and for protecting sensitive subpopulations (10). Also,

a modifying factor is applied to reflect professional judgment of the

entire data base.

The RfC is not a direct or absolute estimator of risk, but rather

a reference point to gauge the potential effects. Doses at or below

the RfC are not likely to be associated with any adverse health

effects. However, exceedance of the RfC does not imply that an

adverse health effect would necessarily occur. As the amount and

frequency of exposures exceeding the RfC increases, the probability

that adverse effects may be observed in the human population also

increases. The RfC is expressed in milligrams of pollutant per cubic

meter of air (mg/m ). If available, the Oral Reference Dose (RfD) is 3

also presented in this section. The RfD is the oral equivalent of the

RfC.

EPA’s confidence in the RfC and/or RfD is also presented in this

section. EPA ranks each RfC and RfD as low, medium, or high in three

areas: (1) confidence in the study on which the RfC or RfD was based;

(2) confidence in the data base; (3) overall confidence in the RfC or

RfD. All three rankings are presented in this section.

E.3.3 Reproductive and Developmental

This section presents the results of reproductive and

developmental studies on the effects of the chemical in animals and

humans. Reproductive effects are those effects that adversely affect

the female or the male reproductive system. Examples in the female

include reduced fertility, a decrease in the survival of offspring,

and alterations in the reproductive cycle. Male reproductive effects

include a decrease in sperm count or an increase in abnormal sperm

morphology. Developmental effects are adverse effects on the

developing organism that result from exposure prior to conception

(either parent), during prenatal development, or postnatally to the

time of sexual maturation. Examples include altered growth, death of

the developing organism, and malformations or birth defects.

Reproductive and developmental effects may be observed after shortterm or long-term exposure to the chemical, as some effects can be

attributed to one time or short-term exposures during a critical

biological cycle.

E.4 ARSENIC HEALTH EFFECTS SUMMARY

Arsenic is a naturally occurring element in the earth’s crust

that is usually found combined with other elements. Arsenic combined

with elements such as oxygen, chlorine, and sulfur is referred to as

inorganic arsenic; arsenic combined with carbon and hydrogen is
0.0002 g/m (concentration corresponding to a 10 risk level) =10 a 3 -6 -6

(risk level)/4.29 x 10 (g/m ) (unit risk estimate). -3 3 -1

E-5

referred to as organic arsenic. In this health effects summary,

arsenic refers to inorganic arsenic and its associated compounds. 

Organic arsenic compounds, such as arsine gas, are not discussed. EPA

has set a Maximum Contaminant Level (MCL) of 0.05 mg/L for inorganic

arsenic.3

E.4.1 CANCER EFFECTS — Arsenic

There is clear evidence that chronic exposure to inorganic

arsenic in humans increases the risk of cancer. Studies have reported

that inhalation of arsenic results in an increased risk of lung

cancer. In addition, ingestion of arsenic has been associated with an

increased risk of nonmelanoma skin cancer and bladder, liver, and lung

cancer. No information is available on the risk of cancer in humans

from dermal exposure to arsenic. Animal studies have not clearly

associated arsenic exposure, via ingestion exposure, with cancer. No

studies have investigated the risk of cancer in animals as a result of

inhalation or dermal exposure.4

EPA has classified inorganic arsenic in Group A - Known Human

Carcinogen. For arsenic, the Group A classification was based on the

increased incidence in humans of lung cancer through inhalation

exposure and the increased risk of skin, bladder, liver, and lung

cancer through drinking water exposure.5

E.4.1.1 Inhalation Cancer Risk for Arsenic. EPA used the

absolute-risk linear extrapolation model to estimate the inhalation

unit risk for inorganic arsenic. Five studies on arsenic-exposed

copper smelter workers were modeled for excess cancer risk. All five

studies showed excess risks of lung cancer that were related to the

intensity and duration of exposure and the duration of the latency

period. The estimates of unit risk obtained from the five studies

were in reasonably good agreement, ranging from 1.25 x 10 to 7.6 x 10 -3 -

 (g/m ) . Using the geometric mean of these data, EPA calculated an 3 3 -1

inhalation unit risk estimate of 4.29 x 10 (g/m ) (EPA). Based on -3 3 -1 6

this unit risk estimate, EPA estimates that if an individual were to

breathe air containing arsenic at 0.0002 g/m over his or her entire 3a

lifetime (70 years), that person would theoretically have an increased

chance of one in a million of developing cancer as a direct result of

breathing air containing this chemical. Similarly, EPA estimates that

breathing air containing 0.002 g/m would result in an increased 3 

chance of up to one in a hundred thousand of developing cancer. EPA

has high confidence in the arsenic cancer unit risk estimate for

inhalation exposure because the studies examined a large number of

people, the exposure assessments included air measurements and urinary

arsenic measurements, and lung cancer incidence was significantly

increased over expected values.7

The Electric Power Research Institute (EPRI) has proposed a

revision to EPA’s IURE for inorganic arsenic. EPRI used standard EPA
E-6

risk assessment methodology to recalculate the estimated risk. They

calculated a new unit risk of 1.43 x 10 (g/m ) , which is one-third -3 3 -1

the value on IRIS presented above. EPRI’s risk estimate is based on

updated exposure data from an epidemiology study of workers at a

smelter in Tacoma, Washington, which indicated that the workers were

much more highly exposed than previously thought. EPRI also used

results from a recent Swedish smelter study.8

E.4.1.2 Oral Cancer Risk for Arsenic. To estimate the risks

posed by ingesting arsenic, EPA obtained in Taiwan concerning skin

cancer incidence, age, and level of exposure via drinking water. In

37 villages that had obtained drinking water for 45 years from artisan

wells with various elevated levels of arsenic, 40,421 individuals were

examined for hyperpigmentation, keratosis, skin cancer, and blackfoot

disease (gangrene of the extremities caused by injury to the

peripheral vasculature). The local well waters were analyzed for

arsenic, and the age-specific cancer prevalence rates were found to

correlate with both local arsenic concentrations and age (duration of

exposure). Based on these data, although EPA has not presented the

calculations for the oral unit risk estimate for arsenic, they did 9

propose that a unit risk estimate of 5 x 10 (μg/L) from oral -5 -1

exposure to arsenic in drinking water be used.10

The Taiwan cancer data have the following limitations: (1) the

water was contaminated with substances such as bacteria and ergot

alkaloids in addition to arsenic; (2) total arsenic exposure was

uncertain because of intake from the diet and other sources; (3) early

deaths from blackfoot disease may have led to an underestimate of

prevalence; and (4) there was uncertainty concerning exposure

durations. Due to these limitations, and also because the diet,

economic status, and mobility of individuals in Taiwan are different

from those of most U.S. citizens, EPA has stated "the uncertainties

associated with ingested inorganic arsenic are such that estimates

could be modified downwards as much as an order of magnitude, relative

to risk estimates associated with most other carcinogens."11

E.4.2 Noncancer Effects — Arsenic

E.4.2.1 Acute (Short-Term) Effects for Arsenic. Arsenic has

been recognized as a human poison since ancient times, and large

doses, approximately 600 g/kg/day or higher, taken orally have

resulted in death. Oral exposure to lower levels of arsenic has

resulted in effects on the gastrointestinal system (nausea, vomiting);

central nervous system (headaches, weakness, delirium); cardiovascular

system (hypotension, shock); and the liver, kidney, and blood (anemia,

leukopenia). Acute arsenic poisoning of humans, through inhalation

exposure, has resulted in similar effects, including effects on the

gastrointestinal system (nausea, diarrhea, abdominal pain), blood, and

central and peripheral nervous system. The only effect noted from

dermal (skin) exposure to arsenic in humans is contact dermatitis,

with symptoms such as erythma and swelling. This effect has been

noted only at high arsenic levels.12
E-7

Because significant information is available on the acute effects

of arsenic poisoning in humans, few animal studies have been carried

out. The limited available data have shown arsenic to have moderate

to high acute toxicity to animals by the oral route. This is based on

data showing the LD for arsenic to range between 50 and 5,000 mg/kg. 50 13

E.4.2.2 Chronic (Long-Term) Effects for Arsenic. The primary

effect noted in humans from chronic exposure to arsenic, through both

inhalation and oral exposure, is effects on the skin. The inhalation

route has resulted primarily in irritation of the skin and mucous

membranes (dermatitis, conjunctivitis, pharyngitis, and rhinitis),

while chronic oral exposure has resulted in a pattern of skin changes

that include the formation of warts or corns on the palms and soles

along with areas of darkened skin on the face, neck, and back. Other

effects noted from chronic oral exposure include peripheral

neuropathy, cardiovascular disorders, liver and kidney disorders, and

blackfoot disease. No information is available on effects in humans

from chronic low-level dermal exposure to arsenic.14

No studies are available on the chronic noncancer effects of

arsenic in animals, from inhalation or dermal exposure. Oral animal

studies have noted effects on the kidney and liver.15

EPA has established an RfD (Reference Dose) for inorganic arsenic

of 0.0003 mg/kg/day, based on a NOAEL (adjusted to include arsenic

exposure from food) of 0.0008 mg/kg/day, an uncertainty factor of 3,

and a modifying factor of 1. This RfD was based on two studies that 16 17

showed that the prevalence of blackfoot disease increased with both

age and dose for individuals exposed to high levels of arsenic in

drinking water. This same population also displayed a greater

incidence of hyperpigmentation and skin lesions. Other human studies

support these findings, with several studies noting an increase in

skin lesions from chronic exposure to arsenic through the drinking

water. The EPA has not established a RfC for inorganic arsenic.18

EPA has medium confidence in the studies on which the RfD was

based and in the RfD. The key studies were extensive epidemiologic

reports that examined effects in a large number of people. However,

doses were not well characterized, other contaminants were present,

and potential exposure from food or other sources was not examined. 

The supporting studies suffer from other limitations, primarily the

small populations studied. However, the general database on arsenic

does support the findings in the key studies; this was the basis for

EPA’s "medium confidence" ranking of the RfD.19

E.4.2.3 Reproductive and Developmental. Limited information is

available on the reproductive or developmental effects of arsenic in

humans. The only available information consists of several studies

that suggest that women who work in, or live near, metal smelters may

have higher than normal spontaneous abortion rates, and their children

may exhibit lower than normal birth weights. However, these studies

are limited and contain significant uncertainties because they were
E-8

designed to evaluate the effects of smelter pollutants in general and

are not specific for arsenic. 20

Animal studies on arsenic exposure via oral and inhalation routes

have reported that arsenic at very high doses may cause death to the

fetus or birth defects. No information is available on reproductive

or developmental effects of arsenic in animals from dermal exposure. 21

E.5 CHROMIUM HEALTH EFFECTS SUMMARY

Chromium is a metallic element that occurs in the environment in

two major valence states: trivalent chromium (chromium III) and

hexavalent chromium (chromium VI). Chromium VI compounds are much

more toxic than chromium III compounds; chromium III is an essential

element in humans, with a daily intake of 50 to 200 micrograms per day

recommended for an adult, while chromium VI is quite toxic. However,

the human body can detoxify some amount of chromium VI to chromium

III. EPA has set a Maximum Contaminant Level (MCL) of 0.1 mg/L for

total chromium.22

E.5.1 Cancer Effects for Chromium

Epidemiological studies of workers have clearly established that

inhaled chromium is a human carcinogen, resulting in an increased risk

of lung cancer. These studies were not able to differentiate between

exposure to chromium III and chromium VI compounds. No information is

available on cancer in humans from oral or dermal exposure to

chromium.23,24

Animal studies have shown chromium VI to cause lung tumors via

inhalation exposure. No studies are available that investigated

cancer in animals from oral or dermal exposure to chromium VI. 

Chromium III has been tested in mice and rats by the oral route, with

several studies reporting no increase in tumor incidence. No studies

are available on cancer in animals from inhalation or dermal exposure

to chromium III.25,26

EPA has classified chromium VI in Group A - Known Human

Carcinogen. Since the human studies could not differentiate between 27

chromium III and chromium VI exposure, and only chromium VI was found

to be carcinogenic in animal studies, EPA concluded that only chromium

VI should be classified as a human carcinogen. EPA has classified 28

chromium III in Group D — Not Classifiable as to Human

Carcinogenicity.29

 

EPA used the multistage extrapolation model, based on data from

an occupational study of chromate production workers, to estimate the

unit cancer risk for chromium VI. EPA calculated an IURE of 1.2 x 10 -2

(g/m ) . Based upon this unit risk estimate, EPA estimates that if 3 -1 30

an individual were to breathe air containing chromium VI at 0.00008
0.00008 g/m (concentration corresponding to a 10 risk level) = 10 b 3 -6 -6

(risk level)/1.2 x 10 (g/m ) (unit risk estimate). -2 3 -1

E-9

g/m over his or her entire lifetime, that person would theoretically 3 b

have an increased chance of up to a one in one million of developing

cancer as a direct result of breathing air containing this chemical. 

Similarly, EPA estimates that breathing air containing 0.0008 g/m 3

would result in an increased chance of up to one in one hundred

thousand of developing cancer. EPA has not calculated a risk 31

estimate from oral exposure to chromium VI or from inhalation or oral 32

exposure to chromium III.33

EPA has confidence in the risk estimate for chromium VI, based on

the fact that the results of studies of chromium exposure are

consistent across investigators and countries and because a dose

response for lung tumors has been established. However, an

overestimation of risk may exist due to the implicit assumption that

the smoking habits of chromate workers were similar to those of the

general white male population, because it is generally accepted that

the proportion of smokers is higher for industrial workers than for

the general population.34

The International Agency for Research on Cancer (IARC) has stated

that there is sufficient evidence in humans for the carcinogenicity of

chromium VI compounds and inadequate evidence in humans for the

carcinogenicity of chromium III compounds.35

E.5.2 Noncancer Effects

This section presents information from human and/or animal

studies on the acute (short-term), chronic (long-term), and

reproductive/developmental effects of chromium VI and chromium III.

E.5.2.1 Acute (Short-Term) for Chromium. The respiratory tract

is the major target organ for chromium VI following inhalation

exposure in humans. Dyspnea, coughing, and wheezing were reported in

cases in which individual inhaled very high concentrations of chromium

VI. Other effects noted from acute inhalation and oral exposure to

very high concentrations of chromium VI include gastrointestinal and

neurological effects, while dermal exposure causes skin burns.36

Acute animal studies have reported chromium VI to have extreme

toxicity from inhalation and oral exposure. This is based on data

showing the LC for chromium VI to be less than 200 mg/m and the LD 50 50 3

to be less than 50 mg/kg. Chromium III has been shown to have

moderate toxicity from oral exposure, based on LD data in the range 50

of 500 to 5,000 mg/kg. The kidney is the major target organ for

chromium VI acute toxicity in animals, with high doses resulting in

kidney failure. Other target organs include the brain and the liver. 37

E.5.2.2 Chronic (Long-Term) for Chromium. Chronic inhalation

exposure to chromium VI in humans results in effects on the

respiratory tract, with perforations and ulcerations of the septum,
E-10

bronchitis, decreased pulmonary function, pneumonia, asthma, and nasal

itching and soreness reported. Chronic exposure to high levels of

chromium VI by inhalation or oral exposure may also produce effects on

the liver, kidney, gastrointestinal and immune systems, and possibly

the blood. Dermal exposure to chromium VI may cause contact

dermatitis, sensitivity, and ulceration of the skin.38

Limited information is available on the chronic effects of

chromium in animals. The available data indicate that, following

inhalation exposure, the lung and kidney have the highest tissue

levels of chromium. No effects were noted in several oral animal

studies with chromium VI and chromium III.39

EPA has established RfD for chromium VI of 0.005 mg/kg/day, based

upon a NOAEL (adjusted) of 2.4 mg/kg/day, an uncertainty factor of

500, and a modifying factor of 1. This was based on a study of rats, 40

which reported no adverse effects after their exposure to chromium VI

in the drinking water for 1 year. Other studies support these

findings; one study reported no significant effects in female dogs

given chromium VI in the drinking water for 4 years, and a case study

on humans reported no adverse health effects in a family of four who

drank water for 3 years from a private well containing chromium VI at

1 mg/L.41

EPA has low confidence in the study on which the RfD for chromium

(VI) was based and in the RfD. Confidence in the key study was ranked

low due to the small number of animals tested, the small number of

parameters measured, and the lack of toxic effects at the highest dose

tested. The low ranking of the RfD was due to lack of high-quality

supporting studies and the fact that developmental and reproductive

effects are not well studied.42

The RfD for chromium III is 1 mg/kg/day, based upon a NOAEL

(adjusted) of 1,468 mg/kg/day, an uncertainty factor of 1,000, and a

modifying factor of 1. This was based on no effects observed in rats 43

fed chromium III in the diet for 2 years. EPA has low confidence in

the study on which the RfD was based and in the RfD. The low ranking

of the key study was due to the lack of explicit detail on study

protocol and results, while the low ranking of the RfD was due to the

lack of supporting data and the lack of an observed effect level in

the key study. EPA has not established an RfC for chromium III or 44 45

chromium VI.46

E.5.2.3 Reproductive and Developmental for Chromium. Limited

information is available on the reproductive or developmental effects

of chromium in humans. The only available data suggest that exposure

to chromium (VI) by inhalation in women may result in complications

during pregnancy and childbirth. 47

Animal studies have not reported reproductive effects from

inhalation exposure to chromium (VI). Oral studies on chromium (VI)

have reported severe developmental effects in mice such as gross

abnormalities and reproductive effects including decreased litter
E-11

size, reduced sperm count, and degeneration of the outer cellular

layer of the seminiferous tubules. No information is available on the

reproductive or developmental effects of chromium (III) in humans or

animals.48

E.6 HYDROCHLORIC ACID HEALTH EFFECTS SUMMARY

Hydrochloric acid is an aqueous solution of hydrogen chloride gas

and is commercially available in several concentrations and purities. 

Because of impurities, commercial varieties of hydrochloric acid are

generally yellow. Hydrochloric acid is used in refining metal ore, as

a lab reagent, and in the removal of scale from boilers.49

E.6.1 Cancer Effects

Limited information is available on the possible carcinogenic

effects of hydrochloric acid. No information is available on the

cancer risk to humans from exposure to hydrochloric acid. The

carcinogenic effects of combined and separate exposures via inhalation

to formaldehyde and hydrochloric acid were investigated in a study on

rats. No carcinogenic response was observed when rats were exposed

only to hydrochloric acid at concentrations of 10 ppm. No studies 50

have investigated risk of cancer in animals as a result of oral or

dermal exposures. 

EPA has not classified hydrochloric acid with respect to

potential carcinogenicity and has not estimated the unit cancer risk

associated with hydrochloric acid.51

E.6.2 Noncancer Effects — Hydrogen Chloride

E.6.2.1 Acute (Short-Term) Effects for Hydrogen Chloride. The

acute effects on humans exposed by inhalation to hydrochloric acid

include coughing, choking, inflammation and ulceration of the

respiratory tract, chest pain, and pulmonary edema. Oral exposure may

result in corrosion of the mucous membranes, esophagus, and stomach,

with nausea, vomiting, intense thirst, and diarrhea. Dermal contact

with hydrochloric acid can cause burns, ulcerations, and scarring. 52

Animals exposed to 320 parts per million (ppm) for 6 minutes

suffered sensory irritation, while levels of 680 ppm or higher for 1

minute resulted in less severe effects; inhalation of air containing

6,400 mg/m hydrochloric acid for 30 minutes resulted in death from 3

laryngeal spasm, laryngeal edema, or rapidly developing pulmonary

edema. Acute inhalation exposure tests resulted in an LC of 1,108 53 50

ppm for exposed mice and 3,124 ppm for exposed rats (moderate to high

acute toxicity). An LD of 900 mg/kg (moderate acute toxicity) was 50

reported for rabbits exposed orally to hydrochloric acid. No 54

information is available on effects in animals from acute dermal

exposure to hydrochloric acid.

E.6.2.2 Chronic (Long-Term) Effects for Hydrogen Chloride. In

humans, cases of gastritis, chronic bronchitis, dermatitis, and

photosensitization have been reported among individuals exposed
E-12

occupationally to hydrochloric acid. No other data are available 55

specifically on the effects of long-term human exposure dermally or

via inhalation or ingestion.

In animals, the only study of the effects of long-term inhalation

of hydrochloric acid reported epithelial or squamous hyperplasia of

the nasal mucosa, larynx, and trachea. In a 90-day inhalation study,

decreased body weight gains, minimum to mild rhinitis, nasal cavity

lesions, and eosinophilic globules in the epithelial lining of the

nasal tissues were reported in test animals. No studies are 56

available on the long-term effects on animals from low-level oral or

dermal exposures to hydrochloric acid.

EPA has established an RfC for hydrochloric acid of 0.02 mg/m . 3

This concentration was based on a rate study in which hyperplasia of

the nasal mucosa, larynx, and trachea were seen. An uncertainty

factor of 300 was applied to an LOAEL of 6.1 mg/m . The EPA has low 3 57

confidence in the study, database, and RfC because the study used only

one dose and the database did not provide any additional chronic or

reproductive studies. 58

E.6.2.3 Reproductive and Developmental for Hydrogen Chloride. 

No information is available on reproductive or developmental effects

of hydrochloric acid in humans. In animal studies in which female

rats were exposed via inhalation prior to mating and during gestation,

severe dyspnea, cyanosis, and altered estrus cycles were noted in the

dams; increased fetal mortality and decreased fetal weight were also

reported in offspring. No animal studies are available on 59

reproductive or developmental effects of oral or dermal exposure.

E.7 HYDROGEN FLUORIDE HEALTH EFFECTS SUMMARY

Hydrogen fluoride (HF) is a colorless gas that is used in making

aluminum and in making chlorofluorocarbons. HF readily dissolves in

water, is present in the air or other media, and, in the dissolved

form, is known as hydrofluoric acid. Air around hazardous waste sites

or factories that use or produce HF may contain this chemical. EPA 60

has set a maximum contaminant level (MCL) of 4 mg/L for HF.61

E.7.1 Cancer Effects — Hydrogen Fluoride

A cohort of workers in Denmark exposed to hydrofluoric fumes or

dust reported an increase in mortality and morbidity from respiratory

cancer. Increased lung cancer rates have been reported in aluminum

industry workers, although no correction was made for smoking and

exposure to other chemicals. Epidemiological studies of populations

exposed to fluorides through drinking water have not shown an

increased risk of cancer. No data are available on cancer in humans

following dermal exposure to HF. No animal studies have been 62

identified regarding the carcinogenic effects of HF. EPA has not

classified HF with respect to carcinogenicity and has not estimated a

unit risk for HF.63
E-13

E.7.2 Noncancer Effects — Hydrogen Fluoride

E.7.2.1 Acute (Short-Term) Effects for Hydrogen Fluoride. Acute

(short-term) inhalation exposure to HF can cause severe respiratory

damage in humans, including severe irritation and pulmonary edema. 

Many of the human studies regarding inhalation of HF also involved

dermal exposure, making it difficult to determine which effects are

specific to the inhalation route. The results of ingestion include

necrosis of the esophagus and stomach with nausea, vomiting, diarrhea,

circulatory collapse, and death. Severe ocular irritation and dermal

burns may occur following eye or skin exposure.64,65

ln animals, acute inhalation exposure has resulted in renal and

hepatic damage. HF produces irritation of the eyes, skin, and

conjunctivae in rats as a result of dermal exposure. No information

was found on the effects on animals from oral exposures to HF.66

E.7.2.2 Chronic (Long-Term) Effects for Hydrogen Fluoride. The

major health effect of chronic inhalation exposure to HF and fluoride

dusts, either individually or in combination, is skeletal fluorosis. 67

Chronic inhalation exposure of humans to HF has resulted in irritation

and congestion of the nose, throat, and bronchi at low levels. In 68

addition, persons exposed occupationally to HF and fluoride dusts in

an aluminum smelter reported reduced expiratory volume and increased

cough and sputum production. No information is available on the

chronic effects of oral or dermal exposure to HF in humans.69

Limited information exists on the chronic effects of HF in

animals. Damage to the liver, kidneys, and lungs has been observed in

animals chronically exposed to HF by inhalation. No information was 70

found on the long-term effects of oral or dermal exposure in animals. 

EPA is reviewing the RfC and RfD for HF.71

E.7.2.3 Reproductive and Developmental Effects for Hydrogen

Fluoride. No studies were located regarding the developmental and

reproductive effects in humans from inhalation, oral, or dermal

exposure to HF.72

Dogs exposed via inhalation to HF developed degenerative

testicular changes and ulceration of the scrotum. No studies were

found regarding the reproductive and developmental effects in animals

from oral or dermal exposure.73

E.8 MERCURY HEALTH EFFECTS SUMMARY

Mercury is a naturally occurring element that exists in three

forms: elemental mercury, inorganic mercury (primarily mercuric

chloride), and organic mercury (primarily methyl mercury). Elemental

mercury is a shiny, silver-white, odorless liquid; inorganic mercury

compounds are usually white powders or crystals; and organic mercury

compounds are white crystalline solids. The majority of mercury in

air is elemental mercury vapor, which is released to the air by

natural and industrial sources. The health effects of mercury and
E-14

mercury compounds are summarized in chapter 7 of this Utility HAP

Study Report and are discussed in greater length and detail in Volume

V of the Mercury Study Report to Congress.74

E.9 NICKEL HEALTH EFFECTS SUMMARY

Nickel is a silvery-white metal that is usually found in nature

as a component of silicate, sulfide, or arsenide ores. Table E-1

presents the physical properties of some of the major forms of nickel.

The most predominant forms of nickel in the atmosphere are

probably nickel sulfate, nickel oxides, and the complex oxides of

nickel. Each form of nickel exhibits different physical properties. 

Nickel compounds may be divided into two groups: soluble and

insoluble nickel compounds. The soluble compounds include nickel

sulfate and nickel acetate. Insoluble compounds include nickel

monoxide, metallic nickel, nickel hydroxide, nickel subsulfide, and

nickel carbonyl. Most nickel is used to make stainless steel; other

uses include the manufacture of batteries, electroplating baths,

textile dyes, coins, spark-plugs, and machinery parts.

E.9.1 Cancer Effects — Nickel

Human studies have reported an increased risk of lung and nasal

cancers among nickel refinery workers exposed to nickel refinery dust

and to nickel sulfate. Nickel refinery dust is defined as the "dust 75

from pyro-metallurgical sulfide nickel matte" refineries and is a

mixture of many nickel compounds, including nickel subsulfide. It is

not clear which compound is carcinogenic in the nickel refinery dust. 76

No information is available on the carcinogenic effects of nickel in

humans from oral or dermal exposure.77,78

Animal studies have reported lung tumors from inhalation exposure

to the following nickel compounds and mixtures: nickel refinery

dusts, nickel sulfate, nickel subsulfide, nickel carbonyl, and

metallic nickel. Studies in animals have reported tumors from

intramuscular and other routes of administration from exposure to

nickel monoxide and nickel hydroxide. Oral animal studies have not

reported tumors from exposure to nickel acetate in the drinking water. 

No information is available on the carcinogenic effects of nickel in

animals from dermal exposure79,80,81,82

E.9.1.1 Cancer Effects for Nickel Refinery Dust. U.S.

Environmental Protection Agency has classified nickel refinery dust in

Group A - Known Human Carcinogen. For nickel refinery dust, the Group

A classification was based on an increased risk of lung and nasal

cancer in humans through inhalation exposure and increased lung tumor

incidences in animals. The International Agency for Research on 83

Cancer (IARC) has classified nickel refinery dust as having sufficient

evidence in humans for carcinogenicity. This is based on the same

information U.S. Environmental Protection Agency used.
E-15

Table E-1. Physical Properties of Some Forms of Nickel

Chemical Name Formula Description Solubility

Metallic Nickel Ni Lustrous white, hard Soluble in dilute nitric acids; slightly

ferromagnetic metal or grey soluble in hydrochloric or sulfuric acids;

powder insoluble in cold or hot water

Nickel Hydroxide Ni (OH) Green crystals or Nearly insoluble in cold water; soluble in 2

amorphous solid acid, ammonium hydroxide

Nickel Subsulfide Ni S Lustrous pale yellow or Insoluble in cold water; soluble in nitric 3 2

bronze metallic crystals acid

Nickel Carbonyl Ni (O) Colorless to yellow liquid Nearly insoluble in water; soluble in 4

ethanol, benzene, and nitric acid;

insoluble in dilute acids or dilute alkali

Nickel Sulfate NiSO Pale-green to yellow crystals Soluble in water; insoluble in ethanol

(anhydrous) 4

Nickel Monoxide NiO Grey, black, or green Insoluble in water; soluble in acid

powder

Nickel Acetate Ni(OCOCH ) Dull-green crystals Soluble in water; insoluble in ethanol 3 2

Source: IARC 199035

U.S. Environmental Protection Agency used the additive and

multiplicative extrapolation method, based on human data, to estimate

the unit cancer risk for nickel refinery dust. U.S. Environmental

Protection Agency calculated an inhalation unit risk estimate of 

2.4 x 10 (g/m . Based upon this unit risk estimate, U.S. -4 3)-1 84

Environmental Protection Agency estimates that if an individual were

to breathe air containing nickel refinery dust at 0.004 μg/m over his 3

or her entire lifetime (70 yrs, 24 hrs/day), that person would

theoretically have an increased chance of up to one in one million of

developing cancer as a direct result of breathing air containing this

chemical. Similarly, U.S. Environmental Protection Agency estimates

that breathing air containing 0.04 μg/m would result in an increased 3

chance of an increased chance of up to one in one hundred thousand of

developing cancer.85

U.S. Environmental Protection Agency used four data sets, all

from human exposure, to calculate the unit risk estimates for nickel

refinery dusts. A range of incremental unit risk estimates was

calculated from these data sets that were consistent with each other. 86

E.9.1.2 Cancer Effects for Nickel Sulfate. The National

Toxicology Program (NTP) has recently completed a draft report on the

carcinogenic effects of nickel sulfate hexahydrate. They have

concluded that there was no evidence of carcinogenic activity of

nickel sulfate hexahydrate in male or female rats or male or female
E-16

mice. These conclusions are based on the results of 2-year inhalation

studies.87

The International Committee on Nickel Carcinogenesis in Man

summarized the available epidemiologic data on nickel and concluded

that there was strong evidence that exposure to soluble nickel

(primarily nickel sulfate) was associated with an increased

respiratory cancer risk.88

The International Agency for Research on Cancer (IARC) has

classified nickel sulfate as having sufficient evidence in humans for

carcinogenicity. This is based on epidemiological studies that 89

showed an increased risk of lung and nasal cancer through inhalation

exposure. In addition, animal studies have reported malignant tumors

in the peritoneal cavity when nickel sulfate was applied by

intraperitoneal injections.90

E.9.1.3 Cancer Effects for Nickel Subsulfide. U.S.

Environmental Protection Agency has also classified nickel subsulfide

in Group A, based upon the same studies as those that were used to

classify nickel refinery dust. For nickel subsulfide, U.S. 91

Environmental Protection Agency also used human data to estimate the

unit cancer risk. U.S. Environmental Protection Agency calculated an

inhalation unit risk estimate of 4.8 x 10 (g/m ) . U.S. -4 3 -1 92

Environmental Protection Agency estimates that if an individual were

to breathe air containing this nickel compound at 0.002 g/m over his 3

or her entire lifetime, that person would theoretically have an

increased chance of up to one in one million chance of developing

cancer as a direct result of breathing air containing this chemical. 

Similarly, U.S. Environmental Protection Agency estimates that

breathing air containing 0.02 g/m would result in an increased chance 3

of up to one in one hundred thousand chance of developing cancer. 

U.S. Environmental Protection Agency has also calculated unit risk

estimates for nickel subsulfide from a rat inhalation study. These

estimates were approximately one order of magnitude greater than those

calculated from the human studies.93

The National Toxicology Program has recently completed a draft

report on the carcinogenic effects of nickel subsulfide. They have

concluded that there was clear evidence of carcinogenic activity of

nickel subsulfide in male and female rats and no evidence of

carcinogenic activity for male and female mice. These conclusions are

based on the results of 2-year inhalation studies.94

IARC has classified nickel subsulfide as having sufficient

evidence in humans and experimental animals for carcinogenicity. The 95

International Committee on Nickel Carcinogenesis in Man concluded that

there was some evidence to suggest that exposure to nickel subsulfide

presents on increased risk of lung and nasal cancer.96

The State of California has calculated an estimated unit risk for

continuous lifetime exposure to nickel subsulfide at 1 g Ni/m . This 3

risk ranges from 2.8 x 10 for the maximum likelihood estimate to -3
E-17

3.7 x 10 for the upper 95 percent confidence limit. This risk -3

estimate was based on animal data.97

 

E.9.1.4 Cancer Effects for Nickel Carbonyl. U.S. Environmental

Protection Agency has classified nickel carbonyl in Group B2 -

Probable Human Carcinogen. For nickel carbonyl, this classification

was based on an increase in lung tumors in animals exposed via

inhalation. IARC has classified nickel carbonyl as having limited 98

evidence in experimental animals for carcinogenicity. This is based 99

on the same information as that U.S. Environmental Protection Agency

used.

U.S. Environmental Protection Agency has not calculated an

inhalation or an oral unit cancer risk estimate for nickel carbonyl,

due to the lack of appropriate data. In one study, the survival rate

of the animals was very low, and another study used the intravenous

route of exposure.100

E.9.1.5 Cancer Effects for Nickel Monoxide. The NTP has

recently completed a draft report on the carcinogenic effects of

nickel monoxide. They have concluded that there was some evidence of

carcinogenic activity of nickel monoxide in male and female rats, no

evidence of carcinogenic activity in male mice, and equivocal evidence

of carcinogenic activity in female mice. These conclusions are based

on the results of 2-year inhalation studies.101

IARC has classified nickel monoxide as having sufficient evidence

in experimental animals for carcinogenicity. This is based on animal 102

studies that showed an increased incidence of tumors in rats exposed

via intrapleural, intramuscular, and intraperitoneal administration. 

The International Committee on Nickel Carcinogenesis summarized the

available epidemiologic data on nickel and concluded that there was

some evidence to suggest that exposure to oxidic nickel (including

nickel monoxide) may result in increased lung and nasal cancer risks. 103

E.9.1.6 Cancer Effects for Nickel Hydroxide. IARC has

classified nickel hydroxide as having sufficient evidence in

experimental animals for carcinogenicity. This is based on animal 104

studies that showed an increase in tumors in rats exposed via

intramuscular injection.

E.9.1.7 Cancer Effects for Metallic Nickel. IARC has classified

metallic nickel as having sufficient evidence in experimental animals

for carcinogenicity. This is based on animal studies that showed an 105

increase in tumors from exposure via inhalation and intratracheal,

intraperitoneal, and intravenous administration. The International

Committee on Nickel Carcinogenesis in Man summarized the available

data on nickel and concluded that the available information gave no

evidence of increased respiratory cancer risks from exposure to

metallic nickel.106

E.9.1.8 Nickel Acetate. IARC has not classified nickel acetate

as to carcinogenicity.107
E-18

E.9.1.9 Overall Assessment for Nickel Compounds. IARC examined

all of the data on nickel and stated that for an overall evaluation,

it considers nickel compounds to be carcinogenic to humans and

metallic nickel to be possibly carcinogenic to humans.108

The State of California has calculated an estimated unit risk for

continuous lifetime exposure to nickel compounds at 1 g/m . This risk 3

ranges from 2.1 x 10 for the maximum likelihood estimate to 2.57 x 10 -4 -

 for the upper 95 percent confidence limit. This risk estimate was 4

based on human data. They also concluded that all nickel compounds

should be considered potentially carcinogenic to humans by

inhalation.109

The American Conference of Governmental Industrial Hygienists

(ACGIH) have stated that all nickel compounds should be considered as

confirmed human carcinogens, based on the weight of evidence from

epidemiologic studies of, or convincing clinical evidence in, exposed

humans.110

The International Committee on Nickel Carcinogenesis in Man

concluded that more than one form of nickel gives rise to lung and

nasal cancer. They stated that although much of the respiratory

cancer risk seen among nickel refinery workers could be attributed to

exposure to a mixture of nickel oxides and sulfides, exposure to large

concentrations of nickel oxides in the absence of nickel sulfides was

also associated with increased lung and nasal cancer risks. In

addition, there was evidence that soluble nickel exposure (such as

nickel sulfate) increased the risk of these cancers. They concluded

that respiratory cancer risks are primarily related to exposure to

soluble nickel at concentrations greater than 1 mg/m and to exposure 3

to less soluble forms at concentrations greater than 10 mg/m .3 111

E.9.2 Noncancer Effects — Nickel

E.9.2.1 Acute (Short-Term) Effects for Nickel. Nickel carbonyl

appears to be the most acutely toxic nickel compound. Symptoms from

acute inhalation exposure in humans include headache, vertigo, nausea,

vomiting, insomnia, and irritability, followed by chest pains, dry

coughing, cyanosis, gastrointestinal symptoms, sweating, visual

disturbances, and severe weakness. Acute oral exposure to high levels

of nickel sulfate and nickel chloride in humans has resulted in

vomiting, cramps, impaired vision, giddiness, headache, and cardiac

arrest in humans. No information is available on the acute effects of

nickel via dermal exposure in humans.112

The lungs and kidneys appear to be target organs for acute nickel

carbonyl toxicity, via inhalation and oral exposure in animals, with

pulmonary fibrosis and renal edema reported. No information is

available on acute effects of nickel via dermal exposure in animals. 113

Acute animal tests, such as the LD test in rats, have shown nickel 50

compounds to exhibit acute toxicity values ranging from low to high,

based upon LD data in the range of 50 mg/kg to greater than 5,000 50

mg/kg. The soluble compounds, such as nickel acetate, were most
E-19

toxic, and the insoluble compounds, such as metallic nickel powder,

were the least toxic. 114

E.9.2.2 Chronic (Long-Term) Effects for Nickel. Contact

dermatitis is the most common effect in humans from exposure to

nickel, via inhalation, oral, and dermal exposure. Cases of nickel

contact dermatitis have been reported following occupational and

nonoccupational exposure, with symptoms of itching of fingers, wrists,

and forearms. Chronic inhalation exposure to nickel in humans also

results in respiratory effects. These effects include direct

respiratory effects such as asthma due to primary irritation or an

allergic response and an increased risk of chronic respiratory tract

infections.115,116

Animal studies have reported effects on the lungs, kidneys, and

immune system from inhalation exposure to nickel, and effects on the

respiratory and gastrointestinal systems, heart, blood, liver, kidney,

and decreased body weight from oral exposure to nickel. Dermal animal

studies have reported effects on the skin.117,118

E.9.3 Essentiality for Nickel

Nickel has been demonstrated to be an essential nutrient for some

mammalian species, and it has been suggested that it may also be

essential for human nutrition. A requirement for nickel has not been

conclusively demonstrated in humans, and a recommended daily allowance

has not been set. By extrapolation from animal data, there have been

various estimates of the human daily requirement for nickel. The

National Academy of Sciences estimated that a 70 kilogram person would

have a daily requirement of 50 g of nickel. Other researchers have 119

estimated requirements ranging from 30 g to 120 g of nickel. 120

E.9.4 Reproductive and Developmental Effects for Nickel

No information is available regarding the reproductive or

developmental effects of nickel in humans. Animal studies have

reported developmental effects, such as a reduction in fetal body

weight, and reproductive effects, including testicular degeneration

from inhalation exposure to nickel. Oral animal studies have reported

deaths in females due to pregnancy complications and a significant

decrease in number of offspring per litter from exposure to nickel. 121

E.9.5 Noncancer Health-based Numbers for Nickel

U.S. Environmental Protection Agency has established a Reference

Dose (RfD) for nickel (soluble salts) of 0.02 mg/kg/day, based upon a

NOAEL (adjusted) of 5 mg/kg/day, an uncertainty factor of 300, and a

modifying factor of 1. This was based on a study in rats that showed 122

decreased body and organ weights from chronic (2-year) exposure to

nickel in the diet. Other studies showed similar results, with

decreased body and organ weights after exposure to nickel chloride via

gavage and through the drinking water. U.S. Environmental Protection

Agency has not established a Reference Concentration (RfC) for any

nickel compound.123
E-20

U.S. Environmental Protection Agency has medium confidence in the

RfD for nickel (soluble salts) and low confidence in the study on

which it was based. The Ambrose et al. 1976 study was properly

designed and provided adequate toxicological endpoints; however, high

mortality occurred in the controls. The database provided adequate 124

supporting subchronic studies; this was the basis for U.S.

Environmental Protection Agency’s medium confidence level in the RfD. 125

The EPRI has recommended a RfC of 2.38 x 10 mg(Ni)/m for all -3 3

nickel compounds. This was based on the ACGIH Threshold Limit Value

(TLV). It was translated for community exposure by scaling for

exposure time differences between community and occupational exposure

assumptions.126

Calabrese has calculated an ambient air level goal (AALG) for

soluble nickel compounds of 0.36 ng (Ni)/m and an AALG for insoluble 3

nickel compounds of 7.1 ng (Ni)/m . An AALG is a health-based 3

guideline based on risk assessment methodology similar to that used by

U.S. Environmental Protection Agency.127

The California Air Resources Board has stated that the most

sensitive noncancer endpoint reported in humans is allergic

sensitization, while immune suppression is the most sensitive endpoint

reported in animal studies. The board has concluded that because

these noncancer effects occur at concentrations greater than 3 orders

of magnitude above a 24-hour maximum concentration of nickel (0.024

ng(Ni)/m ) measured in California near an industrial source, it is 3

unlikely that noncancer health effects would be caused by the levels

of nickel compounds currently in the air.128

The Agency of Toxic Substances and Disease Registry has

recommended a minimum risk level (MRL) for intermediate duration,

inhalation exposure to nickel of 9.5 x 10 mg(Ni)/m . They have stated -5 3

that this MRL may not be protective for some hypersensitive

individuals. An MRL is a health-based guideline based on similar 129

risk assessment methodology to that used by U.S. Environmental

Protection Agency.

E.9.6 Federal Regulations and Guidelines for Nickel

The Occupational Safety and Health Administration (OSHA) has

established a maximum allowable level of nickel in workplace air for

an 8-hour workday, 40-hour workweek of 1 mg(Ni)/m for metallic nickel 3

and insoluble compounds, and 0.1 mg(Ni)/m for soluble nickel 3

compounds.130

The National Institute of Occupational Safety and Health has a

recommended exposure level for workplace air of 0.15 mg (Ni)/m for all 3

nickel compounds except nickel carbonyl and 7 g (Ni)/m for nickel 3

carbonyl.131

The ACGIH has recommended a TLV of 0.05 mg(Ni)/m for an 3

8-hour exposure in the workplace to all nickel compounds (elemental,

insoluble, and soluble).132
E-21

The U.S. Environmental Protection Agency has set a maximum

contaminant level (MCL) of 0.1 mg/L for nickel. This is the maximum

level allowed in drinking water.133

E.10 2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN HEALTH EFFECTS SUMMARY

2,3,7,8-Tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) belongs to the

class of compounds, chlorinated dibenzo-p-dioxins, which are referred

to as dioxins. 2,3,7,8-TCDD is a colorless solid with no known odor. 

It does not occur naturally, nor is it intentionally manufactured by

any industry, although it can be produced inadvertently in small

amounts as an impurity during the manufacture of certain herbicides

and germicides and has been detected in products of incineration of

municipal and industrial wastes. The only present use for

2,3,7,8-TCDD is in chemical research.134

E.10.1 Cancer Effects — Dioxins

An increase in lung cancer risks was observed among Japanese

males exposed as a result of an oil poisoning accident. Human studies

have also found an association between 2,3,7,8-TCDD and soft-tissue

sarcomas, lymphomas, and stomach carcinomas, although for malignant

lymphomas, the increase in risk is not consistent. The increase in

risk is of borderline significance for highly exposed groups and is

less-significant among groups exposed to lower levels of 2,3,7,8-TCDD.

Although there are problems with the studies of human effects, such as

confounding factors, short follow-up period, and lack of exposure

information, the overall weight of evidence from epidemiological

studies suggests that the generally increased risk of cancer in humans

is likely due to 2,3,7,8-TCDD.135

Information on the carcinogenicity of 2,3,7,8-TCDD following

inhalation exposure of animals is not available. In animal studies of

oral exposure to 2,3,7,8-TCDD, multisite tumors in rats and mice

including the tongue, lung, nasal turbinates, liver, and thyroid have

been reported. Estimates derived from human data suggest a unit

risk for lung cancer of 3 x 10 to 5 x 10 pg/kg-day) ; for all -4 -4 -1

cancers combined the unit risk estimate is 2 x 10 to 3 x 10 -3 -3

(pg/kg-day) (U.S. Environmental Protection Agency ). -1 136

E.10.2 Noncancer Effects — Dioxins

E.10.2.1 Acute (Short-Term) Effects for Dioxins. The acute

effects on humans exposed through the spraying in Vietnam of

herbicides that contained 2,3,7,8-TCDD include diarrhea, vomiting,

skin rashes, fever, and abdominal pain. Routes of exposure in these 137

instances are not well defined and may include inhalation as well as

oral and dermal exposures.

No information is available on effects in animals from acute

inhalation exposure to 2,3,7,8-TCDD. In oral exposure studies,

2,3,7,8-TCDD is highly toxic to all laboratory animals tested even

though there are large differences in species sensitivity. LD values 50

range from 0.6 g/kg in male guinea pigs to 5,500 g/kg in hamsters. 
E-22

Other effects on animals from acute oral exposure include loss of body

weight, hepatotoxicity, and decreased thymus weight. Information on 138

the effects of acute dermal exposure in animals is limited, although

dermal effects have been reported.139

E.10.2.2 Chronic (Long Term) Effects for Dioxins. No studies

are available on the inhalation toxicity of 2,3,7,8-TCDD in humans,

although such exposure may have occurred in populations exposed to

chemicals contaminated with 2,3,7,8-TCDD. Oral exposure of humans to

chemicals contaminated with 2,3,7,8-TCDD has resulted in chloracne,

immunotoxicity, hyperpigmentation, hyperkeratosis, possible

hepatotoxicity, aching muscles, loss of appetite, weight loss,

digestive disorders, headaches, neuropathy, insomnia, sensory changes,

and loss of libido.140

Chloracne is the only substantiated effect in humans produced by

dermal exposure to compounds contaminated with 2,3,7,8-TCDD.141

No information on chronic inhalation and dermal exposure is

available for animals. Oral exposure to 2,3,7,8-TCDD has resulted in

dermatitis, extreme loss of body weight, and effects on the liver and

immune system. U.S. Environmental Protection Agency has not 142

established an RfC or RfD for 2,3,7,8-TCDD.

E.10.2.3 Reproductive and Developmental Effects for Dioxins. 

Several studies have investigated the incidence of birth defects and

reproductive effects in humans exposed to 2,3,7,8-TCDD through

accidental releases or the spraying of 2,3,7,8-TCDD-contaminated

herbicides. U.S. Environmental Protection Agency has concluded that

the data were not inconsistent with 2,3,7,8-TCDD’s adversely affecting

development, but as a result of the limitations of the data, these

studies could not prove an association with 2,3,7,8-TCDD exposure and

the observed effect. The major limitations in these human studies

were the concomitant exposure to other potentially toxic chemicals,

the lack of any specific quantitative data on the extent of exposure

of individuals within the study group, and the lack of statistical

power of the studies.143

No studies are available on the reproductive and developmental

effects in animals caused by inhalation or dermal exposure to

2,3,7,8-TCDD. In oral exposure studies, 2,3,7,8-TCDD has produced 144

fetal anomalies, including cleft palate and hydronephrotic kidneys in

mice and internal organ hemorrhage in rats, and resulted in

spontaneous abortions in monkeys and decreased fetal survival.
E-23

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E-24

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E-25

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E-26

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98. Ref. 81, Carcin-1.

99. Ref. 35, p. 410.

100. Ref. 81, Carcin-2.

101. National Toxicology Program. NTP Draft Technical Report on the

Toxicology and Carcinogenesis Studies of Nickel Oxide in F344/N

Rats and B6C3Fl Mice. U.S. Dept. of Health and Human Services. 

Public Health Service. National Institute of Health. 1994. 
E-28

102. Ref. 35, p. 410.

103. Ref. 88, p. 72.

104. Ref. 35, p. 410.

105. Ref. 35, p. 410.

106. Ref. 88, p. 72-73.

107. Ref. 35, p. 410-411.

108. Ref. 35, p. 411.

109. Ref. 97, p. 4-5 Part B.

110. American Conference of Governmental Industrial Hygienists. 

Threshold Limit Values for Chemical Substances and Physical

Agents and Biological Exposure Indices. Cincinnati, OH. 1995.

111. Ref. 88, p. 74-75.

112. U.S. Environmental Protection Agency. Health Assessment Document

for Nickel. Office of Health and Environmental Assessment. U.S.

Environmental Protection Agency/600/8-83/012F. Section 5.5.1. 

1985.

113. Ref. 112, Section 5.1.2.

114. Ref. 78, Section 2.2.2.

115. Ref. 78, Section 2.2.

116. Ref. 112, Section 5.2.2.

117. Ref. 78, Section 2.2.

118. Ref. 112, Section 5.2.1.3.

119. National Academy of Sciences. Drinking Water and Health. 

Volume 3. Safe Drinking Water Committee. National Academy

Press. Washington, D.C. 1980.

120. Ref. 97, p. 26-31.

121. Ref. 78, Section 2.2.

122. U.S. Environmental Protection Agency. Integrated Risk

Information System on Nickel, Soluble salts. RfD-1. 1995.

123. Ref. 122, RfC-1.
E-29

124. Ambrose et. Al. Long-term toxicologic assessment of nickel in

rats and dogs. J. Food Sci. Technol. 13:181-187. 1976.

125. Ref. 122, RfD-6.

126. Ref. 8, p. I-3.

127. Calabrese, E. Air Toxics and Risk Assessment. Lewis Publishers,

Inc. Chesea, MI. p. 463-471. 1991.

128. Ref. 97, Executive Summary.

129. Ref. 78, Section 2.2.

130. Ref. 78, Section 7.

131. Ref. 78, Section 7.

132. Ref. 78, Section 7.

133. Ref. 3.

134. ATSDR. Toxicological Profile for 2,3,7,8-Tetrachlorodibenzo-p-dioxin. Agency for Toxic Substances and Disease

Registry. U.S. Public Health Service. U.S. Department of Health

and Human Services. Section 1. 1989.

135. U.S. Environmental Protection Agency. Health Assessment Document

for 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related

Compounds. Vol II. (Draft). Office of Research and Development. 

Washington, DC. Volume II, Chapter 7. 1994.

136. Ref. 135, Volume I, Chapter 6; Volume III.

137. HSDB. Hazardous Substances Databank (online database). U.S.

Department of Health and Human Services, National Library of

Medicine, National Toxicology Information Program. Bethesda, MD. 

1993.

138. Ref. 134, Section 2.2.

139. Ref. 135, Volume I, Chapter 6.

140. Ref. 134, Section 2.2.

141. Ref. 134, Section 2.2.3.

142. Ref. 134, Section 2.2.2.

143. Ref. 134, Section 2.2.1.

144. Ref. 134, Section 2.2.2.

Appendix F — Documentation of The Inhalation Human Exposure Modeling

for the Utility Study
x
F-1

F.1 INTRODUCTION

The model used to calculate direct inhalation risks from

hazardous air pollutants (HAPs) emitted from utility boilers is the

Human Exposure Model Version 1.5 (HEM 1.5). It was developed by the

Pollutant Assessment Branch (PAB) of the U.S. Environmental Protection

Agency’s (EPA) Office of Air Quality Planning and Standards (OAQPS)

and was designed for screening assessments. The model is used in

source ranking to assess the relative risks associated with exposure

to different pollutants and to characterize human exposure, cancer

risks, and noncarcinogenic hazards for stationary sources that emit

HAPs. The HEM uses the Industrial Source Complex - Long Term Version

2 (ISCLT2) air dispersion model, updated 1990 census population data,

meteorological, temperature, and mixing height databases, and

chemical-specific health effects numbers (see Table F-1.)

The remainder of this technical report contains a description of

ISCLT2, the population and meteorological databases, human exposure

algorithms, and risk estimating methodology applied in HEM 1.5 to

arrive at direct inhalation risk estimates for this utility study.

F.2 ISCLT2 DISPERSION MODELING

Air dispersion modeling is used to estimate atmospheric fate and

transport of pollutants from the point of emission to the location of

exposure to arrive at long-term average ambient air concentrations of

the pollutant. ISCLT2, the air dispersion model used in HEM 1.5, is

the Agency’s regulatory air dispersion model for the types of sources

represented in this study. ISCLT2 is one of the primary models used

to support EPA studies and regulatory programs for air pollutants. 

ISCLT2 uses emission parameters and meteorological data to estimate

the transport and dispersion of pollutants in the atmosphere.

The ISCLT is a steady-state, Gaussian plume, atmospheric

dispersion model that applies to multiple-point, area, and volume

emission sources. It is designed specifically to estimate long-term

ambient concentrations resulting from air emissions from these source

types in a computationally efficient manner. ISCLT2 is recognized by

the Guideline on Air Quality Models as a preferred model for dealing 1

with complicated sources (i.e., facilities with point, area, and

volume sources) when estimating long-term concentrations (i.e.,

monthly or longer).

As described in the Guideline on Air Quality Models, the ISCLT is

appropriate for modeling industrial source complexes in either rural

or urban areas. With this model, long-term ambient concentrations can 1

be estimated for transport distances up to 50 km. The ISCLT2

incorporates separate point, area, and volume source computational

algorithms for calculating ambient concentrations at user-specified

locations (i.e., receptors). The locations of the receptors relative

to the source locations are determined through a user-specified

Cartesian coordinate reference system. For the utility study,
F-2

Table F-1. Summary of HEM 1.5 Features

Characterization Single pollutant, multiple source, nationwide

Dispersion model ISCLT2

Meteorological database Data set from locations/years available on OAQPS TTN and from the

National Weather Service

Population database 1990 Census Databases 

Block level

6.9 million records

Exposure calculations >0.5 km Interpolate air concentration to population

<0.5 km Assign population to air concentration

HEM = Human Exposure Model

ISCLT2 = Industrial Source Complex Model - Long Term Version 2

OAQPS = Office of Air Quality Planning and Standards

TTN = Technology Transfer Network

receptors were placed around the source along 16 radials, spaced every

22.5 degrees, at distances of 0.2, 0.5, 1.0, 2.0, 5.0, 10.0, 20.0,

30.0, 40.0 and 50.0 kilometers from the source.

ISCLT2 source inputs vary according to source type. For the

point sources in this study, the inputs include emission rate,

physical stack height, stack inner diameter, stack gas exit velocity,

and stack gas exit temperature.

The ISCLT2 is a sector-averaged model that uses statistical

summaries of meteorological data to calculate long-term, ground-level

ambient concentrations. The principal meteorological inputs to the

ISCLT2 are STability ARray (STAR) summaries that consist of a

tabulation of the joint frequency of occurrence of wind speed

categories, wind-direction sectors, and Pasquill atmospheric stability

categories. Other meteorological data requirements include average

mixing heights for each stability class and average ambient air

temperatures.

As described above, the ISCLT2 model computes long-term ambient

concentrations at user-specified receptor points that occur as a

result of air emissions from multiple sources. These computations are

done on an emission point (stack)-by-stack basis, such that the

ambient concentration from each stack at each receptor is computed. 

Total ambient concentrations at a particular receptor are obtained by

summing the contributions from each of the stacks. With Gaussian

plume algorithms such as those included in the ISCLT2, the source

contributions at each receptor are directly proportional to the source

emission rate.
Xi  

J

j1

qixij

F-3

Normalized ambient concentrations for each source-receptor

combination were computed such that they would correspond to a unit

emission rate of 1 gram per second (g/s) for each stack in the

facility. The total ambient concentration at a receptor is then

computed as the sum of the contributions from each stack, where the

latter are computed as the product of the normalized concentration and

the desired emission rate. Mathematically, this can be expressed as

follows:

Where:

X = total ambient concentration at receptor i, μg/m i 3

q = emission rate for stack, g/s i

X = normalized contribution from stack j to receptor i, μg/m ij 3

J = total number of stacks.

Thus, the principal output of the dispersion modeling is a set of

normalized stack contributions (i.e., x in the above equation) for ij

each scenario modeled.

F.2.1 Assumptions Used

For the utility study, HEM analysis flat terrain was assumed

because of the lack of information. Building downwash was not

considered because of the tall stacks used by the utility boilers. 

The assumption was made that all particles were small enough to behave

as gases. All emissions from one site are assumed to originate from

stacks that are collocated.

F.2.2 Model Options

Air dispersion is affected by surface roughness. The ISCLT2

model provides two regimes of surface roughness based on land

classification: urban and rural. When there is no information

available regarding the land classification around a particular source

of interest, the air quality modeling guidelines suggest a surrogate,

population density, to make a land classification determination. 

Because the population database which is part of the HEM 1.5 model can

easily provide population density estimates, this option was selected

for the utility study for conducting the more detailed analyses. 

Initial screening analysis assumed the plant setting of “urban,” which

earlier sensitivity analysis indicated would maximize surrounding

ambient concentrations estimates.

EPA’s Guideline on Air Quality Models distinguishes between 2

urban and rural settings based on population density. “Urban” is

defined as a population density greater than 750 people per km in the 2
F-4

area between the point source and a 3 km radius from the source;

“rural” is assumed for a population density of less than 750 people

per km .2

ISCLT2 can be run in a number of different ways by changing

various modeling options. For consistency in regulatory modeling

applications, a set of choices has been defined as the default option. 

The default option set determines how the model calculates ambient air

concentrations and includes:

• default stack-tip downwash calculations

• buoyancy-induced dispersion calculations

• final plume rise in all calculations

• calms processing routines

• upper-bound concentration estimates for sources influenced

by building downwash from super-squat buildings

• default wind profile exponents

• default vertical potential temperature gradients.

The default option set was used in the utility study with one

change. Instead of the final plume rise option of the default

selections, a transitional plume rise was used. Plume rise accounts

for how the plume behaves near the stack as a function of the momentum

of release of the plume and the buoyant rising of the plume resulting

from the high plume temperature in comparison to the surrounding air. 

The use of the transitional plume rise would be expected to produce

more realistic estimates of ambient air concentrations near the stack

where the maximum concentrations occur. Each of these defaults is

defined further in the ISCLT2 User’s Guide.3

F.3 HEM DATABASES

Four databases are contained in the HEM 1.5 model. The

meteorological database contains long-term summaries for selected

locations across the country. HEM pairs plant locations with the

nearest location for meteorological data contained in the database. 

The second database is the population database, which contains

population data from the 1990 census. Ambient air concentrations of

the modeled pollutant are coupled with the population numbers and

location to develop nationwide exposure estimates. The two remaining

databases contain estimates of ambient temperatures and mixing height.

F.3.1 Meteorological Database

The ISCLT2 meteorological database contains long-term

meteorologic data, primarily from National Weather Service (NWS)

airport locations, in the form of STAR summaries. STAR summaries
F-5

display joint frequencies of occurrence of wind direction, wind speed,

and air stability by combining these factors into a frequency

distribution. HEM 1.5 chooses the STAR data set for each plant based

on proximity of the plant to the location where the meteorological

data were collected.4

The meteorological database used for the utility study contains

data from hourly surface observations obtained from the OAQPS

Technology Transfer Network (TTN). The Support Center for Regulatory

Air Models Bulletin Board System (SCRAM-BBS) contains annual data

files of surface observations from 349 NWS locations (primarily

airports) across the United States and its Territories for the years

1984-1989. From each location’s surface observations, STAR summaries

were created that encompass all available years into one long-term

estimate of the location’s dispersion characteristics. Figure F-1

depicts the coverage of the HEM 1.5 meteorological database. The

range of averaging years over which the data are averaged is from 1 to

6 years, with a typical average of 6 years (225 sites).

F.3.2 Population Database

The population database contains “block level” 1990 census data

collected by the U.S. Census Bureau for reapportionment as specified

in Public Law 94-17. It is used by the model to estimate the location

and number of people exposed to the modeled pollutants. The 1990

population has been aggregated into 6.9 million blocks.

F.3.3 Mixing Height Database

The mixing height database is more limited in scope than the

other databases mentioned above. Only 73 sites were available from

the NWS for the years 1984-1989. Also, the mixing heights are

calculated from observations taken once daily. Of the 73 sites, 40

are based on 6 years of observations.

F.3.4 Temperature Database

The temperature database provides an arithmetic average of

ambient temperatures for each atmospheric stability class for each

STAR site. Because the temperature was recorded for every set of wind

speed and direction observations in the NWS raw data, the temperature

database is similar to the meteorological database; that is, each

database has the same number of sites (349), the same number of years

of data to calculate the averages at each site, and the same typical

number of years (6) on which averages are based. By default, the site

closest to the plant is selected for air dispersion calculations and

is, for this database, the nearest STAR site.

F.4 EXPOSURE ALGORITHMS

Exposure is calculated in HEM 1.5 through pairing population

information from the census database with modeled ambient air

concentrations of each specific pollutant. The output of the

dispersion model is an air concentration array around the plant. HEM

1.5 calculates exposure by integrating the HAP air concentration at 

F-7

the population center (centroid) of the census block through

interpolation of the air concentration values at the surrounding

modeled points. All persons residing in the census block are treated

as being exposed to the air concentration at the centroid.

F.4.1 Air Concentration - Population Pairing

ISCLT2 calculates air concentrations at user-specified receptors. 

For the utility study, receptors were placed around the source along

16 radials, spaced every 22.5 degrees, at distances of 0.2, 0.5, 1.0,

2.0, 5.0, 10.0, 20.0, 30.0, 40.0, and 50.0 km from the source, for a

total of 160 receptors. Except for receptors located very close to

the stack, HEM 1.5 calculates exposure by interpolating the air

concentration at the population centroid (the population center of the

census block) between the values at the receptors surrounding the

centroid. There is a linear relationship between the logarithm of the

concentrations and the logarithm of the radial distances. This linear

relationship is used to estimate the concentration along the radial

nearest the centroid at the same distance from the stack as the

centroid. The estimates are then interpolated linearly between the

radials of the receptors surrounding the population centroid. Figure

F-2 depicts the relationship between the receptor locations and a

hypothetical block population centroid.

F.4.2 Exceptions for Population Close to Source

Within 0.5 km of the stack, the exposure is calculated

differently than described above because close to the stack, the

receptors are much closer together. Here, the population is estimated

at the points where the air concentration is calculated, rather than

the air concentrations’ being estimated at the known population point. 

This more complicated scheme is described in detail in the HEM user’s

manual.5

F.5 RISK CALCULATIONS

In general, long-term exposure estimates are paired with

chemical-specific health benchmarks, such as inhalation unit risk

estimates (IUREs), to calculate the risk to the population of

developing cancer or the potential for developing other adverse health

effects. Health benchmarks are input for each chemical modeled. 

Health benchmarks and other toxicity information are discussed in

Appendix E (Health Effects Summaries: Overview). Risk is calculated

for the exposed population on a single-pollutant basis. For

carcinogens, HEM 1.5 produces distributions of exposure and risk, as

well as estimates of annual incidence, number of people exposed at

various risk levels, and maximum individual risk (MIR). A comparison

of the modeled ambient air concentration to the reference

concentration is used to estimate the extent of adverse health effects

for noncarcinogens. Aggregate risk associated with exposure to

multiple pollutants is evaluated by adding the risks from individual

pollutants.
5 10 20 30 40 50 km

North

South

Block Centroid

Census Block

1

2

3

4

F-8

Figure F-2. The exposure algorithms interpolate between the estimated air concentrations and

the population data. Air concentrations are calculated at the points where the

circles and lines intersect. Population is known at the block centroid locations. 

The concentration at the centroid is calculated based on the concentration

estimated at the 4 points surrounding the centroid.
F-9

The utility boiler HEM modeling application requires the input of

chemical-specific toxicity information. HEM 1.5 uses the IUREs for

carcinogens to estimate cancer risks or other adverse health effects

for each individual chemical according to that chemical’s particular

level of toxicity. The more toxic a chemical, the lower the ambient

air concentration necessary to produce high risk levels.

F.5.1 Required Health Number Input

An IURE is entered in the risk calculation for each carcinogenic

pollutant. The IURE represents an estimate of the increased cancer

risk from a lifetime (70-year) exposure to a concentration of one unit

of exposure. The IURE for inhalation is normally expressed as risk

per μg/m of air contaminant. 3

Hazard quotients for noncarcinogens are calculated by comparing

the ambient air concentration of the pollutant with its reference

concentration (RfC). The RfC is an estimate (with uncertainty

spanning perhaps an order of magnitude) of the daily exposure of the

human population to the chemical by inhalation (including sensitive

subpopulations) that is likely to be without deleterious effects

during a lifetime.

F.5.2 Risk Calculations

HEM 1.5 calculates carcinogenic risk using standard EPA risk

equations and assumptions. Maximum individual risk (MIR) is defined

as the increased probability of an individual to develop cancer

following exposure to a pollutant at the maximum modeled long-term

ambient concentration assuming a lifetime of exposure. It is

calculated by multiplying the estimated ambient air concentration of a

HAP by the IURE. 

Unlike cancer risk characterization, noncancer risks typically

are not expressed as a probability of an individual suffering an

adverse effect. Instead, the estimated exposure concentration is

compared with a noncancer health benchmark such as an RfC. This is

usually expressed as a hazard quotient. The hazard quotient is the

ratio of the exposure (ambient air concentration of the pollutant) to

the RfC. The RfC represents 

the highest protective concentration, and a ratio value greater than

or equal to one would represent an exposure that may be a public

health concern and should be evaluated further.

For additional information on the carcinogenic and

noncarcinogenic effects of HAPs, refer to Appendix E (Health Effects

Summaries: Overview).

F.6 ASSUMPTIONS

Simplifying assumptions are used in the HEM utility boiler

analysis to enable estimation of the potential health effects due to
F-10

HAP emissions from utility boilers. The following assumptions are

made from HEM 1.5:

1. Direct inhalation of pollutants is the only source of

exposure.

2. Average exposures are equivalent to those experienced if one

constantly stayed at home; no adjustment is made for

exposure changes resulting from population movement between

home, school, work, etc.

3. Homes are located at population-weighted centers (centroids)

of census blocks (or at nodes of the polar grid within 0.5

km) because the locations of actual residences are not

included in the database.

4. For the most exposed individuals, it is assumed that people

reside at the home for their entire lifetimes (in modeling

carcinogens, a lifetime is assumed to be 70 years).

5. Indoor concentrations are the same as outdoor

concentrations.

6. The plant emits pollutants at the same level for the 70-year

lifetime of exposure.

7. No resuspension of pollutants via dust occurs.

8. There is no population migration or growth.

9. Varying exposures that might arise as a result of

differences in age, sex, health status, degree of activity,

etc. do not exist.

10. Because the model does not handle complex terrain, each

plant is located in flat terrain. An additional complex

terrain analysis was conducted using specially-designed models.

11. The nearest meteorological location provides the most

appropriate STAR, temperature, and mixing height data for

the plant.

12. No pollutants are emitted from point sources other than stacks.

F.7 HEM 1.5 OUTPUT

For carcinogens, HEM 1.5 produces estimates of annual incidence

(population risk), number of people exposed to various risk levels,

and maximum individual lifetime risk. For noncarcinogens, HEM 1.5

estimates the number of people exposed at various concentrations and

the maximum individual concentration. These values are for individual

pollutants; no summing of risks across chemicals is performed.
F-11

1. U.S. Environmental Protection Agency. Guideline on Air Quality

Models (Revised). EPA-450/2-78-027R. Office of Air Quality

Planning and Standards, Research Triangle Park, NC. 1993,

revised 1995. pp. 4-4.

2. Ref. 1, pp. 8-10.

3. U.S. Environmental Protection Agency. User’s Guide for the

Industrial Source Complex (ISC2) Dispersion Models, Volume 1 -

User’s Instructions. EPA-450/4-92-008a. Office of Air Quality

Planning and Standards, Research Triangle Park, NC. 1992.

4. Ref. 3, pp. 3-58 to 3-60

5. U.S. Environmental Protection Agency. User’s Manual for the

Human Exposure Model (HEM). EPA-450/5-86-001. Office of Air

Quality Planning and Standards, Research Triangle Park, NC. 

1986. pp. 2-12 to 2-19.

F.8 REFERENCES
mm
Appendix G — Data Tables for Dioxin Multipathway Assessment
xx
G-1

Appendix G-1

IEM Model Input Parameters

For Dioxin Screening Level

Multipathway Assessment

(Chapter 11)
xx
G-3

Table G-1. Scenario Dependent Exposure/Risk Modeling Parameters

Parameter fisher farmer resident resident Source 

Subsistence Subsistence Adult Child

Exposure duration, years 30 30 30 18 EPA, 1996cDistance from source, meters 500 500 5000 5000 EPA, 1994a; EPA, 1997bAir temperature, celsius 9.14 9.14 13.1 13.1 SAMSON, 1993Runoff, cm/y 50 50 25 25 Gerhaghty, 1973Wind speed, m/s 4.38 4.38 4.19 4.19 SAMSON, 1993Reduction in meat conc. due to fattening 1.0 N/A N/A N/A EPA, 1994aprior to slaughterBody weight, kg 70 70 70 17 EPA, 1996cInhalation rate, m /day 20 20 20 16 EPA, 1996c 3

Plant consumption, g/kg body weight/day grains 1.87 1.87 1.87 3.77 EPA, 1996c legumes 0.381 0.381 0.381 0.666 EPA, 1996c potatoes 0.170 0.170 0.170 0.274 EPA, 1996c root vegetables 0.024 0.024 0.024 0.036 EPA, 1996c fruits 0.57 0.57 0.57 0.223 EPA, 1996c fruiting vegetables 0.064 0.064 0.064 0.120 EPA, 1996c leafy vegetables 0.0281 0.0281 0.0281 0.008 EPA, 1996cAnimal consumption, g/kg body weight/d beef N/A 0.341 N/A N/A EPA, 1996c dairy N/A 0.599 N/A N/A EPA, 1996c poultry N/A 0.111 N/A N/A EPA, 1996c eggs N/A 0.073 N/A N/A EPA, 1996cSoil ingestion, g/d 0.1 0.1 0.1 0.2 EPA, 1996c
Table G-1. (Continued)

Parameter fisher farmer resident resident Source 

Subsistence Subsistence Adult Child

G-4

Water consumption, L/kg body weight/d 0.0285 0.0285 0.0285 0.02 EPA, 1996c

Fish ingestion, kg/kg body weight/d 0.000857 N/A N/A N/A EPA, 1996cDermal soil contact time, hr/day 1.5 8.33 1.5 5.6 EPA, 1996aDermal area exposed to soil, cm 5000 5000 5000 850 EPA, 1996a; 2

EPA, 1990

Dermal soil cont. amount, g/ cm 0.02 0.02 0.02 1.0 EPA, 1990 2

Fraction contaminated contact grains 0.667 1.0* 0.195 0.195 EPA, 1996c legumes 0.8 1.0* 0.5 0.5 EPA, 1996c potatoes 0.225 1.0* 0.031 0.031 EPA, 1996c root vegetables 0.268 1.0* 0.073 0.073 EPA, 1996c fruits 0.233 1.0* 0.076 0.076 EPA, 1996c fruiting vegetables 0.623 1.0* 0.317 0.317 EPA, 1996c leafy vegetables N/A 1.0* N/A N/A EPA, 1996c beef N/A 1.0* N/A N/A EPA, 1996c dairy N/A 1.0* N/A N/A EPA, 1996c eggs N/A 1.0* N/A N/A EPA, 1996c poultry N/A 1.0* N/A N/A EPA, 1996c

*Indicates a value not taken from the stated source, rather assumed appropriate for the scenario.
G-5

Table G-1a. Scenario Independent, Dioxin Congener Independent Parameters

Input parameter Input value Source of curr

Ideal gas constant, L-atm/mole-K 8.21e-02 EPA, 1996dAir density, g/cm3 1.19e-03 EPA, 1994aSolids density, kg/L, or g/cm3 2.65e+00 EPA, 1993Drag coefficient 1.10e-03 EPA, 1993Von Karman's coefficient 7.40e-01 EPA, 1993Boundry thickness 4.00e+00 EPA, 1993Interception fraction, unitless grains 0.00e+00 EPA, 1996d legumes 8.00e-03 EPA, 1996d potatoes 0.00e+00 EPA, 1996d root vegetables 0.00e+00 EPA, 1996d fruits 5.00e-02 EPA, 1996d

 fruiting vegetables 5.00e-02 EPA, 1996d leafy vegetables 1.50e-01 EPA, 1996d forage 4.70e-01 EPA, 1996d silage 4.40e-01 EPA, 1996dPlant surface loss coefficient, yr-1 1.80e+01 EPA, 1993Fraction wet deposition adhering 6.40e-01 EPA, 1994aLength of plants' exposure, yrs grains 0.00e+00 EPA, 1996d legumes 1.23e-01 EPA, 1996d potatoes 0.00e+00 EPA, 1996d root vegetables 0.00e+00 EPA, 1996d fruits 1.23e-01 EPA, 1996d fruiting vegetables 1.23e-01 EPA, 1996d leafy vegetables 1.57e-01 EPA, 1996d forage 1.23e-01 EPA, 1996d silage 1.23e-01 EPA, 1996dYield of plants, kg dw/m2
Table G-1a. (Continued)

Input parameter Input value Source of curr

G-6

 grains 3.00e-01 EPA, 1996d

 legumes 1.04e-01 EPA, 1996d potatoes 4.80e-01 EPA, 1996d root vegetables 3.34e-01 EPA, 1996d fruits 1.07e-01 EPA, 1996d fruiting vegetables 1.07e-01 EPA, 1996d leafy vegetables 1.77e-01 EPA, 1996d forage 3.10e-01 EPA, 1996d silage 8.40e-01 EPA, 1996d Surface Area Volume to Whole Plant Volume Correction grains 5.00e-01 EPA, 1994a legumes 1.00e-02 EPA, 1994a 2.50e-01 EPA, 1994a root vegetables 2.50e-01 EPA, 1994a fruits 1.00e-02 EPA, 1994a fruiting vegetables 1.00e-02 EPA, 1994a leafy vegetables 1.00e-02 EPA, 1994a forage 5.00e-01 EPA, 1994a silage 5.00e-01 EPA, 1994aFeed (grain & silage) and grass (forage) diet fractions, unitless beef-grain & silage 4.80e-01 EPA, 1994a beef-forage 4.80e-01 EPA, 1994a dairy-grain & silage 9.00e-01 EPA, 1994a dairy-forage 8.00e-02 EPA, 1994a poultry & eggs-silage 8.50e-01 Lorber, 1997 poultry & eggs-forage 5.00e-02 Lorber, 1997Soil diet fraction; unitless beef 4.00e-02 EPA, 1994a dairy 2.00e-02 EPA, 1994a
Table G-1a. (Continued)

Input parameter Input value Source of curr

G-7

 poultry & eggs 1.00e-01 Lorber, 1997

Bioavialibility of contaminant in soil, unitless 6.50e-01 EPA, 1994aTotal soil loss constant, yr-1 2.77e-02 Lorber, 1997Fraction dermally absorbed, unitless 3.00e-02 EPA, 1994aCancer potency factor, kg-d/mg 1.56e+05 EPA, 1994aFish lipid content correction term 7.00e-02 EPA, 1994aFraction org. C in bottom sediment, unitless 3.00e-02 EPA, 1994aAveraging time, days 2.56e+04 EPA, 1994aLand area of fallout, km 3.73e+01 EPA, 1996d 2

Impervious surface area, km 3.73e+00 10% Land Area Fallout 2Effective diameter of cont. area, m 2.00e+02 Lorber, 1997Soil bulk density, g/cm 1.40e+00 EPA, 1996d 3Watershed depth of incorporation, cm (no till) 2.00e+00 Brzuzy, 1995Mixing depth for soil tillage, cm 2.00e+01 EPA, 1996dSoil vol. water content, mL/cm 2.00e-01 EPA, 1993 3Erosivity factor, kg/km -yr 1.60e+02 EPA, 1996d 2Erodibility factor, t/ac 3.00e-01 EPA, 1996dTopographic factor, unitless 2.00e-01 EPA, 1996dCover management factor, unitless 1.00e-01 EPA, 1996dSupport practice factor, unitless 1.00e+00 EPA, 1996dSediment delivery ratio to water body 2.00e-01 EPA, 1993 equation rounded to eqPollutant enrichment factor 2.00e+00 EPA, 1996dWater body surface area, km 2.49e+00 EPA, 1996d 2Water column volume, m 1.24e+07 EPA, 1996d 3Annual volumetric flow rate, m /yr 1.00e+07 EPA, 1993 3Average current velocity, m/sec 7.50e-01 EPA, 1993Average water temperature, (C) 1.50e+01 EPA, 1993

Suspended solids concentration, mg/L 5.00e+00 EPA, 1993Benthic sediment concentration, kg/L 1.00e+00 EPA, 1996d
Table G-1a. (Continued)

Input parameter Input value Source of curr

G-8

Upper benthic sediment depth, m 2.00e-02 EPA, 1993

Lifetime, years 7.00e+01 EPA, 1994aExposure time, hours/day 1.80e+01 EPA, 1994aExposure frequency, days/yr 3.65e+02 Max conservative estimate
G-9

Table G-1b. Scenario Independent, Dioxin Congener Dependent Parameters

Input parameter

 CDDs CDFs 

2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 12346

Toxicity

equivalency 1 0.5 0.1 0.1 0.1 0.01 0 0.1 0.5 0.05 0.1 0.1 0.1 0.1 0.01 factor

Molecular 3.22 3.56 3.91 3.91 3.91 4.25 4.61 3.06 3.40 3.40 3.75 3.75 3.75 3.75 4.09 weight, g/mole e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+02 e+0Henry'sconstant,atm-m3/mole

1.60 2.60 1.20 1.20 1.20 7.50 7.00 8.60 6.20 6.20 1.40 6.10 1.00 1.00 5.30 

e-05 e-06 e-05 e-05 e-05 e-06 e-09 e-06 e-06 e-06 e-05 e-06 e-05 e-05 e-05 

Soil water part.coeff., ml/g(L/kg)

2.70 2.70 3.80 1.20 1.20 9.80 2.40 2.10 5.10 3.80 1.20 1.20 1.20 1.20 4.90 

e+04 e+04 e+05 e+05 e+05 e+05 e+05 e+04 e+04 e+04 e+05 e+05 e+05 e+07 e+0

Susp. sed-water 1.35 1.35 1.90 6.00 6.00 4.90 1.20 1.05 2.55 1.90 6.00 6.00 6.00 6.00 2.45 part. coeff., L/kg e+05 e+05 e+06 e+05 e+05 e+06 e+06 e+05 e+05 e+05 e+05 e+05 e+05 e+07 e+0Benthicsed-water part.coeff., L/kg

8.10 8.10 1.14 3.60 3.60 2.94 7.20 6.30 1.53 1.14 3.60 3.60 3.60 3.60 1.47 

e+04 e+04 e+06 e+05 e+05 e+06 e+05 e+04 e+05 e+05 e+05 e+05 e+05 e+07 e+0

Plant-soil bioconcentration factor grains 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05 

e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 

 legumes 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05 

e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 

 potatoes 3.92 3.92 3.02 3.02 1.26 6.27 2.11 3.23 6.42 5.13 1.26 1.26 1.26 1.26 3.64 

e+03 e+03 e+04 e+04 e+04 e+04 e+04 e+03 e+03 e+03 e+04 e+04 e+04 e+04 e+0

 root 3.92 3.92 3.02 3.02 1.26 6.27 2.11 3.23 6.42 5.13 1.26 1.26 1.26 1.26 3.64  vegetables e+03 e+03 e+04 e+04 e+04 e+04 e+04 e+03 e+03 e+03 e+04 e+04 e+04 e+04 e+0 fruits 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05 

e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 

 fruiting 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05  vegetables e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03  leafy 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05  vegetables e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03  forage 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05 

e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 

 silage 5.60 5.60 1.21 1.21 2.33 7.01 1.59 6.50 3.88 4.59 2.33 2.33 2.33 2.33 1.05 

e-03 e-03 e-03 e-03 e-03 e-04 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 e-03 

Air-to-plant biotransfer factor grains 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30 

e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0
Table G-1b. (Continued)

Input parameter

 CDDs CDFs 

2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 12346

G-10

 legumes 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30 

e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0

 fruits 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30 

e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0

 fruiting 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30  vegetables e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0 leafy 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30  vegetables e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0 forage 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30 

e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0

 silage 6.55 2.39 5.20 5.20 5.20 1.00 2.36 4.57 9.75 9.75 1.62 1.62 1.62 1.62 8.30 

e+04 e+05 e+05 e+05 e+05 e+06 e+06 e+04 e+04 e+04 e+05 e+05 e+05 e+05 e+0

Bioconcentration ratio of contaminant, unitless beef 5.76 5.55 2.69 2.99 2.32 0.48 0.69 1.25 4.13 0.97 3.12 2.67 2.72 2.37 0.55 

 dairy 5.76 5.55 2.69 2.99 2.32 0.48 0.69 1.25 4.13 0.97 3.12 2.67 2.72 2.37 0.55  poultry 8.8 6.8 3.6 2.4 5.6 1.4 0.3 3.1 7.4 18 4.8 5.3 4.1 2.1 1  eggs 7.8 6 5.4 4.5 10.2 4.8 4.3 2.7 7.8 20.5 7.4 8.2 6.2 3 3.1 
G-11

Appendix G-2

Dispersion Modeling Results (Chapter 11) 

for Dioxin Congener Specific Emission Rates
xx
G-13

Table G-2. Dioxin Congener Specific Annual Average Dispersion Modeling Res

Plants

ISCST3 Dispersion Modeling Results

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

LCH - fisher & farmer - 500 m 

Particles

Inhalable

air conc. 00000000000000

(μg/m3)

Wet dep. 

(g/m2) 1.62e-11 6.92e-10 1.98e-09 1.58e-09 1.22e-09 1.24e-09 1.28e-08 2.52e-10 1.60e-09 3.12e-10 2.67e-09 8.24e-10 1.65e-09 3.32e-09 

Dry dep. 

(g/m2) 00000000000000

Vapors

Air conc.

(μg/m3) 00000000000000

Wet dep. 

(g/m2) 1.61e-11 1.97e-10 1.21e-10 2.61e-11 4.14e-11 2.05e-11 0 5.00e-10 5.57e-10 1.83e-10 1.38e-10 4.27e-11 1.66e-10 2.03e-10 

Dry dep. 

(g/m2) 00000000000000

LOH - fisher & farmer - 500 m

Particles

Inhalable

air conc. 00000000000000

(μg/m3)

Wet dep. 

(g/m2) 3.68e-10 5.35e-10 1.45e-09 1.03e-09 6.55e-10 2.44e-09 2.91e-09 1.67e-10 4.21e-10 3.15e-10 7.14e-10 4.52e-10 6.48e-10 5.56e-10 

Dry dep. 

(g/m2) 00000000000000

Vapors

Air conc.

(μg/m3) 00000000000000

Wet dep. 

(g/m2) 2.49e-10 1.04e-10 6.03e-11 1.16e-11 1.51e-11 2.76e-11 0 2.26e-10 1.00e-10 1.26e-10 2.52e-11 1.60e-11 4.43e-11 2.32e-11 

Dry dep. 

(g/m2) 00000000000000

SCH - fisher & farmer - 500 m

Particles

Inhalable

air conc. 00000000000000

(μg/m3)

Wet dep. 

(g/m2) 2.45e-10 1.05e-10 2.99e-10 2.38e-10 1.85e-10 1.87e-10 1.94e-09 3.80e-11 2.42e-10 4.72e-11 4.04e-10 1.25e-10 2.50e-10 5.00e-10 

Dry dep. 

(g/m2) 00000000000000

Vapors

Air conc.

(μg/m3) 00000000000000
Table G-2. (Continued)

ISCST3 Dispersion Modeling Results

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-14

Wet dep. 

(g/m2) 2.44e-10 3.00e-11 1.83e-11 3.96e-12 6.29e-12 3.11e-12 0 7.58e-11 8.44e-11 2.79e-11 2.10e-11 6.48e-12 2.52e-11 3.07e-11 

Dry dep. 

(g/m2) 00000000000000

SOH - fisher & farmer - 500 m

Particles

Inhalable

air conc. 1.07e-15 1.55e-15 4.21e-15 2.99e-15 1.91e-15 7.11e-15 8.45e-15 4.86e-16 1.23e-15 9.17e-16 2.07e-15 1.32e-15 1.89e-15 1.62e-15 

(μg/m3)

Wet dep. 

(g/m2) 2.02e-11 2.93e-11 7.95e-11 5.65e-11 3.60e-11 1.34e-10 1.60e-10 9.18e-12 2.32e-11 1.73e-11 3.91e-11 2.49e-11 3.57e-11 3.05e-11 

Dry dep. 

(g/m2) 1.24e-15 1.81e-15 4.90e-15 3.48e-15 2.22e-15 8.27e-15 9.83e-15 5.66e-16 1.43e-15 1.07e-15 2.41e-15 1.53e-15 2.20e-15 1.88e-15 

Vapors

Air conc.

(μg/m3) 1.52e-15 6.35e-16 3.68e-16 7.10e-17 9.23e-17 1.69e-16 0 1.38e-15 6.13e-16 7.72e-16 1.54e-16 9.78e-17 2.72e-16 1.42e-16 

Wet dep. 

(g/m2) 1.38e-11 5.77e-12 3.35e-12 6.45e-13 8.39e-13 1.53e-12 0 1.26e-11 5.56e-12 7.02e-12 1.40e-12 8.88e-13 2.47e-12 1.29e-12 

Dry dep. 

(g/m2) 9.58e-17 4.00e-17 2.32e-17 4.48e-18 5.82e-18 1.06e-17 0 8.73e-17 3.86e-17 4.87e-17 9.71e-18 6.17e-18 1.71e-17 8.93e-18 

LCC - resident & child - 5000 m

Particles

Inhalable

air conc. 4.68e-14 2.00e-12 5.71e-12 4.55e-12 3.53e-12 3.57e-12 3.70e-11 7.27e-13 4.63e-12 9.00e-13 7.70e-12 2.38e-12 4.78e-12 9.59e-12 

(μg/m3)

Wet dep. 

(g/m2) 5.20e-13 2.22e-11 6.35e-11 5.06e-11 3.93e-11 3.97e-11 4.11e-10 8.08e-12 5.14e-11 1.00e-11 8.56e-11 2.65e-11 5.31e-11 1.07e-10 

Dry dep. 

(g/m2) 3.89e-14 1.66e-12 4.75e-12 3.79e-12 2.94e-12 2.97e-12 3.08e-11 6.05e-13 3.85e-12 7.49e-13 6.41e-12 1.98e-12 3.97e-12 7.98e-12 

Vapors

Air conc.

(μg/m3) 1.08e-13 1.33e-12 8.12e-13 1.76e-13 2.78e-13 1.38e-13 0 3.37e-12 3.75e-12 1.23e-12 9.29e-13 2.87e-13 1.12e-12 1.37e-12 

Wet dep. 

(g/m2) 6.68e-13 8.20e-12 5.02e-12 1.09e-12 1.72e-12 8.50e-13 0 2.08e-11 2.32e-11 7.61e-12 5.74e-12 1.78e-12 6.90e-12 8.44e-12 

Dry dep. 

(g/m2) 6.82e-15 8.37e-14 5.12e-14 1.11e-14 1.76e-14 8.68e-15 0 2.12e-13 2.36e-13 7.77e-14 5.86e-14 1.81e-14 7.04e-14 8.61e-14 

LOC - resident & child - 5000 m

Particles

Inhalable

air conc. 1.39e-12 2.02e-12 5.46e-12 3.87e-12 2.47e-12 9.21e-12 1.10e-11 6.29e-13 1.59e-12 1.19e-12 2.69e-12 1.71e-12 2.44e-12 2.10e-12 

(μg/m3)

Wet dep. 

(g/m2) 9.63e-12 1.40e-11 3.79e-11 2.69e-11 1.72e-11 6.39e-11 7.61e-11 4.37e-12 1.10e-11 8.26e-12 1.87e-11 1.18e-11 1.70e-11 1.46e-11 

Dry dep. 

(g/m2) 2.29e-12 3.32e-12 8.98e-12 6.37e-12 4.07e-12 1.51e-11 1.81e-11 1.04e-12 2.62e-12 1.96e-12 4.43e-12 2.81e-12 4.02e-12 3.45e-12 
Table G-2. (Continued)

ISCST3 Dispersion Modeling Results

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-15

Vapors

Air conc.

(μg/m3) 3.40e-12 1.42e-12 8.22e-13 1.58e-13 2.06e-13 3.76e-13 0 3.08e-12 1.36e-12 1.72e-12 3.44e-13 2.18e-13 6.05e-13 3.16e-13 

Wet dep. 

(g/m2) 1.01e-11 4.21e-12 2.43e-12 4.68e-13 6.11e-13 1.11e-12 0 9.14e-12 4.04e-12 5.11e-12 1.02e-12 6.45e-13 1.79e-12 9.36e-13 

Dry dep. 

(g/m2) 2.14e-13 8.96e-14 5.18e-14 9.97e-15 1.30e-14 2.37e-14 0 1.94e-13 8.60e-14 1.09e-13 2.17e-14 1.37e-14 3.81e-14 1.99e-14 

SCC - resident & child - 5000 m

Particles

Inhalable

air conc. 4.25e-12 1.82e-12 5.19e-12 4.13e-12 3.21e-12 3.24e-12 3.36e-11 6.59e-13 4.19e-12 8.20e-13 7.01e-12 2.16e-12 4.34e-12 8.67e-12 

(μg/m3)

Wet dep. 

(g/m2) 7.24e-12 3.10e-12 8.85e-12 7.05e-12 5.48e-12 5.53e-12 5.73e-11 1.12e-12 7.15e-12 1.40e-12 1.20e-11 3.69e-12 7.40e-12 1.48e-11 

Dry dep. 

(g/m2) 3.79e-12 1.62e-12 4.63e-12 3.69e-12 2.87e-12 2.89e-12 2.99e-11 5.88e-13 3.74e-12 7.31e-13 6.26e-12 1.93e-12 3.87e-12 7.74e-12 

Vapors

Air conc.

(μg/m3) 9.87e-12 1.21e-12 7.43e-13 1.60e-13 2.55e-13 1.26e-13 0 3.07e-12 3.42e-12 1.13e-12 8.51e-13 2.63e-13 1.02e-12 1.24e-12 

Wet dep. 

(g/m2) 9.44e-12 1.16e-12 7.10e-13 1.53e-13 2.43e-13 1.20e-13 0 2.93e-12 3.27e-12 1.08e-12 8.14e-13 2.51e-13 9.75e-13 1.19e-12 

Dry dep. 

(g/m2) 6.23e-13 7.65e-14 4.68e-14 1.01e-14 1.61e-14 7.93e-15 0 1.94e-13 2.16e-13 7.12e-14 5.37e-14 1.66e-14 6.44e-14 7.83e-14 

SOC - resident & child - 5000 m

Particles

Inhalable

air conc. 6.35e-13 9.23e-13 2.50e-12 1.78e-12 1.13e-12 4.22e-12 5.02e-12 2.89e-13 7.30e-13 5.45e-13 1.23e-12 7.82e-13 1.12e-12 9.61e-13 

(μg/m3)

Wet dep. 

(g/m2) 4.76e-13 6.92e-13 1.87e-12 1.33e-12 8.49e-13 3.17e-12 3.76e-12 2.17e-13 5.47e-13 4.09e-13 9.24e-13 5.87e-13 8.42e-13 7.21e-13 

Dry dep. 

(g/m2) 1.13e-12 1.64e-12 4.44e-12 3.16e-12 2.01e-12 7.51e-12 8.93e-12 5.14e-13 1.30e-12 9.69e-13 2.19e-12 1.39e-12 2.00e-12 1.71e-12 

Vapors

Air conc.

(μg/m3) 1.60e-12 6.67e-13 3.87e-13 7.46e-14 9.70e-14 1.77e-13 0 1.45e-12 6.43e-13 8.11e-13 1.62e-13 1.03e-13 2.85e-13 1.49e-13 

Wet dep. 

(g/m2) 4.77e-13 1.99e-13 1.16e-13 2.23e-14 2.90e-14 5.30e-14 0 4.35e-13 1.92e-13 2.43e-13 4.83e-14 3.07e-14 8.53e-14 4.45e-14 

Dry dep. 

(g/m2) 1.01e-13 4.20e-14 2.44e-14 4.70e-15 6.12e-15 1.12e-14 0 9.17e-14 4.06e-14 5.12e-14 1.02e-14 6.48e-15 1.80e-14 9.38e-15 
x
G-17

Appendix G-3

IEM Modeled Dioxin Congener Specific Environmental Media Concentrations
xx
G-19

Table G-3. IEM Modeled Dioxin Congener Specific Environmental Media Concentrati

PCDDs PCDFs

Scenario 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

LCH - fisher & farmer

Soil 2.3e-11 6.5e-10 1.5e-09 1.2e-09 9.2e-10 9.1e-10 9.3e-09 5.5e-10 1.6e-09 3.6e-10 2.0e-09 6.3e-10 1.3e-09 2.6e-09 concentration,mg/gWater 3.7e-12 1.0e-10 1.1e-10 1.0e-10 7.9e-11 6.2e-11 7.0e-10 9.9e-11 1.8e-10 4.7e-11 1.8e-10 5.4e-11 1.1e-10 1.7e-10 concentration,mg/LWhole Fish 3.7e-08 1.0e-06 1.1e-06 8.4e-07 6.6e-07 8.3e-08 1.7e-08 8.6e-07 2.5e-06 5.8e-07 1.5e-06 4.5e-07 9.5e-07 1.9e-06 concentration,mg/kgPlant concentration, μg/g grains 1.3e-11 3.6e-10 1.8e-10 1.4e-10 2.1e-10 6.4e-11 1.5e-09 3.6e-10 6.1e-10 1.7e-10 4.8e-10 1.5e-10 3.1e-10 6.0e-10  legumes 9.2e-11 2.5e-09 5.3e-09 4.0e-09 3.3e-09 3.1e-09 3.3e-08 2.2e-09 5.9e-09 1.4e-09 7.3e-09 2.3e-09 4.7e-09 9.2e-09 

 potatoes 8.5e-11 2.3e-09 3.0e-09 7.3e-09 2.4e-09 1.5e-09 2.0e-08 2.1e-09 4.9e-09 1.2e-09 5.4e-09 1.7e-09 3.5e-09 6.7e-11  root 8.5e-11 2.3e-09 3.0e-09 7.3e-09 2.4e-09 1.5e-09 2.0e-08 2.1e-09 4.9e-09 1.2e-09 5.4e-09 1.7e-09 3.5e-09 6.7e-11  vegetables fruits 4.9e-10 1.4e-08 3.1e-08 2.4e-08 1.9e-08 1.9e-08 1.9e-07 1.1e-08 3.3e-08 7.5e-09 4.2e-08 1.3e-08 2.7e-08 5.3e-08  fruiting 4.9e-10 1.4e-08 3.1e-08 2.4e-08 1.9e-08 1.9e-08 1.9e-07 1.1e-08 3.3e-08 7.5e-09 4.2e-08 1.3e-08 2.7e-08 5.3e-08  vegetables leafy 9.3e-10 2.6e-08 6.0e-08 4.6e-08 3.6e-08 3.6e-08 3.6e-07 2.2e-08 6.2e-08 1.4e-08 8.0e-08 2.5e-08 5.2e-08 1.0e-07 

 vegetables forage 1.7e-09 4.6e-08 1.0e-07 7.8e-08 6.3e-08 6.1e-08 6.3e-07 4.0e-08 1.1e-07 2.5e-08 1.4e-07 4.3e-08 9.0e-08 1.8e-07  silage 5.5e-10 1.5e-08 3.5e-08 2.7e-08 2.1e-08 2.1e-08 2.1e-07 1.3e-08 3.6e-08 8.4e-09 4.7e-08 1.5e-08 3.0e-08 5.9e-08 Animal concentrations, μg/g beef 6.7e-09 1.8e-07 2.0e-07 1.7e-07 1.0e-07 2.1e-08 3.1e-07 3.4e-08 3.2e-07 1.7e-08 3.1e-07 8.1e-08 1.7e-07 2.9e-07  dairy 3.0e-09 7.9e-08 8.9e-08 7.5e-08 4.6e-08 9.4e-09 1.4e-07 1.5e-08 1.4e-07 7.7e-09 1.4e-07 3.6e-08 7.7e-08 1.3e-07  poultry 1.8e-08 3.9e-07 4.8e-07 2.5e-07 4.5e-07 1.1e-07 2.5e-07 1.5e-07 1.0e-06 5.7e-07 8.6e-07 2.9e-07 4.8e-07 4.7e-07  eggs 1.6e-08 3.4e-07 7.2e-07 4.6e-07 8.3e-07 3.8e-07 3.5e-06 1.3e-07 1.1e-06 6.5e-07 1.3e-06 4.6e-07 7.2e-07 6.8e-07 

LOH - fisher & farmer

Soil 4.5e-10 4.7e-10 1.1e-09 7.5e-10 4.9e-10 1.8e-09 2.1e-09 2.9e-10 3.8e-10 3.2e-10 5.4e-10 3.4e-10 5.0e-10 4.2e-10 concentration,mg/gWater 7.0e-11 7.3e-11 7.8e-11 6.5e-11 4.2e-11 1.2e-10 1.6e-10 5.2e-11 4.3e-11 4.2e-11 4.6e-11 2.9e-11 4.3e-11 2.7e-11 concentration,mg/L
Table G-3. (Continued)

PCDDs PCDFs

Scenario 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-20

Whole fish 7.1e-07 7.4e-07 7.9e-07 5.4e-07 3.5e-07 1.6e-07 3.8e-09 4.5e-07 6.1e-07 5.1e-07 3.9e-07 2.5e-07 3.6e-07 3.1e-07 

concentration,mg/kgPlant concentration, μg/g grains 2.5e-10 2.6e-10 1.3e-10 9.1e-11 1.1e-10 1.3e-10 3.4e-10 1.9e-10 1.5e-10 1.5e-10 1.3e-10 7.9e-11 1.2e-10 9.8e-11 

 legumes 1.8e-09 1.8e-09 3.8e-09 2.6e-09 1.7e-09 6.1e-09 7.4e-09 1.1e-09 1.4e-09 1.2e-09 1.9e-09 1.2e-09 1.8e-09 1.5e-09  potatoes 1.6e-09 1.7e-09 2.2e-09 4.7e-09 1.3e-09 2.9e-09 4.6e-09 1.1e-09 1.2e-09 1.1e-09 1.4e-09 8.9e-10 1.3e-09 1.1e-11  root 1.6e-09 1.7e-09 2.2e-09 4.7e-09 1.3e-09 2.9e-09 4.6e-09 1.1e-09 1.2e-09 1.1e-09 1.4e-09 8.9e-10 1.3e-09 1.1e-11  vegetables fruits 9.4e-09 9.7e-09 2.2e-08 1.5e-08 1.0e-08 3.7e-08 4.3e-08 6.0e-09 7.9e-09 6.7e-09 1.1e-08 7.0e-09 1.0e-08 8.7e-09  fruiting 9.4e-09 9.7e-09 2.2e-08 1.5e-08 1.0e-08 3.7e-08 4.3e-08 6.0e-09 7.9e-09 6.7e-09 1.1e-08 7.0e-09 1.0e-08 8.7e-09  vegetables

 leafy 1.8e-08 1.8e-08 4.3e-08 2.9e-08 1.9e-08 7.0e-08 8.3e-08 1.1e-08 1.5e-08 1.3e-08 2.1e-08 1.3e-08 2.0e-08 1.6e-08 

 vegetables forage 3.2e-08 3.3e-08 7.4e-08 5.1e-08 3.3e-08 1.2e-07 1.4e-07 2.1e-08 2.6e-08 2.3e-08 3.7e-08 2.3e-08 3.4e-08 2.9e-08  silage 1.0e-08 1.1e-08 2.5e-08 1.7e-08 1.1e-08 4.1e-08 4.9e-08 6.7e-09 8.8e-09 7.5e-09 1.2e-08 7.8e-09 1.2e-08 9.7e-09 Animal concentrations, μg/g beef 1.3e-07 1.3e-07 1.4e-07 1.1e-07 5.5e-08 4.1e-08 7.0e-08 1.8e-08 7.7e-08 1.5e-08 8.1e-08 4.4e-08 6.5e-08 4.8e-08  dairy 5.7e-08 5.7e-08 6.4e-08 4.9e-08 2.5e-08 1.9e-08 3.2e-08 8.0e-09 3.4e-08 6.9e-09 3.6e-08 2.0e-08 2.9e-08 2.2e-08 

 poultry 3.5e-07 2.8e-07 3.5e-07 1.6e-07 2.4e-07 2.2e-07 5.6e-08 7.8e-08 2.5e-07 5.1e-07 2.3e-07 1.6e-07 1.8e-07 7.8e-08  eggs 3.1e-07 2.5e-07 5.2e-07 3.0e-07 4.4e-07 7.6e-07 8.0e-07 6.8e-08 2.6e-07 5.8e-07 3.5e-07 2.5e-07 2.7e-07 1.1e-07 

SCH - fisher & farmer

Soil 3.6e-10 9.8e-11 2.3e-10 1.8e-10 1.4e-10 1.4e-10 1.4e-09 8.3e-11 2.4e-10 5.5e-11 3.1e-10 9.5e-11 2.0e-10 3.9e-10 concentration,mg/gWater 5.6e-11 1.5e-11 1.6e-11 1.5e-11 1.2e-11 9.3e-12 1.1e-10 1.5e-11 2.7e-11 7.1e-12 2.7e-11 8.2e-12 1.7e-11 2.5e-11 concentration,mg/LWhole fish 5.6e-07 1.6e-07 1.7e-07 1.3e-07 1.0e-07 1.3e-08 2.5e-09 1.3e-07 3.8e-07 8.7e-08 2.2e-07 6.9e-08 1.4e-07 2.8e-07 concentration,mg/kgPlant concentration, μg/g grains 2.0e-10 5.5e-11 2.8e-11 2.1e-11 3.2e-11 9.7e-12 2.2e-10 5.4e-11 9.2e-11 2.5e-11 7.2e-11 2.2e-11 4.7e-11 9.0e-11  legumes 1.4e-09 3.8e-10 8.0e-10 6.1e-10 5.0e-10 4.7e-10 4.9e-09 3.3e-10 8.9e-10 2.1e-10 1.1e-09 3.4e-10 7.2e-10 1.4e-09 
Table G-3. (Continued)

PCDDs PCDFs

Scenario 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-21

 potatoes 1.3e-09 3.6e-10 4.6e-10 1.1e-09 3.7e-10 2.2e-10 3.1e-09 3.2e-10 7.5e-10 1.8e-10 8.1e-10 2.5e-10 5.3e-10 1.0e-11 

 root 1.3e-09 3.6e-10 4.6e-10 1.1e-09 3.7e-10 2.2e-10 3.1e-09 3.2e-10 7.5e-10 1.8e-10 8.1e-10 2.5e-10 5.3e-10 1.0e-11 

 vegetables fruits 7.4e-09 2.0e-09 4.7e-09 3.6e-09 2.9e-09 2.8e-09 2.9e-08 1.7e-09 4.9e-09 1.1e-09 6.4e-09 2.0e-09 4.1e-09 7.9e-09  fruiting 7.4e-09 2.0e-09 4.7e-09 3.6e-09 2.9e-09 2.8e-09 2.9e-08 1.7e-09 4.9e-09 1.1e-09 6.4e-09 2.0e-09 4.1e-09 7.9e-09  vegetables leafy 1.4e-08 3.9e-09 9.0e-09 6.9e-09 5.5e-09 5.4e-09 5.5e-08 3.3e-09 9.3e-09 2.2e-09 1.2e-08 3.7e-09 7.8e-09 1.5e-08  vegetables sforage 2.5e-08 7.0e-09 1.6e-08 1.2e-08 9.5e-09 9.2e-09 9.5e-08 6.0e-09 1.7e-08 3.9e-09 2.1e-08 6.5e-09 1.4e-08 2.6e-08  silage 8.3e-09 2.3e-09 5.3e-09 4.0e-09 3.2e-09 3.2e-09 3.2e-08 1.9e-09 5.5e-09 1.3e-09 7.1e-09 2.2e-09 4.6e-09 8.9e-09 Animal concentrations, μg/g beef 1.0e-07 2.7e-08 3.0e-08 2.5e-08 1.6e-08 3.2e-09 4.7e-08 5.2e-09 4.8e-08 2.6e-09 4.7e-08 1.2e-08 2.6e-08 4.4e-08 

 dairy 4.5e-08 1.2e-08 1.3e-08 1.1e-08 7.0e-09 1.4e-09 2.1e-08 2.3e-09 2.1e-08 1.2e-09 2.1e-08 5.5e-09 1.2e-08 2.0e-08  poultry 2.8e-07 5.9e-08 7.3e-08 3.7e-08 6.9e-08 1.7e-08 3.7e-08 2.3e-08 1.5e-07 8.7e-08 1.3e-07 4.4e-08 7.2e-08 7.1e-08  eggs 2.5e-07 5.2e-08 1.1e-07 7.0e-08 1.3e-07 5.8e-08 5.3e-07 2.0e-08 1.6e-07 9.9e-08 2.0e-07 6.9e-08 1.1e-07 1.0e-07 

SOH - fisher & farmer

Soil 2.5e-11 2.6e-11 6.0e-11 4.2e-11 2.7e-11 9.9e-11 1.2e-10 1.6e-11 2.1e-11 1.8e-11 3.0e-11 1.9e-11 2.8e-11 2.3e-11 

concentration,mg/g

Water 3.9e-12 4.0e-12 4.3e-12 3.6e-12 2.3e-12 6.7e-12 8.7e-12 2.9e-12 2.4e-12 2.3e-12 2.5e-12 1.6e-12 2.4e-12 1.5e-12 

concentration,mg/LWhole Fish 3.9e-08 4.1e-08 4.4e-08 3.0e-08 1.9e-08 8.9e-09 2.1e-10 2.5e-08 3.4e-08 2.8e-08 2.1e-08 1.4e-08 2.0e-08 1.7e-08 concentration,mg/kgPlant concentration, μg/g grains 1.4e-11 1.4e-11 7.4e-12 5.0e-12 6.3e-12 7.0e-12 1.8e-11 1.0e-11 8.1e-12 8.2e-12 6.9e-12 4.4e-12 6.5e-12 5.4e-12  legumes 9.7e-11 1.0e-10 2.1e-10 1.4e-10 9.6e-11 3.4e-10 4.1e-10 6.3e-11 7.8e-11 6.7e-11 1.1e-10 6.7e-11 9.9e-11 8.3e-11  potatoes 9.0e-11 9.3e-11 1.2e-10 2.6e-10 7.0e-11 1.6e-10 2.6e-10 6.1e-11 6.6e-11 6.0e-11 7.7e-11 4.9e-11 7.3e-11 6.1e-13  root 9.0e-11 9.3e-11 1.2e-10 2.6e-10 7.0e-11 1.6e-10 2.6e-10 6.1e-11 6.6e-11 6.0e-11 7.7e-11 4.9e-11 7.3e-11 6.1e-13  vegetables fruits 5.2e-10 5.3e-10 1.2e-09 8.5e-10 5.5e-10 2.0e-09 2.4e-09 3.3e-10 4.3e-10 3.7e-10 6.1e-10 3.9e-10 5.7e-10 4.8e-10  fruiting 5.2e-10 5.3e-10 1.2e-09 8.5e-10 5.5e-10 2.0e-09 2.4e-09 3.3e-10 4.3e-10 3.7e-10 6.1e-10 3.9e-10 5.7e-10 4.8e-10  vegetables
Table G-3. (Continued)

PCDDs PCDFs

Scenario 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-22

 leafy 9.8e-10 1.0e-09 2.4e-09 1.6e-09 1.0e-09 3.9e-09 4.5e-09 6.3e-10 8.2e-10 7.0e-10 1.2e-09 7.3e-10 1.1e-09 9.1e-10 

 vegetables forage 1.8e-09 1.8e-09 4.0e-09 2.8e-09 1.8e-09 6.6e-09 7.8e-09 1.1e-09 1.5e-09 1.2e-09 2.0e-09 1.3e-09 1.9e-09 1.6e-09 

 silage 5.8e-10 6.0e-10 1.4e-09 9.5e-10 6.2e-10 2.3e-09 2.7e-09 3.7e-10 4.9e-10 4.1e-10 6.8e-10 4.3e-10 6.4e-10 5.3e-10 Animal concentrations, μg/g beef 7.1e-09 7.0e-09 7.8e-09 5.9e-09 3.0e-09 2.3e-09 3.9e-09 9.9e-10 4.2e-09 8.5e-10 4.4e-09 2.4e-09 3.6e-09 2.7e-09  dairy 3.1e-09 3.1e-09 3.5e-09 2.7e-09 1.3e-09 1.0e-09 1.7e-09 4.4e-10 1.9e-09 3.8e-10 2.0e-09 1.1e-09 1.6e-09 1.2e-09  poultry 1.9e-08 1.5e-08 1.9e-08 8.8e-09 1.3e-08 1.2e-08 3.1e-09 4.3e-09 1.4e-08 2.8e-08 1.2e-08 8.7e-09 1.0e-08 4.3e-09  eggs 1.7e-08 1.4e-08 2.9e-08 1.6e-08 2.4e-08 4.2e-08 4.4e-08 3.8e-09 1.4e-08 3.2e-08 1.9e-08 1.4e-08 1.5e-08 6.1e-09 

LCC - Resident & Child

Soil 9.0e-13 2.3e-11 5.3e-11 4.0e-11 3.2e-11 3.2e-11 3.2e-10 2.2e-11 5.7e-11 1.3e-11 7.1e-11 2.2e-11 4.7e-11 9.0e-11 

concentration,

mg/gWater 1.4e-13 3.6e-12 3.8e-12 3.4e-12 2.7e-12 2.1e-12 2.4e-11 3.8e-12 6.4e-12 1.7e-12 6.1e-12 1.9e-12 4.0e-12 5.8e-12 

concentration,mg/LPlant concentration, μg/g grains 3.5e-12 1.5e-10 1.8e-10 4.3e-11 6.8e-11 6.0e-11 5.1e-11 7.9e-11 1.8e-10 5.7e-11 8.0e-11 2.5e-11 8.7e-11 1.1e-10  legumes 3.6e-12 9.6e-11 1.9e-10 1.5e-10 1.2e-10 1.1e-10 1.2e-09 8.9e-11 2.2e-10 5.3e-11 2.6e-10 8.2e-11 1.7e-10 3.3e-10 

 potatoes 3.3e-12 8.5e-11 1.1e-10 2.5e-10 8.4e-11 5.1e-11 7.1e-10 8.3e-11 1.8e-10 4.5e-11 1.9e-10 5.8e-11 1.2e-10 2.4e-12  root 3.3e-12 8.5e-11 1.1e-10 2.5e-10 8.4e-11 5.1e-11 7.1e-10 8.3e-11 1.8e-10 4.5e-11 1.9e-10 5.8e-11 1.2e-10 2.4e-12  vegetables fruits 1.9e-11 5.1e-10 1.1e-09 8.6e-10 6.8e-10 6.7e-10 6.8e-09 4.6e-10 1.2e-09 2.9e-10 1.5e-09 4.7e-10 9.9e-10 1.9e-09  fruiting 1.9e-11 5.1e-10 1.1e-09 8.6e-10 6.8e-10 6.7e-10 6.8e-09 4.6e-10 1.2e-09 2.9e-10 1.5e-09 4.7e-10 9.9e-10 1.9e-09  vegetables leafy 3.6e-11 9.5e-10 2.2e-09 1.6e-09 1.3e-09 1.3e-09 1.3e-08 8.7e-10 2.3e-09 5.4e-10 2.9e-09 8.9e-10 1.9e-09 3.6e-09  vegetables

LOC - Resident & Child

Soil 1.6e-11 1.6e-11 3.6e-11 2.5e-11 1.6e-11 5.8e-11 6.9e-11 1.1e-11 1.3e-11 1.1e-11 1.8e-11 1.1e-11 1.7e-11 1.4e-11 concentration,mg/gWater 2.5e-12 2.4e-12 2.6e-12 2.1e-12 1.4e-12 3.9e-12 5.1e-12 1.9e-12 1.5e-12 1.4e-12 1.5e-12 9.5e-13 1.4e-12 9.0e-13 concentration,mg/L
Table G-3. (Continued)

PCDDs PCDFs

Scenario 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-23

Plant concentration, μg/g

 grains 1.0e-10 1.5e-10 1.8e-10 3.8e-11 4.9e-11 1.6e-10 1.1e-11 6.6e-11 6.1e-11 7.6e-11 2.8e-11 1.7e-11 4.5e-11 2.5e-11 

 legumes 6.8e-11 6.9e-11 1.4e-10 9.4e-11 6.3e-11 2.2e-10 2.6e-10 4.6e-11 5.3e-11 4.7e-11 6.9e-11 4.4e-11 6.6e-11 5.5e-11  potatoes 5.9e-11 5.7e-11 7.1e-11 1.5e-10 4.2e-11 9.3e-11 1.5e-10 4.1e-11 4.1e-11 3.8e-11 4.6e-11 2.9e-11 4.4e-11 3.6e-13  root 5.9e-11 5.7e-11 7.1e-11 1.5e-10 4.2e-11 9.3e-11 1.5e-10 4.1e-11 4.1e-11 3.8e-11 4.6e-11 2.9e-11 4.4e-11 3.6e-13  vegetables fruits 3.6e-10 3.6e-10 8.1e-10 5.6e-10 3.6e-10 1.3e-09 1.6e-09 2.4e-10 2.9e-10 2.5e-10 4.0e-10 2.5e-10 3.8e-10 3.1e-10  fruiting 3.6e-10 3.6e-10 8.1e-10 5.6e-10 3.6e-10 1.3e-09 1.6e-09 2.4e-10 2.9e-10 2.5e-10 4.0e-10 2.5e-10 3.8e-10 3.1e-10  vegetables leafy 6.8e-10 6.8e-10 1.5e-09 1.1e-09 6.9e-10 2.5e-09 3.0e-09 4.4e-10 5.5e-10 4.8e-10 7.6e-10 4.8e-10 7.2e-10 6.0e-10  vegetables

SCC - Resident & Child

Soil 1.5e-11 4.3e-12 1.0e-11 7.9e-12 6.3e-12 6.2e-12 6.3e-11 3.5e-12 1.0e-11 2.4e-12 1.4e-11 4.3e-12 9.0e-12 1.7e-11 

concentration,mg/gWater 2.3e-12 6.6e-13 7.4e-13 6.8e-13 5.3e-13 4.2e-13 4.8e-12 6.2e-13 1.2e-12 3.1e-13 1.2e-12 3.7e-13 7.6e-13 1.1e-12 concentration,mg/LPlant concentration, μg/g

 grains 2.8e-10 1.2e-10 1.6e-10 3.6e-11 5.7e-11 5.3e-11 1.0e-11 6.1e-11 1.4e-10 4.7e-11 6.1e-11 1.9e-11 7.2e-11 8.9e-11 

 legumes 7.1e-11 2.2e-11 4.6e-11 3.3e-11 2.7e-11 2.6e-11 2.6e-10 1.6e-11 4.7e-11 1.1e-11 5.9e-11 1.8e-11 3.9e-11 7.4e-11  potatoes 5.6e-11 1.6e-11 2.1e-11 5.0e-11 1.6e-11 9.9e-12 1.4e-10 1.4e-11 3.3e-11 8.1e-12 3.6e-11 1.1e-11 2.4e-11 4.5e-13  root 5.6e-11 1.6e-11 2.1e-11 5.0e-11 1.6e-11 9.9e-12 1.4e-10 1.4e-11 3.3e-11 8.1e-12 3.6e-11 1.1e-11 2.4e-11 4.5e-13  vegetables fruits 3.6e-10 1.1e-10 2.5e-10 1.9e-10 1.5e-10 1.5e-10 1.5e-09 8.1e-11 2.5e-10 5.7e-11 3.4e-10 1.0e-10 2.2e-10 4.2e-10  fruiting 3.6e-10 1.1e-10 2.5e-10 1.9e-10 1.5e-10 1.5e-10 1.5e-09 8.1e-11 2.5e-10 5.7e-11 3.4e-10 1.0e-10 2.2e-10 4.2e-10  vegetables leafy 6.8e-10 2.0e-10 4.8e-10 3.7e-10 2.9e-10 2.9e-10 3.0e-09 1.5e-10 4.8e-10 1.1e-10 6.5e-10 2.0e-10 4.1e-10 8.0e-10  vegetables

SOC - Resident & Child

Soil 1.6e-12 1.9e-12 4.7e-12 3.3e-12 2.1e-12 7.8e-12 9.2e-12 9.1e-13 1.5e-12 1.2e-12 2.3e-12 1.5e-12 2.1e-12 1.8e-12 concentration,mg/g
Table G-3. (Continued)

PCDDs PCDFs

Scenario 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 

G-24

Water 2.4e-13 2.9e-13 3.4e-13 2.8e-13 1.8e-13 5.3e-13 6.9e-13 1.6e-13 1.7e-13 1.6e-13 2.0e-13 1.3e-13 1.8e-13 1.2e-13 

concentration,mg/LPlant concentration, μg/g grains 4.5e-11 6.8e-11 8.5e-11 1.7e-11 2.2e-11 7.5e-11 1.5e-12 2.9e-11 2.7e-11 3.4e-11 1.2e-11 7.3e-12 2.0e-11 1.1e-11 

 legumes 8.8e-12 1.1e-11 2.4e-11 1.6e-11 1.1e-11 3.8e-11 4.5e-11 5.0e-12 8.0e-12 6.7e-12 1.1e-11 7.3e-12 1.1e-11 9.0e-12  potatoes 5.8e-12 6.8e-12 9.3e-12 2.1e-11 5.5e-12 1.2e-11 2.0e-11 3.5e-12 4.8e-12 4.1e-12 6.1e-12 3.8e-12 5.6e-12 4.7e-14  root 5.8e-12 6.8e-12 9.3e-12 2.1e-11 5.5e-12 1.2e-11 2.0e-11 3.5e-12 4.8e-12 4.1e-12 6.1e-12 3.8e-12 5.6e-12 4.7e-14  vegetables fruits 4.4e-11 5.4e-11 1.3e-10 9.4e-11 6.1e-11 2.2e-10 2.6e-10 2.5e-11 4.3e-11 3.4e-11 6.6e-11 4.2e-11 6.1e-11 5.2e-11  fruiting 4.4e-11 5.4e-11 1.3e-10 9.4e-11 6.1e-11 2.2e-10 2.6e-10 2.5e-11 4.3e-11 3.4e-11 6.6e-11 4.2e-11 6.1e-11 5.2e-11  vegetables

 leafy 8.3e-11 1.0e-10 2.6e-10 1.8e-10 1.2e-10 4.3e-10 5.0e-10 4.6e-11 8.1e-11 6.5e-11 1.3e-10 8.0e-11 1.2e-10 9.8e-11 

 vegetables
G-25

Appendix G-4

Dioxin Congener Specific Lifetime Average Daily Dose (LADD) 

for Each Scenario and Pathway
x

xx
G-27

Table G-4. Dioxin Congener Specific Lifetime Average Daily Dose (LADD) for 

Pathway

Lifetime Average Daily Doses (LADDs)

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234

PCDDs PCDFs 

LCH-fisher

 Soil dermal 2.70e-16 7.43e-15 1.75e-14 1.34e-14 1.06e-14 1.05e-14 1.07e-13 6.28e-15 1.80e-14 4.13e-15 2.34e-14 7.24e-15 1.52e-14 2.94

 Soil ingestion 1.44e-14 3.96e-13 9.35e-13 7.14e-13 5.64e-13 5.59e-13 5.71e-12 3.35e-13 9.62e-13 2.20e-13 1.25e-12 3.86e-13 8.11e-13 1.57

Veget. 5.76e-14 1.59e-12 3.20e-12 2.54e-12 2.02e-12 1.87e-12 1.98e-11 1.38e-12 3.64e-12 8.54e-13 4.47e-12 1.38e-12 2.90e-12 5.49

ingestion

 Water ingest 4.48e-14 1.24e-12 1.33e-12 1.22e-12 9.65e-13 7.52e-13 8.57e-12 1.21e-12 2.18e-12 5.73e-13 2.14e-12 6.61e-13 1.39e-12 2.04

 Fish ingestion 1.37e-11 3.77e-10 4.06e-10 3.09e-10 2.44e-10 3.04e-11 6.19e-12 3.17e-10 9.28e-10 2.11e-10 5.40e-10 1.67e-10 3.51e-10 6.83

 Direct 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e

inhalation

LOH-fisher

 Soil dermal 5.15e-15 5.34e-15 1.26e-14 8.67e-15 5.60e-15 2.06e-14 2.43e-14 3.28e-15 4.36e-15 3.69e-15 6.17e-15 3.91e-15 5.78e-15 4.84

 Soil 2.75e-13 2.85e-13 6.71e-13 4.62e-13 2.99e-13 1.10e-12 1.29e-12 1.75e-13 2.32e-13 1.97e-13 3.29e-13 2.08e-13 3.08e-13 2.58

ingestion

Veget. 1.10e-12 1.14e-12 2.30e-12 1.64e-12 1.07e-12 3.68e-12 4.50e-12 7.21e-13 8.78e-13 7.62e-13 1.18e-12 7.46e-13 1.10e-12 9.02

ingestion

 Water ingest 8.57e-13 8.88e-13 9.56e-13 7.91e-13 5.11e-13 1.48e-12 1.95e-12 6.34e-13 5.27e-13 5.12e-13 5.63e-13 3.57e-13 5.28e-13 3.34

 Fish ingestion 2.62e-10 2.71e-10 2.91e-10 2.00e-10 1.29e-10 5.97e-11 1.40e-12 1.65e-10 2.24e-10 1.89e-10 1.42e-10 9.01e-11 1.33e-10 1.12

 Direct 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e

inhalation

SCH-fisher

 Soil dermal 4.08e-15 1.12e-15 2.65e-15 2.02e-15 1.60e-15 1.59e-15 1.62e-14 9.50e-16 2.72e-15 6.28e-16 3.55e-15 1.10e-15 2.30e-15 4.43

 Soil ingestion 2.18e-13 6.00e-14 1.41e-13 1.08e-13 8.53e-14 8.46e-14 8.62e-13 5.07e-14 1.45e-13 3.35e-14 1.89e-13 5.84e-14 1.23e-13 2.36

Veget. 8.72e-13 2.40e-13 4.84e-13 3.84e-13 3.05e-13 2.83e-13 3.00e-12 2.09e-13 5.50e-13 1.30e-13 6.78e-13 2.09e-13 4.39e-13 8.27

ingestion
Table G-4. (Continued)

Lifetime Average Daily Doses (LADDs)

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234

PCDDs PCDFs 

G-28

 Water 6.79e-13 1.87e-13 2.01e-13 1.85e-13 1.46e-13 1.14e-13 1.30e-12 1.84e-13 3.30e-13 8.70e-14 3.24e-13 1.00e-13 2.10e-13 3.06

ingestion

 Fish ingestion 2.07e-10 5.71e-11 6.14e-11 4.66e-11 3.69e-11 4.60e-12 9.35e-13 4.79e-11 1.40e-10 3.21e-11 8.19e-11 2.53e-11 5.30e-11 1.03

 Direct 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e

inhalation

SOH-fisher

 Soil dermal 2.84e-16 2.93e-16 6.92e-16 4.77e-16 3.08e-16 1.13e-15 1.33e-15 1.82e-16 2.40e-16 2.03e-16 3.39e-16 2.15e-16 3.19e-16 2.66

 Soil ingestion 1.51e-14 1.56e-14 3.69e-14 2.55e-14 1.64e-14 6.05e-14 7.11e-14 9.69e-15 1.28e-14 1.08e-14 1.81e-14 1.15e-14 1.70e-14 1.42

Veget. 6.06e-14 6.27e-14 1.26e-13 9.06e-14 5.87e-14 2.03e-13 2.47e-13 4.00e-14 4.85e-14 4.20e-14 6.46e-14 4.11e-14 6.08e-14 4.96

ingestion

 Water ingest 4.72e-14 4.88e-14 5.26e-14 4.36e-14 2.81e-14 8.12e-14 1.07e-13 3.51e-14 2.91e-14 2.82e-14 3.09e-14 1.96e-14 2.91e-14 1.84

 Fish 1.44e-11 1.49e-11 1.60e-11 1.10e-11 7.09e-12 3.29e-12 7.71e-14 9.16e-12 1.24e-11 1.04e-11 7.81e-12 4.96e-12 7.35e-12 6.17

ingestion

 Direct 2.38e-19 2.01e-19 4.20e-19 2.81e-19 1.84e-19 6.68e-19 7.76e-19 1.72e-19 1.69e-19 1.55e-19 2.05e-19 1.30e-19 1.99e-19 1.62

inhalation

LCH-farmer

 Soil dermal 1.50e-15 4.13e-14 9.73e-14 7.44e-14 5.87e-14 5.82e-14 5.94e-13 3.49e-14 1.00e-13 2.29e-14 1.30e-13 4.02e-14 8.44e-14 1.64

 Soil ingestion 1.44e-14 3.96e-13 9.35e-13 7.14e-13 5.64e-13 5.59e-13 5.71e-12 3.35e-13 9.62e-13 2.20e-13 1.25e-12 3.86e-13 8.11e-13 1.57

Veget. 1.77e-13 4.88e-12 1.05e-11 8.41e-12 6.49e-12 6.17e-12 6.45e-11 4.20e-12 1.14e-11 2.66e-12 1.44e-11 4.45e-12 9.33e-12 1.75

ingestion

 Animal 3.12e-12 7.56e-11 9.71e-11 6.97e-11 7.44e-11 2.29e-11 2.03e-10 2.02e-11 1.65e-10 5.21e-11 1.63e-10 4.95e-11 8.99e-11 1.21

ingestion

 Water ingest 4.48e-14 1.24e-12 1.33e-12 1.22e-12 9.65e-13 7.52e-13 8.57e-12 1.21e-12 2.18e-12 5.73e-13 2.14e-12 6.61e-13 1.39e-12 2.04
Table G-4. (Continued)

Lifetime Average Daily Doses (LADDs)

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234

PCDDs PCDFs 

G-29

 Direct 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e

inhalation

LOH-farmer

 Soil dermal 2.86e-14 2.97e-14 6.99e-14 4.81e-14 3.11e-14 1.14e-13 1.35e-13 1.82e-14 2.42e-14 2.05e-14 3.43e-14 2.17e-14 3.21e-14 2.69

 Soil ingestion 2.75e-13 2.85e-13 6.71e-13 4.62e-13 2.99e-13 1.10e-12 1.29e-12 1.75e-13 2.32e-13 1.97e-13 3.29e-13 2.08e-13 3.08e-13 2.58

Veget. 3.38e-12 3.51e-12 7.51e-12 5.44e-12 3.44e-12 1.21e-11 1.46e-11 2.19e-12 2.76e-12 2.38e-12 3.79e-12 2.40e-12 3.55e-12 2.88

ingestion

 Animal 5.97e-11 5.44e-11 6.97e-11 4.51e-11 3.94e-11 4.49e-11 4.61e-11 1.05e-11 4.00e-11 4.65e-11 4.29e-11 2.67e-11 3.42e-11 1.98

ingestion

 Water ingest 8.57e-13 8.88e-13 9.56e-13 7.91e-13 5.11e-13 1.48e-12 1.95e-12 6.34e-13 5.27e-13 5.12e-13 5.63e-13 3.57e-13 5.28e-13 3.34

 Direct 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e

inhalation

SCH-farmer

 Soil dermal 2.27e-14 6.24e-15 1.47e-14 1.12e-14 8.88e-15 8.81e-15 8.98e-14 5.28e-15 1.51e-14 3.48e-15 1.97e-14 6.08e-15 1.28e-14 2.46

 Soil ingestion 2.18e-13 6.00e-14 1.41e-13 1.08e-13 8.53e-14 8.46e-14 8.62e-13 5.07e-14 1.45e-13 3.35e-14 1.89e-13 5.84e-14 1.23e-13 2.36

Veget. 2.68e-12 7.38e-13 1.58e-12 1.27e-12 9.81e-13 9.34e-13 9.74e-12 6.35e-13 1.73e-12 4.04e-13 2.18e-12 6.72e-13 1.41e-12 2.64

ingestion

 Animal 4.73e-11 1.14e-11 1.47e-11 1.05e-11 1.13e-11 3.46e-12 3.07e-11 3.05e-12 2.50e-11 7.91e-12 2.47e-11 7.48e-12 1.36e-11 1.81

ingestion

 Water ingest 6.79e-13 1.87e-13 2.01e-13 1.85e-13 1.46e-13 1.14e-13 1.30e-12 1.84e-13 3.30e-13 8.70e-14 3.24e-13 1.00e-13 2.10e-13 3.06

 Direct 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e+00 0.0e

inhalation

SOH-farmer

 Soil dermal 1.58e-15 1.63e-15 3.84e-15 2.65e-15 1.71e-15 6.30e-15 7.40e-15 1.01e-15 1.33e-15 1.13e-15 1.88e-15 1.20e-15 1.77e-15 1.48

 Soil ingestion 1.51e-14 1.56e-14 3.69e-14 2.55e-14 1.64e-14 6.05e-14 7.11e-14 9.69e-15 1.28e-14 1.08e-14 1.81e-14 1.15e-14 1.70e-14 1.42
Table G-4. (Continued)

Lifetime Average Daily Doses (LADDs)

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234

PCDDs PCDFs 

G-30

Veget. 1.86e-13 1.93e-13 4.13e-13 3.00e-13 1.89e-13 6.67e-13 8.03e-13 1.21e-13 1.52e-13 1.31e-13 2.08e-13 1.32e-13 1.96e-13 1.58

ingestion

 Animal 3.29e-12 2.98e-12 3.83e-12 2.49e-12 2.16e-12 2.47e-12 2.53e-12 5.83e-13 2.20e-12 2.56e-12 2.36e-12 1.47e-12 1.88e-12 1.09

ingestion

 Water ingest 4.72e-14 4.88e-14 5.26e-14 4.36e-14 2.81e-14 8.12e-14 1.07e-13 3.51e-14 2.91e-14 2.82e-14 3.09e-14 1.96e-14 2.91e-14 1.84

 Direct 2.38e-19 2.01e-19 4.20e-19 2.81e-19 1.84e-19 6.68e-19 7.76e-19 1.72e-19 1.69e-19 1.55e-19 2.05e-19 1.30e-19 1.99e-19 1.62

inhalation

LCC-resident

 Soil dermal 1.03e-17 2.69e-16 6.12e-16 4.64e-16 3.67e-16 3.63e-16 3.69e-15 2.48e-16 6.57e-16 1.54e-16 8.17e-16 2.53e-16 5.35e-16 1.03

 Soil ingestion 5.50e-16 1.43e-14 3.27e-14 2.47e-14 1.96e-14 1.94e-14 1.97e-13 1.32e-14 3.50e-14 8.22e-15 4.36e-14 1.35e-14 2.85e-14 5.49

Veget. 1.38e-15 4.51e-14 7.66e-14 4.33e-14 3.97e-14 3.75e-14 2.96e-13 3.26e-14 8.02e-14 2.13e-14 7.70e-14 2.38e-14 5.58e-14 9.83

ingestion

 Water ingest 1.67e-15 4.36e-14 4.64e-14 4.21e-14 3.33e-14 2.60e-14 2.95e-13 4.64e-14 7.85e-14 2.10e-14 7.41e-14 2.29e-14 4.85e-14 7.11

 Direct 1.42e-17 3.05e-16 5.99e-16 4.34e-16 3.50e-16 3.40e-16 3.40e-15 3.76e-16 7.69e-16 1.96e-16 7.93e-16 2.45e-16 5.41e-16 1.01

inhalation

LOC-resident

 Soil dermal 1.85e-16 1.81e-16 4.12e-16 2.82e-16 1.83e-16 6.70e-16 7.87e-16 1.23e-16 1.49e-16 1.29e-16 2.02e-16 1.28e-16 1.91e-16 1.58

 Soil ingestion 9.89e-15 9.64e-15 2.20e-14 1.50e-14 9.74e-15 3.57e-14 4.20e-14 6.56e-15 7.92e-15 6.88e-15 1.08e-14 6.82e-15 1.02e-14 8.45

Veget. 3.18e-14 3.95e-14 6.30e-14 2.95e-14 2.30e-14 8.06e-14 6.71e-14 2.08e-14 2.21e-14 2.28e-14 2.12e-14 1.34e-14 2.30e-14 1.70

ingestion

 Water ingest 3.01e-14 2.93e-14 3.13e-14 2.55e-14 1.66e-14 4.79e-14 6.28e-14 2.31e-14 1.77e-14 1.76e-14 1.83e-14 1.16e-14 1.73e-14 1.09

 Direct 4.39e-16 3.16e-16 5.77e-16 3.70e-16 2.46e-16 8.80e-16 1.01e-15 3.41e-16 2.71e-16 2.68e-16 2.79e-16 1.77e-16 2.80e-16 2.22

inhalation

SCC-resident

 Soil dermal 1.76e-16 4.97e-17 1.19e-16 9.10e-17 7.19e-17 7.14e-17 7.29e-16 4.04e-17 1.20e-16 2.74e-17 1.59e-16 4.92e-17 1.03e-16 1.99
Table G-4. (Continued)

Lifetime Average Daily Doses (LADDs)

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234

PCDDs PCDFs 

G-31

 Soil ingestion 9.40e-15 2.65e-15 6.34e-15 4.85e-15 3.83e-15 3.81e-15 3.89e-14 2.16e-15 6.40e-15 1.46e-15 8.50e-15 2.62e-15 5.49e-15 1.06

Veget. 5.99e-14 2.42e-14 3.64e-14 1.39e-14 1.55e-14 1.47e-14 6.66e-14 1.32e-14 3.35e-14 9.92e-15 2.40e-14 7.39e-15 2.05e-14 3.17

ingestion

 Water ingest 2.87e-14 8.07e-15 9.02e-15 8.26e-15 6.52e-15 5.11e-15 5.82e-14 7.61e-15 1.44e-14 3.74e-15 1.45e-14 4.46e-15 9.34e-15 1.37

 Direct 1.30e-15 2.78e-16 5.45e-16 3.94e-16 3.18e-16 3.09e-16 3.08e-15 3.42e-16 6.99e-16 1.79e-16 7.22e-16 2.23e-16 4.92e-16 9.11

inhalation

SOC-resident

 Soil dermal 1.82e-17 2.15e-17 5.40e-17 3.78e-17 2.42e-17 8.97e-17 1.06e-16 1.05e-17 1.74e-17 1.40e-17 2.65e-17 1.68e-17 2.46e-17 2.07

 Soil ingestion 9.72e-16 1.15e-15 2.88e-15 2.01e-15 1.29e-15 4.78e-15 5.65e-15 5.60e-16 9.26e-16 7.44e-16 1.41e-15 8.98e-16 1.31e-15 1.11

Veget. 8.96e-15 1.31e-14 1.91e-14 6.60e-15 5.97e-15 2.11e-14 1.13e-14 5.57e-15 6.08e-15 6.80e-15 4.60e-15 2.92e-15 5.72e-15 3.82

ingestion

 Water ingest 2.98e-15 3.49e-15 4.10e-15 3.43e-15 2.20e-15 6.42e-15 8.47e-15 1.99e-15 2.08e-15 1.91e-15 2.40e-15 1.53e-15 2.23e-15 1.43

 Direct 2.05e-16 1.46e-16 2.65e-16 1.70e-16 1.13e-16 4.04e-16 4.61e-16 1.60e-16 1.26e-16 1.25e-16 1.28e-16 8.13e-17 1.29e-16 1.02

inhalation

LCC-child

 Soil dermal 2.69e-17 7.02e-16 1.60e-15 1.21e-15 9.60e-16 9.49e-16 9.65e-15 6.48e-16 1.72e-15 4.03e-16 2.14e-15 6.60e-16 1.40e-15 2.69

 Soil ingestion 9.06e-16 2.36e-14 5.38e-14 4.07e-14 3.23e-14 3.19e-14 3.24e-13 2.18e-14 5.77e-14 1.35e-14 7.18e-14 2.22e-14 4.70e-14 9.04

Veget. 4.17e-16 1.45e-14 2.26e-14 1.12e-14 1.10e-14 1.03e-14 6.97e-14 9.76e-15 2.34e-14 6.49e-15 2.00e-14 6.17e-15 1.53e-14 2.58

ingestion

 Water ingest 6.90e-16 1.80e-14 1.92e-14 1.74e-14 1.37e-14 1.07e-14 1.22e-13 1.92e-14 3.24e-14 8.66e-15 3.06e-14 9.46e-15 2.00e-14 2.93

 Direct 9.37e-18 2.01e-16 3.95e-16 2.86e-16 2.31e-16 2.24e-16 2.24e-15 2.48e-16 5.07e-16 1.29e-16 5.22e-16 1.61e-16 3.57e-16 6.63

inhalation

LOC-child

 Soil dermal 4.85e-16 4.72e-16 1.08e-15 7.36e-16 4.77e-16 1.75e-15 2.06e-15 3.22e-16 3.88e-16 3.37e-16 5.28e-16 3.34e-16 4.98e-16 4.14

 Soil ingestion 1.63e-14 1.59e-14 3.62e-14 2.47e-14 1.60e-14 5.88e-14 6.91e-14 1.08e-14 1.30e-14 1.13e-14 1.77e-14 1.12e-14 1.67e-14 1.39
Table G-4. (Continued)

Lifetime Average Daily Doses (LADDs)

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234

PCDDs PCDFs 

G-32

Veget. 1.02e-14 1.33e-14 1.95e-14 7.84e-15 6.64e-15 2.29e-14 1.58e-14 6.64e-15 6.78e-15 7.35e-15 5.65e-15 3.58e-15 6.54e-15 4.60

ingestion

 Water ingest 1.24e-14 1.21e-14 1.29e-14 1.05e-14 6.84e-15 1.98e-14 2.59e-14 9.52e-15 7.32e-15 7.25e-15 7.56e-15 4.78e-15 7.14e-15 4.52

 Direct 2.90e-16 2.08e-16 3.80e-16 2.44e-16 1.62e-16 5.80e-16 6.64e-16 2.25e-16 1.79e-16 1.76e-16 1.84e-16 1.16e-16 1.85e-16 1.46

inhalation

SCC-child

 Soil dermal 4.60e-16 1.30e-16 3.11e-16 2.38e-16 1.88e-16 1.87e-16 1.90e-15 1.06e-16 3.14e-16 7.16e-17 4.17e-16 1.28e-16 2.69e-16 5.19

 Soil Iigest. 1.55e-14 4.37e-15 1.04e-14 8.00e-15 6.31e-15 6.27e-15 6.40e-14 3.55e-15 1.05e-14 2.41e-15 1.40e-14 4.32e-15 9.04e-15 1.75

Veget. 2.15e-14 8.97e-15 1.28e-14 4.18e-15 5.13e-15 4.84e-15 1.56e-14 4.72e-15 1.17e-14 3.58e-15 7.20e-15 2.22e-15 6.68e-15 9.71

ingestion

 Water ingest 1.18e-14 3.33e-15 3.72e-15 3.41e-15 2.69e-15 2.11e-15 2.40e-14 3.14e-15 5.93e-15 1.54e-15 5.97e-15 1.84e-15 3.85e-15 5.67

 Direct 8.54e-16 1.83e-16 3.59e-16 2.60e-16 2.10e-16 2.04e-16 2.03e-15 2.26e-16 4.61e-16 1.18e-16 4.76e-16 1.47e-16 3.24e-16 6.00

inhalation

SOC-child

 Soil dermal 4.76e-17 5.62e-17 1.41e-16 9.87e-17 6.32e-17 2.34e-16 2.77e-16 2.74e-17 4.54e-17 3.65e-17 6.92e-17 4.40e-17 6.42e-17 5.42

 Soil ingestion 1.60e-15 1.89e-15 4.74e-15 3.32e-15 2.13e-15 7.88e-15 9.31e-15 9.22e-16 1.53e-15 1.23e-15 2.33e-15 1.48e-15 2.16e-15 1.82

Veget. 3.29e-15 4.86e-15 6.70e-15 1.98e-15 1.96e-15 6.89e-15 2.63e-15 2.06e-15 2.13e-15 2.49e-15 1.37e-15 8.73e-16 1.86e-15 1.17

ingestion

 Water ingest 1.23e-15 1.44e-15 1.69e-15 1.41e-15 9.06e-16 2.65e-15 3.49e-15 8.20e-16 8.58e-16 7.88e-16 9.92e-16 6.30e-16 9.21e-16 5.92

 Direct 1.35e-16 9.61e-17 1.75e-16 1.12e-16 7.43e-17 2.66e-16 3.04e-16 1.05e-16 8.31e-17 8.20e-17 8.43e-17 5.35e-17 8.52e-17 6.71

inhalation
G-33

Appendix G-5

Dioxin Congener Specific Cancer Risk Results for Each 

Scenario and Pathway
xxx
G-35

Table G-5. Dioxin Congener Specific Risk Results for Each Scenario and Pat

Cancer Risks

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 1234678 

PCDDs PCDFs 

LCH-fisher

Soil dermal 4e-11 6e-10 3e-10 2e-10 2e-10 2e-11 2e-11 1e-10 1e-09 3e-11 4e-10 1e-10 2e-10 5e-10 6e-11 

Soil ingestion 2e-09 3e-08 1e-08 1e-08 9e-09 9e-10 9e-10 5e-09 8e-08 2e-09 2e-08 6e-09 1e-08 2e-08 3e-09 

Vegetable

ingestion 9e-09 1e-07 5e-08 4e-08 3e-08 3e-09 3e-09 2e-08 3e-07 7e-09 7e-08 2e-08 5e-08 9e-08 1e-08 

Water

ingestion 7e-09 1e-07 2e-08 2e-08 2e-08 1e-09 1e-09 2e-08 2e-07 4e-09 3e-08 1e-08 2e-08 3e-08 4e-09 

Fish ingestion 2e-06 3e-05 6e-06 5e-06 4e-06 5e-08 1e-09 5e-06 7e-05 2e-06 8e-06 3e-06 5e-06 1e-05 2e-07 

Direct

inhalation 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 

LOH-fisher

Soil dermal 8e-10 4e-10 2e-10 1e-10 9e-11 3e-11 4e-12 5e-11 3e-10 3e-11 1e-10 6e-11 9e-11 8e-11 2e-11 

Soil ingestion 4e-08 2e-08 1e-08 7e-09 5e-09 2e-09 2e-10 3e-09 2e-08 2e-09 5e-09 3e-09 5e-09 4e-09 8e-10 

Vegetable

ingestion 2e-07 9e-08 4e-08 3e-08 2e-08 6e-09 7e-10 1e-08 7e-08 6e-09 2e-08 1e-08 2e-08 1e-08 3e-09 

Water

ingestion 1e-07 7e-08 1e-08 1e-08 8e-09 2e-09 3e-10 1e-08 4e-08 4e-09 9e-09 6e-09 8e-09 5e-09 1e-09 

Fish ingestion 4e-05 2e-05 5e-06 3e-06 2e-06 9e-08 2e-10 3e-06 2e-05 1e-06 2e-06 1e-06 2e-06 2e-06 4e-08 

Direct

inhalation 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 

SCH-fisher

Soil dermal 6e-10 9e-11 4e-11 3e-11 2e-11 2e-12 3e-12 1e-11 2e-10 5e-12 6e-11 2e-11 4e-11 7e-11 8e-12 

Soil ingestion 3e-08 5e-09 2e-09 2e-09 1e-09 1e-10 1e-10 8e-10 1e-08 3e-10 3e-09 9e-10 2e-09 4e-09 5e-10 

Vegetable

ingestion 1e-07 2e-08 8e-09 6e-09 5e-09 4e-10 5e-10 3e-09 4e-08 1e-09 1e-08 3e-09 7e-09 1e-08 2e-09 

Water

ingestion 1e-07 1e-08 3e-09 3e-09 2e-09 2e-10 2e-10 3e-09 3e-08 7e-10 5e-09 2e-09 3e-09 5e-09 6e-10 

Fish ingestion 3e-05 4e-06 1e-06 7e-07 6e-07 7e-09 1e-10 7e-07 1e-05 3e-07 1e-06 4e-07 8e-07 2e-06 2e-08 

Direct

inhalation 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 

SOH-fisher

Soil dermal 4e-11 2e-11 1e-11 7e-12 5e-12 2e-12 2e-13 3e-12 2e-11 2e-12 5e-12 3e-12 5e-12 4e-12 8e-13 

Soil ingestion 2e-09 1e-09 6e-10 4e-10 3e-10 9e-11 1e-11 2e-10 1e-09 8e-11 3e-10 2e-10 3e-10 2e-10 4e-11 

Vegetable

ingestion 9e-09 5e-09 2e-09 1e-09 9e-10 3e-10 4e-11 6e-10 4e-09 3e-10 1e-09 6e-10 9e-10 8e-10 2e-10 
Table G-5. (Continued)

Cancer Risks

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 1234678 

PCDDs PCDFs 

G-36

Water

ingestion 7e-09 4e-09 8e-10 7e-10 4e-10 1e-10 2e-11 5e-10 2e-09 2e-10 5e-10 3e-10 5e-10 3e-10 6e-11 

Fish ingestion 2e-06 1e-06 2e-07 2e-07 1e-07 5e-09 1e-11 1e-07 1e-06 8e-08 1e-07 8e-08 1e-07 1e-07 2e-09 

Direct

inhalation 4e-14 2e-14 7e-15 4e-15 3e-15 1e-15 1e-16 3e-15 1e-14 1e-15 3e-15 2e-15 3e-15 3e-15 5e-16 

LCH-farmer

Soil dermal 2e-10 3e-09 2e-09 1e-09 9e-10 9e-11 9e-11 5e-10 8e-09 2e-10 2e-09 6e-10 1e-09 3e-09 3e-10 

Soil ingestion 2e-09 3e-08 1e-08 1e-08 9e-09 9e-10 9e-10 5e-09 8e-08 2e-09 2e-08 6e-09 1e-08 2e-08 3e-09 

Vegetable

ingestion 3e-08 4e-07 2e-07 1e-07 1e-07 1e-08 1e-08 7e-08 9e-07 2e-08 2e-07 7e-08 1e-07 3e-07 3e-08 

Animal

ingestion 5e-07 6e-06 2e-06 1e-06 1e-06 4e-08 3e-08 3e-07 1e-05 4e-07 3e-06 8e-07 1e-06 2e-06 1e-07 

Water

ingestion 7e-09 1e-07 2e-08 2e-08 2e-08 1e-09 1e-09 2e-08 2e-07 4e-09 3e-08 1e-08 2e-08 3e-08 4e-09 

Direct

inhalation 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 

LOH-farmer

Soil dermal 4e-09 2e-09 1e-09 8e-10 5e-10 2e-10 2e-11 3e-10 2e-09 2e-10 5e-10 3e-10 5e-10 4e-10 8e-11 

Soil ingestion 4e-08 2e-08 1e-08 7e-09 5e-09 2e-09 2e-10 3e-09 2e-08 2e-09 5e-09 3e-09 5e-09 4e-09 8e-10 

Vegetable

ingestion 5e-07 3e-07 1e-07 8e-08 5e-08 2e-08 2e-09 3e-08 2e-07 2e-08 6e-08 4e-08 6e-08 4e-08 9e-09 

Animal

ingestion 9e-06 4e-06 1e-06 7e-07 6e-07 7e-08 7e-09 2e-07 3e-06 4e-07 7e-07 4e-07 5e-07 3e-07 3e-08 

Water

ingestion 1e-07 7e-08 1e-08 1e-08 8e-09 2e-09 3e-10 1e-08 4e-08 4e-09 9e-09 6e-09 8e-09 5e-09 1e-09 

Direct

inhalation 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 

SCH-farmer

Soil dermal 4e-09 5e-10 2e-10 2e-10 1e-10 1e-11 1e-11 8e-11 1e-09 3e-11 3e-10 9e-11 2e-10 4e-10 5e-11 

Soil ingestion 3e-08 5e-09 2e-09 2e-09 1e-09 1e-10 1e-10 8e-10 1e-08 3e-10 3e-09 9e-10 2e-09 4e-09 5e-10 

Vegetable

ingestion 4e-07 6e-08 2e-08 2e-08 2e-08 1e-09 2e-09 1e-08 1e-07 3e-09 3e-08 1e-08 2e-08 4e-08 5e-09 

Animal

ingestion 7e-06 9e-07 2e-07 2e-07 2e-07 5e-09 5e-09 5e-08 2e-06 6e-08 4e-07 1e-07 2e-07 3e-07 1e-08 

Water

ingestion 1e-07 1e-08 3e-09 3e-09 2e-09 2e-10 2e-10 3e-09 3e-08 7e-10 5e-09 2e-09 3e-09 5e-09 6e-10 
Table G-5. (Continued)

Cancer Risks

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 1234678 

PCDDs PCDFs 

G-37

Direct

inhalation 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 0e+00 

SOH-farmer

Soil dermal 2e-10 1e-10 6e-11 4e-11 3e-11 1e-11 1e-12 2e-11 1e-10 9e-12 3e-11 2e-11 3e-11 2e-11 5e-12 

Soil ingestion 2e-09 1e-09 6e-10 4e-10 3e-10 9e-11 1e-11 2e-10 1e-09 8e-11 3e-10 2e-10 3e-10 2e-10 4e-11 

Vegetable

ingestion 3e-08 2e-08 6e-09 5e-09 3e-09 1e-09 1e-10 2e-09 1e-08 1e-09 3e-09 2e-09 3e-09 2e-09 5e-10 

Animal

ingestion 5e-07 2e-07 6e-08 4e-08 3e-08 4e-09 4e-10 9e-09 2e-07 2e-08 4e-08 2e-08 3e-08 2e-08 1e-09 

Water

ingestion 7e-09 4e-09 8e-10 7e-10 4e-10 1e-10 2e-11 5e-10 2e-09 2e-10 5e-10 3e-10 5e-10 3e-10 6e-11 

Direct

inhalation 4e-14 2e-14 7e-15 4e-15 3e-15 1e-15 1e-16 3e-15 1e-14 1e-15 3e-15 2e-15 3e-15 3e-15 5e-16 

LCC-resident

Soil dermal 2e-12 2e-11 1e-11 7e-12 6e-12 6e-13 6e-13 4e-12 5e-11 1e-12 1e-11 4e-12 8e-12 2e-11 2e-12 

Soil ingestion 9e-11 1e-09 5e-10 4e-10 3e-10 3e-11 3e-11 2e-10 3e-09 6e-11 7e-10 2e-10 4e-10 9e-10 1e-10 

Vegetable

ingestion 2e-10 4e-09 1e-09 7e-10 6e-10 6e-11 5e-11 5e-10 6e-09 2e-10 1e-09 4e-10 9e-10 2e-09 1e-09 

Water

ingestion 3e-10 3e-09 7e-10 7e-10 5e-10 4e-11 5e-11 7e-10 6e-09 2e-10 1e-09 4e-10 8e-10 1e-09 1e-10 

Direct

inhalation 2e-12 2e-11 9e-12 7e-12 5e-12 5e-13 5e-13 6e-12 6e-11 2e-12 1e-11 4e-12 8e-12 2e-11 2e-12 

LOC-resident

Soil dermal 3e-11 1e-11 6e-12 4e-12 3e-12 1e-12 1e-13 2e-12 1e-11 1e-12 3e-12 2e-12 3e-12 2e-12 5e-13 

Soil ingestion 2e-09 8e-10 3e-10 2e-10 2e-10 6e-11 7e-12 1e-10 6e-10 5e-11 2e-10 1e-10 2e-10 1e-10 3e-11 

Vegetable

ingestion 5e-09 3e-09 1e-09 5e-10 4e-10 1e-10 1e-11 3e-10 2e-09 2e-10 3e-10 2e-10 4e-10 3e-10 3e-10 

Water

ingestion 5e-09 2e-09 5e-10 4e-10 3e-10 7e-11 1e-11 4e-10 1e-09 1e-10 3e-10 2e-10 3e-10 2e-10 4e-11 

Direct

inhalation 7e-11 2e-11 9e-12 6e-12 4e-12 1e-12 2e-13 5e-12 2e-11 2e-12 4e-12 3e-12 4e-12 3e-12 7e-13 

SCC-resident

Soil dermal 3e-11 4e-12 2e-12 1e-12 1e-12 1e-13 1e-13 6e-13 9e-12 2e-13 2e-12 8e-13 2e-12 3e-12 4e-13 

Soil ingestion 1e-09 2e-10 1e-10 8e-11 6e-11 6e-12 6e-12 3e-11 5e-10 1e-11 1e-10 4e-11 9e-11 2e-10 2e-11 

Vegetable

ingestion 9e-09 2e-09 6e-10 2e-10 2e-10 2e-11 1e-11 2e-10 3e-09 8e-11 4e-10 1e-10 3e-10 5e-10 8e-10 
Table G-5. (Continued)

Cancer Risks

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 1234678 

PCDDs PCDFs 

G-38

Water

ingestion 4e-09 6e-10 1e-10 1e-10 1e-10 8e-12 9e-12 1e-10 1e-09 3e-11 2e-10 7e-11 1e-10 2e-10 3e-11 

Direct

inhalation 2e-10 2e-11 8e-12 6e-12 5e-12 5e-13 5e-13 5e-12 5e-11 1e-12 1e-11 3e-12 8e-12 1e-11 2e-12 

SOC-resident

Soil dermal 3e-12 2e-12 8e-13 6e-13 4e-13 1e-13 2e-14 2e-13 1e-12 1e-13 4e-13 3e-13 4e-13 3e-13 7e-14 

Soil ingestion 2e-10 9e-11 4e-11 3e-11 2e-11 7e-12 9e-13 9e-12 7e-11 6e-12 2e-11 1e-11 2e-11 2e-11 3e-12 

Vegetable

ingestion 1e-09 1e-09 3e-10 1e-10 9e-11 3e-11 2e-12 9e-11 5e-10 5e-11 7e-11 5e-11 9e-11 6e-11 2e-10 

Water

ingestion 5e-10 3e-10 6e-11 5e-11 3e-11 1e-11 1e-12 3e-11 2e-10 1e-11 4e-11 2e-11 3e-11 2e-11 5e-12 

Direct

inhalation 3e-11 1e-11 4e-12 3e-12 2e-12 6e-13 7e-14 2e-12 1e-11 1e-12 2e-12 1e-12 2e-12 2e-12 3e-13 

LCC-child

Soil dermal 4e-12 5e-11 2e-11 2e-11 1e-11 1e-12 2e-12 1e-11 1e-10 3e-12 3e-11 1e-11 2e-11 4e-11 5e-12 

Soil ingestion 1e-10 2e-09 8e-10 6e-10 5e-10 5e-11 5e-11 3e-10 5e-09 1e-10 1e-09 3e-10 7e-10 1e-09 2e-10 

Vegetable

ingestion 7e-11 1e-09 4e-10 2e-10 2e-10 2e-11 1e-11 2e-10 2e-09 5e-11 3e-10 1e-10 2e-10 4e-10 4e-10 

Water

ingestion 1e-10 1e-09 3e-10 3e-10 2e-10 2e-11 2e-11 3e-10 3e-09 7e-11 5e-10 1e-10 3e-10 5e-10 6e-11 

Direct

inhalation 1e-12 2e-11 6e-12 4e-12 4e-12 3e-13 3e-13 4e-12 4e-11 1e-12 8e-12 3e-12 6e-12 1e-11 1e-12 

LOC-child

Soil dermal 8e-11 4e-11 2e-11 1e-11 7e-12 3e-12 3e-13 5e-12 3e-11 3e-12 8e-12 5e-12 8e-12 6e-12 1e-12 

Soil ingestion 3e-09 1e-09 6e-10 4e-10 3e-10 9e-11 1e-11 2e-10 1e-09 9e-11 3e-10 2e-10 3e-10 2e-10 4e-11 

Vegetable

ingestion 2e-09 1e-09 3e-10 1e-10 1e-10 4e-11 2e-12 1e-10 5e-10 6e-11 9e-11 6e-11 1e-10 7e-11 1e-10 

Water

ingestion 2e-09 9e-10 2e-10 2e-10 1e-10 3e-11 4e-12 1e-10 6e-10 6e-11 1e-10 7e-11 1e-10 7e-11 2e-11 

Direct

inhalation 5e-11 2e-11 6e-12 4e-12 3e-12 9e-13 1e-13 4e-12 1e-11 1e-12 3e-12 2e-12 3e-12 2e-12 4e-13 

SCC-child

Soil dermal 7e-11 1e-11 5e-12 4e-12 3e-12 3e-13 3e-13 2e-12 2e-11 6e-13 6e-12 2e-12 4e-12 8e-12 1e-12 

Soil ingestion 2e-09 3e-10 2e-10 1e-10 1e-10 1e-11 1e-11 6e-11 8e-10 2e-11 2e-10 7e-11 1e-10 3e-10 3e-11 

Vegetable

ingestion 3e-09 7e-10 2e-10 7e-11 8e-11 8e-12 2e-12 7e-11 9e-10 3e-11 1e-10 3e-11 1e-10 2e-10 3e-10 
Table G-5. (Continued)

Cancer Risks

Congener 2378 12378 123478 123789 123678 1234678 Octa 2378 23478 12378 123478 123678 123789 234678 1234678 

PCDDs PCDFs 

G-39

Water

ingestion 2e-09 3e-10 6e-11 5e-11 4e-11 3e-12 4e-12 5e-11 5e-10 1e-11 9e-11 3e-11 6e-11 9e-11 1e-11 

Direct

inhalation 1e-10 1e-11 6e-12 4e-12 3e-12 3e-13 3e-13 4e-12 4e-11 9e-13 7e-12 2e-12 5e-12 9e-12 1e-12 

SOC-child

Soil dermal 7e-12 4e-12 2e-12 2e-12 1e-12 4e-13 4e-14 4e-13 4e-12 3e-13 1e-12 7e-13 1e-12 8e-13 2e-13 

Soil ingestion 2e-10 1e-10 7e-11 5e-11 3e-11 1e-11 1e-12 1e-11 1e-10 1e-11 4e-11 2e-11 3e-11 3e-11 6e-12 

Vegetable

ingestion 5e-10 4e-10 1e-10 3e-11 3e-11 1e-11 4e-13 3e-11 2e-10 2e-11 2e-11 1e-11 3e-11 2e-11 6e-11 

Water

ingestion 2e-10 1e-10 3e-11 2e-11 1e-11 4e-12 5e-13 1e-11 7e-11 6e-12 2e-11 1e-11 1e-11 9e-12 2e-12 

Direct

inhalation 2e-12 7e-12 3e-12 2e-12 1e-12 4e-13 5e-14 2e-12 6e-12 6e-13 1e-12 8e-13 1e-12 1e-12 2e-13 
xxx
G-41

Appendix G-6

Comparison of Dioxin Local-Scale ISCST3 and Long-Range RELMAP Modeling

Exposures and Risks
xxx
G-43

Table G-6. Comparison of Local-Scale ISCST3 and Long-Range

RELMAP Modeling Exposures and Risks for Dioxin

Total TEQ exposures Total TEQ RELMAP % of Total 

 (LADD in mg/kg-d) Risk (mean data)

LCH-fisher

 Soil Dermal 2.9e-14 5e-09 0.14%

 Soil Ingestion 1.6e-12 2e-07 0.14%

 Vegetation Ingestion 6.0e-12 9e-07 4.92%

 Water Ingestion 3.2e-12 5e-07 0.22%

 Fish Ingestion 9.9e-10 2e-04 0.22%

 Direct Inhalation 1.8e-15 3e-10 100.00%

LOH-fisher

 Soil Dermal 1.6e-14 2e-09 0.26%

 Soil Ingestion 8.4e-13 1e-07 0.26%

 Vegetation Ingestion 3.5e-12 5e-07 8.50%

 Water Ingestion 2.1e-12 3e-07 0.33%

 Fish Ingestion 6.5e-10 1e-04 0.33%

 Direct Inhalation 1.8e-15 3e-10 100.00%

SCH-fisher

 Soil Dermal 8.5e-15 1e-09 0.48%

 Soil Ingestion 4.5e-13 7e-08 0.48%

 Vegetation Ingestion 2.0e-12 3e-07 14.64%

 Water Ingestion 1.2e-12 2e-07 0.61%

 Fish Ingestion 3.6e-10 6e-05 0.60%

 Direct Inhalation 1.8e-15 3e-10 100.00%

SOH-fisher

 Soil Dermal 9.1e-16 1e-10 4.48%

 Soil Ingestion 4.8e-14 8e-09 4.48%

 Vegetation Ingestion 4.7e-13 7e-08 62.78%

 Water Ingestion 1.2e-13 2e-08 5.75%

 Fish Ingestion 3.8e-11 6e-06 5.68%

 Direct Inhalation 1.8e-15 3e-10 99.97%

LCH-farmer

 Soil Dermal 1.6e-13 3e-08 0.14%

 Soil Ingestion 1.6e-12 2e-07 0.14%

 Vegetation Ingestion 1.9e-11 3e-06 2.48%

 Animal Product Ingestion 2.0e-10 3e-05 0.85%

 Water Ingestion 3.2e-12 5e-07 0.22%

 Direct Inhalation 1.8e-15 3e-10 100.00%
Table G-6. (Continued)

Total TEQ exposures Total TEQ RELMAP % of Total 

 (LADD in mg/kg-d) Risk (mean data)

G-44

LOH-farmer

 Soil Dermal 8.7e-14 1e-08 0.26%

 Soil Ingestion 8.4e-13 1e-07 0.26%

 Vegetation Ingestion 1.0e-11 2e-06 4.42%

 Animal Product Ingestion 1.4e-10 2e-05 1.24%

 Water Ingestion 2.1e-12 3e-07 0.33%

 Direct Inhalation 1.8e-15 3e-10 100.00%

SCH-farmer

 Soil Dermal 4.7e-14 7e-09 0.48%

 Soil Ingestion 4.5e-13 7e-08 0.48%

 Vegetation Ingestion 5.9e-12 9e-07 7.89%

 Animal Product Ingestion 7.9e-11 1e-05 2.20%

 Water Ingestion 1.2e-12 2e-07 0.61%

 Direct Inhalation 1.8e-15 3e-10 100.00%

SOH-farmer

 Soil Dermal 5.0e-15 8e-10 4.48%

 Soil Ingestion 4.8e-14 8e-09 4.48%

 Vegetation Ingestion 1.0e-12 2e-07 45.66%

 Animal Product Ingestion 9.4e-12 1e-06 18.54%

 Water Ingestion 1.2e-13 2e-08 5.75%

 Direct Inhalation 1.8e-15 3e-10 99.97%

LCC-resident

 Soil Dermal 1.1e-15 2e-10 3.74%

 Soil Ingestion 5.8e-14 9e-09 3.74%

 Vegetation Ingestion 2.4e-13 4e-08 37.37%

 Water Ingestion 1.2e-13 2e-08 5.76%

 Direct Inhalation 2.9e-15 5e-10 61.32%

LOC-resident

 Soil Dermal 5.8e-16 9e-11 7.01%

 Soil Ingestion 3.1e-14 5e-09 7.01%

 Vegetation Ingestion 1.8e-13 3e-08 49.78%

 Water Ingestion 7.8e-14 1e-08 8.80%

 Direct Inhalation 2.8e-15 4e-10 63.60%

SCC-resident

 Soil Dermal 4.1e-16 6e-11 9.85%
Table G-6. (Continued)

Total TEQ exposures Total TEQ RELMAP % of Total 

 (LADD in mg/kg-d) Risk (mean data)

G-45

 Soil Ingestion 2.2e-14 3e-09 9.85%

 Vegetation Ingestion 2.2e-13 3e-08 39.83%

 Water Ingestion 5.6e-14 9e-09 12.19%

 Direct Inhalation 4.1e-15 6e-10 43.52%

SOC-resident

 Soil Dermal 1.0e-16 2e-11 39.65%

 Soil Ingestion 5.5e-15 9e-10 39.65%

 Vegetation Ingestion 1.1e-13 2e-08 77.27%

 Water Ingestion 1.5e-14 2e-09 46.34%

 Direct Inhalation 2.2e-15 3e-10 79.02%

LCC-child

 Soil Dermal 2.8e-15 4e-10 3.74%

 Soil Ingestion 9.5e-14 1e-08 3.74%

 Vegetation Ingestion 8.1e-14 1e-08 43.34%

 Water Ingestion 4.9e-14 8e-09 5.76%

 Direct Inhalation 1.9e-15 3e-10 61.32%

LOC-child

 Soil Dermal 1.5e-15 2e-10 7.01%

 Soil Ingestion 5.1e-14 8e-09 7.01%

 Vegetation Ingestion 6.3e-14 1e-08 55.25%

 Water Ingestion 3.2e-14 5e-09 8.80%

 Direct Inhalation 1.8e-15 3e-10 63.60%

SCC-child

 Soil Dermal 1.1e-15 2e-10 9.85%

 Soil Ingestion 3.6e-14 6e-09 9.85%

 Vegetation Ingestion 8.3e-14 1e-08 42.00%

 Water Ingestion 2.3e-14 4e-09 12.19%

 Direct Inhalation 2.7e-15 4e-10 43.52%

SOC-child

 Soil Dermal 2.7e-16 4e-11 39.65%

 Soil Ingestion 9.0e-15 1e-09 39.65%

 Vegetation Ingestion 4.4e-14 7e-09 78.82%

 Water Ingestion 6.1e-15 1e-09 46.34%

 Direct Inhalation 1.5e-15 2e-10 79.02%
G-46

G.7 REFERENCES FOR APPENDIX G

Brzuzy, L. P. and R. A. Hites. Estimating the atmospheric deposition

of polychlorinated dibenzo-p-dioxins and furans from soils. 

Environmental Science and Technology. Volume 29. 1995. Pp.2090-2098. 

1995.

Geraghty, J. J., D. W. Miller, F. Fander Leeden, and F. L. Troise. 

Water Atlas of the United States. Plate 21. 1973.

Lorber, Matthew. EPA Office of Research and Development. Personal

Communication. 1997.

Lorber, Matthew. Development of an Air-to-Leaf Phase Transfer Factor

of Dioxins and Furans. In: Organohalogen Compounds, Volume 24. 

(Dioxin '95: 15th International Symposium on Chlorinated Dioxins and

Related Compounds, Short Papers). Pages 179-186. 1995.

SAMSON. Solar and Meteorological Surface Observation Network. 1961 -

1990 data, Version 1. United States Department of Commerce, National

Climatic Data Center, Asheville, NC. 1993.

Stephens, R.D, M.X. Petreas, and D.G. Haward. Biotransfer and

Bioaccumulation of Dioxins and Furans from Soil: Chickens as a Model

for Grazing Animals. Science of the total Environment. Volume 175. 

1995. Pp. 253-273. 1995.

U.S. Environmental Protection Agency. Methodology for Assessing

Health Risks Associated with Indirect Exposure to Combustor Emissions.

Interim Final Report. EPA/600/6-90/003. Office of Health and

Environmental Assessment. Washington, DC. 1990.

U.S. EPA. Addendum to the Methodology for Assessing Health Risks

Associated with Indirect Exposure to Combustor Emissions. External

Review Draft. EPA/600/AP-93/003. Office of Health and Environmental

Assessment. Washington, DC. 1993.

U.S. Environmental Protection Agency. Estimating Exposure to DioxinLike Compounds, Vol. III: Site-Specific Assessment Procedures. 

External Review Draft. EPA/600/6-88/005Cc. Office of Research and

Development. Washington, DC. 1994a.

U.S. Environmental Protection Agency. Mercury Study Report to

Congress, Volume III: An Assessment of Exposure from Anthropogenic

Mercury Emissions in the United States. SAB Review Draft. EPA-452/R96-001c. Office of Air Quality Planning & Standards and Office of

Research and Development. 1996c.
G-47

U.S. Environmental Protection Agency. Study of Hazardous Air

Pollutant Emissions from Electric Utility Steam Generating Units.

Interim Final Report. EPA-453/R-96-013c. Office of Air Quality

Planning and Standards. Research Triangle Park, NC. 1996d.
mm
Appendix H - Literature Review of the Potential Impacts of Hydrogen

Chloride and Hydrogen Fluoride
x
H-1

H.1 OVERVIEW

The information presented in this appendix was collected to

expand the EPA's knowledge of the potential impacts of HCl and HF

emissions from utilities. The EPA is updating its current state of

knowledge of health impacts (including dose/response relationships);

atmospheric chemistry (e.g., half-life, impacts on the acid rain

phenomenon); potential human exposure through pathways other than

direct inhalation; and possible ecological harm. The EPA’s goal is to

understand the potential impacts from HCl and HF emissions to any and

all health and environmental areas. This appendix is not intended to

provide a detailed, comprehensive treatise on the above subject area;

rather, it is designed to provide general technical information that

will identify possible problem areas that may call for additional,

more detailed research.

Published evidence for potential impacts of HCl and HF was

evaluated from a wide variety of sources. Overall, there is extensive

information available on the toxicology of these two pollutants;

however, literature pertaining specifically to HF and HCl atmospheric

chemistry is relatively scarce, especially that pertaining to fine

particulate matter and acid rain. Literature on HCl and HF from

sources outside the United States and pertaining to emissions sources

other than utilities has also been evaluated. 

This appendix is organized so that the findings for HCl are

presented first, followed by the findings for HF. Within each

section, evidence from the literature for transport and transformation

through atmospheric, terrestrial, and aquatic processes is presented

first, followed by evidence for impacts on human health; vegetation;

and wild, domestic, and aquatic animals.

H.2 FINDINGS FOR HYDROGEN CHLORIDE

H.2.1 HCl Emissions and Formation

The information on nationwide utility HCl emissions was obtained

from Table 3-3 in Chapter 3 of this report (1990 estimate). Emissions

are reported to be 148,000 tons/year for coal and natural gas boilers

combined. Utility emissions are the most significant anthropogenic

source of atmospheric HCl. Other important sources are industrial coal

combustion, and solid waste combustion. 

Atmospheric HCl is emitted by both natural and anthropogenic

sources. For instance, anthropogenic sources contributing to measured

concentrations of HCl in an urban area of Switzerland were found to

include automobiles, heating units, and a garbage incinerator. 1

Wegner, et al. cite coal combustion and waste incineration as the main

anthropogenic HCl sources. Puxbaum, et al. cite coal combustion as 2

the primary source of HCl in central Europe. Other sources include 3

biomass burning and the photolysis of Cl-atom precursors such as HCl

and Cl , followed by hydrocarbon reactions. Natural sources of HCl 2 4

emissions include volcanic activity, marine plants/microorganisms,
H-2

land plant combustion-generated methyl chloride, and sea-salt

reactions. 5

Graedel, et al. estimated the global acid-equivalent fluxes and

reported on observations for the period 1977-1990. They predicted HCl 6

emissions growth to the year 2100. The dominant global source of

atmospheric HCl is believed to be marine production by direct

volatilization from deliquescent sea-salt aerosol that has been

acidified by the incorporation of HNO and/or H SO . Total global 3 24

emissions of HCl are estimated at 55 Tg Cl per year, in a year with

average volcanic activity. Acid-equivalent fluxes are calculated to

be 2.0 Teq H per year for SO (83 percent anthropogenic), 2.2 Teq H + + x

per year for NO (57 percent anthropogenic), and 1.6 Teq H per year x +

for HCl. However, because most of the HCl is thought to be generated

by acid-displacement reactions involving anthropogenically derived

precursors, much of this HCl does not correspond to a net production

of atmospheric acidity. Thus, the net influence of this acidity is

already accounted for in the sum of SO and NO . Because SO and NO xx xx

emissions as a proxy for HCl emissions decreased from 1975 to 1995 in

more developed countries, the 1 to 11 percent per year increase in HCl

concentrations observed during the 1977-1990 period is believed to be

from enhanced volatilization of sea salt. Interestingly, Graedel, et

al. predict that HCl emissions will grow from an estimated 55 Tg

chlorine per year in 1990 to 158 Tg chlorine per year in 2100, and

that acid-equivalent emissions will more than double in this period

due to development.

HCl emissions are believed to be the third largest source of

anthropogenic atmospheric acidity. In the United Kingdom, HCl

emissions sources are coal-fired boilers, waste incineration,

chlorinated hydrocarbons, automobile exhaust, glass-making, fuel oil

combustion, steel pickling acid, and regeneration. Coal burning, 5

however, is responsible for 93 percent of total HCl emissions in the

United Kingdom, with waste incineration emitting another 6 percent. 

All other sources combined emit the remaining 1 percent of HCl. HCl

contributes only 4 percent of the United Kingdom’s potential

atmospheric acidity, while 71 percent and 25 percent are attributed to

SO and NO , respectively. When evaluated at the scale of Western 2 x

Europe, the contribution by HCl to atmospheric potential acidity is

estimated at 2 percent.

The past, current, and future quality and availability of

emissions inventories for acid-related compounds was evaluated by

Graedel, et al. Information available on atmospheric fluxes of HCl 6

was determined to be of poor quality. Because of the poor

availability and quality of global emissions inventories of

atmospheric acid-related compounds determined by a survey in 1992, the

Global Emissions Inventory Activity (GEIA) was introduced under the

auspices of the International Global Atmospheric Chemistry Project. 

GEIA was developed with the goal of establishing a framework for the

development and evaluation of global emissions inventories, along with

the generation and publication of inventories for use by the global
H-3

science and policy communities. In the future, GEIA inventories are

expected to be a significant aid to the characterization of global

emissions of atmospheric species.

Information on ambient concentrations of HCl was relatively

scarce, particularly for the United States. This review includes

information on both urban and rural concentrations, as available. 

Kelly, et al. reviewed concentrations and transformations of

Hazardous Air Pollutants (HAPs). They noted that, as of August 1994,

eight HCl monitoring stations had been operated in the United States

and over 74 samples had been measured, ranging from none detected to 4 7

g/m .3

Wegner, et al. found tropospheric HCl concentrations to vary

substantially as a function of location and time, and to be strongly

correlated with CH O concentrations ® = 0.93, 0.90, and 0.95 for polar, 2

midlatitude maritime, and midlatitude continental, respectively). 2

Average HCl concentrations measured in a tropospheric column were 1.15

x 10 molecules per square centimeter. The concentrations of HCl 15

found in the three regions were as follows: continental maritime <

midlatitude maritime < polar maritime. The reason posited for the

highest HCl concentrations being observed in polar maritime air is

high reaction rates of non-methane hydrocarbons with Cl atoms,

yielding HCl as a reaction product. There was no correlation between

HF and HCl concentrations.

High intermittent concentrations of HCl observed in an urban

environment in Switzerland (up to 3.2 g/m ) were believed to originate 3

largely from an incinerator located approximately 3 km away from the

monitor. Hutchinson, et al., in a paper on HCl-induced stone 1

degradation, estimate typical atmospheric HCl concentrations ranges in

North America of 0.1-1.4 g/m in rural areas and 0.2-3 g/m in urban 3 3

areas. Puxbaum, et al. measured HCl and several other chemical 8

species at a rural site in northeast Austria. HCl concentrations 3

exhibited substantial variation between winter and summer, with

elevated values of 0.7 g/m found in winter. Annual average HCl 3

concentrations were 0.3 g/m .3

The ambient concentrations of HCl will be of interest as part of

the effort to achieve compliance with the new particulate matter

standards. If a significant portion of ambient HCl is in the fine

fraction, it could conceivably contribute to PM exceedances. The 2.5

remainder of this section addresses the topics of HCl formation and

formation by-products.

HCl can be formed several ways in the atmosphere. The burning of

coal can yield HCl as a combustion product, with the quantity of HCl

emitted in this manner being a function of coal composition, method of

combustion, and air pollution control methods. A study conducted by 9

the U.S. Bureau of Mines indicated that the majority of chlorine

contained in coal volatilizes to form HCl. Additional tests found 10

that only small amounts of chlorine remain in the combustion ash. 
H-4

Other studies reveal the processes of chlorine and HCl formation in

anthropogenic and natural systems.

Nonanthropogenic HCl is emitted by volcanoes, or it can be formed

from deliquescent sea salt in the marine environment by the following

process:4

 HNO + NaCl 3(g) (s) (g) 3(s) HCl + NaNO (1)

H SO + 2NaCl 2 4(s) (s) (g) 2 4(s) 2HCl + Na SO (2)

Thus, sulfuric acid and nitric acid in the atmosphere can react with

sea salt spray to generate hydrogen chloride. HCl in marine

environments can also be formed by indirect pathways, which generate

various chlorinated species such as ClNO or Cl . Once this happens, 2 2

photolysis of the chlorinated species will produce chlorine radicals,

which react as follows to produce HCl:

 Cl + RH HCl + R (3)

Rupert and Sigg investigated the interaction between fogwater and

aerosols, which can create or destroy HCl. The following reaction was 1

found to account for the transition of chlorine between the aerosol

and gaseous phase:

NH Cl 4 aerosol 3(g) (g)  NH + HCl (4)

In their study of fogwater and aerosols in an urban area of

Switzerland, Ruprecht and Sigg found empirical values of NH Cl to be 4

lower than theoretical values. One possible explanation is that

insufficient amounts of NH Cl were available to sustain equilibrium 4

concentrations. HCl was observed to dissolve in fogwater, causing

high aqueous concentrations, which were subsequently released in the

gas phase upon fog dissipation. It is thought that concentrations of

gaseous NH must have been too low to neutralize the HCl present, 3 

hindering the formation of NH Cl. A significant portion of atmospheric 4

Cl is present as fine aerosol (<2.4 m) under fog-free conditions. -

Fogwater acts as an ephemeral sink and possible reaction environment

for HCl and other soluble gases.

Another source of atmospheric HCl is anthropogenic chlorocarbons,

which can react principally with OH radicals to produce HCl, with a

reaction rate of approximately 0.5 percent per day. However, due to 5

this slow reaction rate, which permits widespread dispersion of the

HCl produced, it is believed that chlorocarbon-produced HCl emissions

contribute a negligible amount to atmospheric acidity.

In their review of atmospheric transformations of HAPs, Spicer,

et al. describe the identification of HCl reaction products (i.e.,

chloride salts) as qualitative in most cases, because few studies

reported mass balance information.11
H-5

Based on thermodynamic equilibrium calculations performed in the

U.S. Bureau of Mines study previously discussed, coal combustion can

generate small amounts of other gaseous chlorine compounds, including

Cl , HOCl, and NOCl. Analysis during laboratory simulation did not, 2 10

however, detect the presence of these compounds; the emissions were

virtually all HCl.

In summary, HCl can be formed during coal combustion, waste

incineration, the reaction and purification steps in the propylene

oxide manufacturing process, hydrocarbon chlorination and

dehydrochlorination, and combustion of the chlorinated hydrocarbons

found in some gasolines. Nonanthropogenic HCl is formed from

deliquescent sea salt in the marine environment or emitted by

volcanoes. HCl can also be created or destroyed through the

interaction between fogwater and aerosols. Reactions generating HCl

can produce the following by-products in the atmosphere: NaNO , Na SO , 3 24

hydrocarbon radicals, and NH .3

H.2.2 HCl Atmospheric Processes

In a review of the stability and persistence of atmospheric HCl,

two references were found that dealt directly with the atmospheric

lifetime of HCl. No information was reviewed on the stability and

persistence of HCl by-products. 

In their review paper on atmospheric transformation of HAPs,

Spicer, et al. estimate the atmospheric lifetime of HCl to be between

1 and 5 days. Lifetime estimates are defined as the time required 11

for a given HAP’s concentration to decrease to 1/e (37 percent) of its

original value, via atmospheric reaction or removal. Of the 178

chemicals on which lifetime estimates were obtained, 83 had lifetimes

of less than one day, 25 had lifetimes in the 1-5 day category, and 57

had lifetimes greater than 5 days. Thirteen of the chemicals had

conflicting estimates of lifetimes. Lifetime estimates were described

as relative rather than absolute estimates of HAP transformation

lifetimes, because of the varying information sources and calculation

methods. Wegner, et al. report a typical HCl tropospheric lifetime of

1 to 2 days under conditions allowing photochemistry.2

Based on the lifetime information found in this research, HCl

does not appear to be very persistent in the atmosphere. However, it

is possible that chemical lifetimes of 1 to 5 days may result in

utility emissions of HCl reaching acid rain-, or PM -sensitive 2.5

receptors. Future research could address this question.

Several sources were found on the atmospheric chemistry and

removal of HCl. Removal rate is an important factor when considering

HCl’s ability to be transported.

HCl is a highly reactive gas which is rapidly removed from the

atmosphere by most surfaces, particularly those that are moist. 5,8

HCl’s dry deposition rate is thought to be controlled by atmospheric

turbulence, rather than surface conditions. In general, HCl gas will
H-6

be removed from the atmosphere much faster than SO or NO and will be 2 2

deposited in close proximity to the emissions source. Because HCl is

highly soluble, washout is an efficient removal mechanism. Wet

deposition is also likely to deposit HCl close to the source. A study

by Patrinos, et al. found nearly all HCl to be wet-deposited within 15

km from the coal-burning power plant source.12

HCl has been found to be responsible for significantly enhancing

the acidity of cloud water. In the absence of clouds or rain, HCl

stack emissions are likely to be gaseous upon deposition. Chloride

ions can catalyze the oxidation of sulphite to sulphate in

concentrations commonly found in stack plumes. This can cause SO 2

suspended in chloride-containing water droplets to be oxidized more

rapidly than if in pure water. If this is the case, SO deposition 2

close to the source will be decreased, but rain and cloudwater acidity

will be increased, subject to further transport. However, because the

solubility of SO is pH dependent, if HCl lowers the pH below 2

approximately 3, chloride ion catalysis may be reduced. HCl’s

solubility is the reason it is more efficiently removed by rain than

either SO or NO . Thus the acidity in rainwater near a coal-fired 2 x

power plant may be predominately the result of HCl, rather than SO or 2

NO . x

More evidence of chloride ion catalyzation of SO was found in a 2

study by Clarke and Radojevic. This study was designed to provide

more applicable kinetic data concerning the oxidation of SO in fresh 2

and salt water. Clarke and Radojevic found that various chloride 13

salts significantly increased the rate of SO oxidation. The effects 2

increased with increasing salt concentrations. Other nonchloride

salts were also studied and did not significantly affect the reaction

rate. The levels at which the reaction rate was affected, however,

were greater than typical concentrations of chloride in cloud and

rainwater. Consequently, the main significance for HCl atmospheric

chemistry lies in reactions of marine or coastal aerosols and in HClenriched plumes at high humidity. In these environments, rates of

oxidation of SO in droplets will be elevated above the rates of pure 2

water by up to four orders of magnitude.

Tropospheric HCl can be generated by reactions between sea-salt

aerosol and atmospheric acids, volcanic eruptions, and the oxidation

of nonmethane hydrocarbons (NMHC). In these reactions, Cl is produced

by photolysis of species such as Cl , HOCl, BrCl, and ClNO , which are 2 2

all volatilized from sea-salt aerosol. In the troposphere, HCl is

primarily removed via wet and dry deposition, while hydroxyl radical

(OH) and ocean hydrolysis reactions are minor sinks. The hydroxyl

radical reaction proceeds as follows:

HCl + HO H O + Cl 2  (5)

HCl can impact the atmospheric chemistry of other species,

including other HAPs. For example, Selgneur, et al. identified the
H-7

reaction between HCl and elemental mercury, Hg(0), as one of the

relevant gas-phase reactions involving Hg(0) in the atmosphere. 14

 Hg(0) + HCl (g) (g) products (6)

Reaction rate parameter = 1.0 x 10 cm molecule s -19 3 -1 -1

In their theoretical study, Selgneur, Wrobel, and Constantinou

noted that specific products of this reaction have yet to be

identified. The authors found that aqueous-phase simulations of the

Hg(0)/Hg(II) concentration ratio, based on an atmospheric chemical

kinetic mechanism they developed, were quite sensitive to HCl

concentrations. When HCl is not present, the Hg(0)/Hg(II) ratio

calculated by the model ranged between 10 and 1,000,000. When HCl is

present, this ratio is believed to be on the order of 10. Liquid

water content, pH, and SO concentration were all found to have a large 2

and complex impact on the aqueous atmospheric chemistry of Hg(0). HCl

may thus affect the toxicity of mercury emissions from utilities. 

In summary, HCl is a highly reactive gas that is removed from the

atmosphere via wet and dry deposition. In general, because of its

high solubility, HCl will be removed from the atmosphere much faster

than SO or NO and will be deposited in close proximity to the 2 2

emissions source. It is possible that HCl may affect the toxicity of

mercury emissions from utilities.

H.2.3 HCl Atmospheric Transport

The remainder of this section discusses the propensity of HCl to

be transported in the atmosphere, which is an essential factor when

evaluating the impact of utility HCl emissions. The partitioning of

one chemical, such as HCl, in the atmosphere can influence what

compounds other HAP species become. 

HCl dissolved in clouds that are not precipitating can be

transported long distances, and thus may impact both acid rain and

PM . One way in which HCl may move long distances to sensitive 2.5 5

receptors is via the process of chloride ion catalyzation of SO to SO 2 4,

as described in the previous section. HCl is highly soluble, rapidly

dissolving in clouds or rain, and has been found to significantly

acidify cloud waters. A study conducted by March estimated that HCl

that had mixed with power station plumes contributed 57 percent of the

acidity measured in clouds.15

In a study of HCl at a rural site in northeast Austria, Puxbaum,

et al. found that elevated HCl concentrations were coming from air

parcels originating from the north and east. The authors believe the 3

source of this HCl is coal combustion. HCl’s winter transport was

believed to been enhanced by the shallower boundary layer and smoother

surface as a result of snow cover. The authors examined 48-hour back

trajectories. If they are correct in their assertion that the higher

HCl concentrations originating in the northern and eastern directions

are from coal combustion, this would lend credence to the possibility

of HCl transport, at least for up to a 48-hour period.
H-8

HCl was found to affect the gas/liquid partitioning of Hg(0). 14

For instance, when HCl is present, nearly all Hg(II) is present as

HgCl . However, when HCl is absent, most Hg(II) is present in the form 2

of Hg(SO ) with a small portion present as HgSO . When liquid water 3 2 3 2-

content, pH, and total SO concentration are kept constant at 0.1 g/m 2 3

and 4 and 10 ppb, respectively, moving from 0 g/m HCl to 1 g/m HCl 3 3

was found to raise the Hg(II) gas/liquid equilibrium ratio by up to two

orders of magnitude (from 0.0035 to 0.36).

According to the reports discussed in this section, conditions do

exist under which HCl can alter the spatial and temporal deposition of

acidic species. HCl concentrations were found to be among the factors

impacting the atmospheric chemistry of mercury.

H.2.4 HCl Terrestrial Processes

A chemical’s ability to accumulate in food chains and cause longterm harm is linked to its stability and persistence. Information on

terrestrial HCl chemistry was found on three topics: fog events,

damage to limestone, and the mobile anion hypothesis.

Fog events, by altering the oxidation rate of SO to SO and by 2 4

producing strong acidity, may have an impact on acid deposition. 

Concentrations of gaseous HCl are limited by the equilibrium with

solid aerosol phase NH Cl (see reaction #4). During fog events, high 4 1

concentrations of sulfate may be found in small aerosols as a result

of high-pH, aqueous-phase SO oxidation. The presence of HCl can lower 2

the pH to the point that SO oxidation is delayed, possibly altering 2

the spatial deposition of acid species. Acid aerosol deposition may

thus impact vegetation and soils in regions that experience fog.

Hutchinson, et al. demonstrated damage to limestone by gaseous

HCl. Humidity, degree of surface wetness, and temperature were all 8

shown to affect the intake of acids by limestone. HCl is deposited on

the stone by dry deposition which occurs in two stages. First, the

pollutant is transferred to the surface of the stone, then, it is

either absorbed or adsorbed by the stone. Relative humidity increases

the absorption of the pollutant, as does surface wetness due to the

solubility of many acids such as HCl. It was found that the reaction

of HCl and limestone is very rapid, occurs very close to the source,

and is more prevalent on moist surfaces. The calcium chloride

produced in the HCl/limestone reaction was easily removed by runoff. 

Degradation gave the limestone the appearance of acid rain damage,

which was attributed to the action of runoff.

As water flows through soils, it dissolves equivalent amounts of

anions and cations. Typically, these are bicarbonate and organic

anions balanced by base cations, hydrogen, and aluminum. The mobile

anion hypothesis, as described in the NAPAP State of the Science

Report 10, proposes that cation leaching in soils is controlled by the

availability of mobile anions. Acidic deposition may increase the 16

concentration of mobile strong-acid anions, thus hastening base cation

leaching in soils with medium to high base saturation, or leaching of
H-9

acid cations (H , Al ) into surface waters from low base saturation + n+

soils. The extent or magnitude of soil acidification depends on the

relative abundance of hydrogen, aluminum, and base cations, as

influenced by vegetation uptake and cycling and weathering rates of

soil minerals. Soils and soil water can be very acidic under natural

conditions without resulting in acidic surface waters. What controls

the concentrations of surface water acidity and base cations is the

limited mobility of anions from the soils to surface waters. Many

organic anions are retained in lower horizons of some soils or

oxidized by soil microorganisms. Nitrate and sulfate are commonly

retained through several biogeochemical mechanisms. Studies have

confirmed the important influence of anion production and mobility on

nutrient leaching following harvesting, fertilization, wastewater

application and atmospheric sulfuric acid inputs.16

Deposition of sea salts on acid soils has been demonstrated to

result in naturally acidic surface waters, which has been explained

through exchange of base cations for hydrogen and aluminum in the

soils and subsequent leaching balanced by mobile chloride anions. The

proposed mechanism for this effect has been discounted as a cause of

chronic acidification because it should result in a long-term buildup

of base cations or an alkalization of the soil-vegetation system. 

Such an increase in the base cation content of soils has not been

observed, and since such a buildup would eliminate the exchangeable

hydrogen and aluminum, and hence acidic runoff, the mechanism seems to

violate logic.16

Vegetation takes up and retains sea-salt-deposited base cations,

producing strong mineral acid (HCl) and acidifying the soil. As would

be the case for bicarbonate or organic anions, the presence of a

mobile anion (Cl ) not balanced by base cations could allow transport -

of hydrogen and aluminum from soils into surface waters if water did

not flow over weatherable materials. Erosion would also prevent the

buildup of base cations, maintaining soil acidity. There is no

evidence to indicate that vegetation grows faster, or that erosion of

forested systems is more rapid, in coastal areas that receive sea-salt

spray than in inland systems. Whether sea-salts acidify systems would

depend on whether vegetation uptake and removal of base cations

outpace salt spray and weathering. In near-coastal systems with low

weathering rates, or on poorly drained soils where roots and water do

not contact weatherable materials, salt spray may be the primary

source of base cations for vegetation uptake. In such systems

chloride would leach hydrogen and aluminum from acid soils, and

chronic acidification by this mechanism could occur. No near-coastal

whole-ecosystem study exists to evaluate rates of base cation supply

and cycling on a site-specific or regional basis. The sea-salt 16

mechanism may be an unlikely cause of regional acidification of inland

systems; however, it has implications for the potential contribution

to inland surface water acidification from atmospheric chloride

deposition resulting from utility emissions.
H-10

In summary, by altering SO oxidation, HCl can possibly impact 2

the spatial and temporal deposition of acidic species. HCl is also

capable of causing damage to limestone structures. Terrestrial HCl

transport has the potential to affect numerous features of ecosystems,

including soils, plants, and animals, both directly and indirectly.

The mobile anion hypothesis is a possible explanation for how utility

HCl emissions might impact

terrestrial processes.

H.2.5 HCl Aquatic Processes

No information specific to the aquatic stability or persistence

of HCl was found during this review. The aquatic chemistry of HCl is

important to the toxic burden, if any, it presents to aquatic

organisms. Two reports containing information on aquatic chemistry

were identified for this review. 

Stewart, et al. report that the chemistry of chlorinated

compounds in natural waters is complex. The pH, water temperature, 17

nitrogenous compounds, and types and amounts of organic matter present

all affect the persistence and toxicity of interim species of

ubiquitous chlorine-containing compounds. The oxidation of chlorinecontaining compounds such as OCl ultimately releases chloride ion. -

Skeffington lists a potential equilibrium constant (pK ) of -3 a

for the following reaction of hydrochloric acid and water:18

 HCl + H O2 3 H O + Cl (7) + -

A pK value in this range means that the above reaction goes to a

completion and HCl, as a strong acid, fully ionizes to hydronium and

chloride in water.

Part of the understanding we have of acid rain is based on what

we know about the behavior of chloride in aqueous environments. This

review includes a study on acid rain and a study investigating whether

chloride in waterbodies can be correlated with various anthropogenic

activities.

Peters examined the factors controlling chloride anion (Cl ) in -

two New York watersheds by examining precipitation, throughfall, soil

water, groundwater, and surface water. By combining previous 19

research on each of these water types, he was able to determine that

Cl cycling is more complex than the generally held view that there is -

rapid transport of atmospheric Cl deposition. In the Adirondack study -

system, it was shown that an additional Cl source in the watershed is -

the weathering of Cl from hornblende in surface minerals. Once Cl is - -

available in the watershed, the author hypothesizes that the biotic

system controls cycling; however, there was no direct quantification

of this hypothesis, only an estimated comparison. Peters also found

that annual Cl throughfall flux is two to five times that of -

precipitation, and that the watershed containing more hornblende in

the surface minerals had a three-fold greater net flux of Cl . -
H-11

Specifically, important findings include the observations that

atmospheric deposition is not the only primary source of Cl , and that -

the biotic cycling of Cl plays a significant role in transportation/ -

retention in the system. A better understanding of Cl cycling would -

be beneficial for estimating the impacts of acid deposition on a

watershed.

In a subsequent study of the interference of dissolved organic

carbon (DOC) on colorimetric Cl measurements, Norton, et al. challenge -

the Cl cycling findings of several studies, including that of Peters, -

above. The positive DOC interference was found to bias colorimetric 20

Cl measurements, leading to overestimates of Cl concentrations. - -

Norton, et al. present recalculated Cl concentrations from the Peters -

study, using their own empirical relationship between DOC and Cl . -

These authors conclude that the bias in data published in the

literature may have lead to spurious conclusions about Cl budgets in -

forests and watersheds, and that the issue of dry deposition and

recycling of Cl is far from settled. -

Zahn and Grimm developed a qualitative spatial analysis of land

use (agricultural, urban, forest, and groundwater protection areas) as

compared to nitrate and chloride concentrations in underlying

groundwater. Within the study area in southern Germany, it was found 21

that groundwater concentrations of nitrate ranged from 2 mg/L to >80

mg/L, with an average of 26.2 mg/L. Similarly, the concentration of

chloride in the groundwater samples ranged from 1 to 86 mg/L and had

an average value of 16.6 mg/L. There was no observed correlation

between depth of sample and concentration. Expected values of nitrate

and chloride, when considering regional geology, would be expected in

the range of 10 and 5 mg/L, respectively. In an attempt to correlate

the observed increase with anthropogenic sources, each observed

increase was correlated with an anthropogenic factor if possible

(i.e., agricultural runoff, roads, industries, waste dumps, etc.). 

When the results of these correlations were analyzed, it was found

that elevated chloride levels were most closely associated with

settlements, industries, waste dumps, and roadways, although the

correlation between land use and groundwater concentrations was not

apparent in all localities. It was determined that the only

substantial correlation was in nitrate and chloride concentrations

linked to waste dumps. The authors state that the anthropogenic

impacts are greater for chloride than for nitrate.

Zahn and Grimm note that one of the largest anthropogenic sources

of nitrate and chloride in soils (and hence groundwater) in Germany is

atmospheric deposition. Depending on land use, nitrate deposition

ranges from 1,500 to 2,050 kg/km /yr, and chloride deposition varies 2

from 670 to 1,170 kg/km /yr. Although it is qualitative, this work 2

indicates that there is a spatial correlation between land use

practices and the nitrate and chloride levels observed in groundwater. 

A more quantitative analysis would be necessary to validate these

results. The extensive development of quantitative spatial analysis

with global imaging systems will make the correlation of groundwater
H-12

pollution and land use or atmospheric deposition a potential screening

instrument for groundwater contamination by anthropogenic activities,

such as utility HCl emissions in future research.

In summary, chloride cycling in watersheds was found to be more

complex than researchers previously thought. The traditional view has

been that atmospherically deposited chloride is rapidly transported. 

This research indicates that a better understanding of chloride

cycling is necessary. A qualitative spatial analysis found that

elevated concentrations of chloride in groundwater were most strongly

correlated with waste dumps.

H.2.6 HCl Human Health Impacts

Inhalation of HCl can cause injury to the respiratory tract. 

Risk from exposure to high concentrations of HCl is most likely to

arise in occupational settings or through an industrial or

transportation accident, while exposure to relatively low HCl

concentrations occurs over a wide area subject to emissions from

anthropogenic sources, including utility, industrial, and waste

combustion, as well as natural sources. 22

Chapter 6 of this report uses the RfC for HCl of 20 i g/m to 3

estimate the chronic noncancer hazard quotient (HQ), and lists the

California Air Pollution Control Officers Association (CAPCOA) acute

Reference Exposure Level (REL) as 3,000 g/m for HCl. Additional 3

information on HCl acute toxicity was obtained from a Lab Chemical

Safety Summary compiled by the National Research Council and EPA

Region III. The National Research Council reported a Threshold Limit 23

Value (TLV) of 5 ppm and a Permissible Exposure Level (PEL) of 5 ppm

for HCl. Toxicity values for chlorine range from 0.5 ppm to 1.0 ppm. 

Table H-1 contains the National Research Council’s reported toxicity

values. EPA Region III’s Superfund Risk Based Concentration (RBC) for

chlorine is 370 g/m .3

The results of a WHO review on hydrogen chloride exposure to

humans yielded evidence of local irritation to the upper respiratory

tract. Mucous membranes were found to be especially susceptible. 22

Exposure at higher concentrations caused conjunctival irritation and

superficial corneal damage. Hydrogen chloride can cause transitory

epidermal inflammation if it comes in contact with damp clothing or

skin. Long-term exposure may induce erosion of the inciso-labial

surfaces of the teeth. Even though one study did show possible

induction of a tumor growth using HCL, the authors conclude that there

are no mutagenic, carcinogenic, or teratogenic effects related to HCl. 

The WHO task force could not, due to the dearth of data, determine a

not-to-be-exceeded ambient HCl concentration.

In summary, the National Research Council’s Lab Chemical Safety

Summary contains pertinent toxicity data on HCl. Evidence of local

irritation to the upper respiratory tract by HCl was found and longterm exposure may cause tooth erosion. The WHO concluded in a review 
H-13

Table H-1. National Research Council Acute Toxicity

Values for Chlorine

Chemical TLV-TWA PEL

Hydrogen chloride 5 ppm 5 ppm

Chlorine 0.5 ppm 1.0 ppm

PEL = Permissible Exposure Limit

TLV = Threshold Limit Value

TWA = Time Weighted Average

on HCl that there are no mutagenic, carcinogenic, or teratogenic

effects related to HCl.

H.2.7 HCl Vegetation Impacts

The behavior of vegetation with regard to uptake, absorption,

translocation, distribution, and accumulation of chloride might all

have important effects on survival or growth. One study directly

pertaining to chloride’s effects on vegetation was found.

In a Texas study designed to ascertain the impact of marine salt

on vegetation, McWilliams and Sealy found a strong correlation between

atmospheric chloride levels and leaf chloride levels in Spanish moss

(Tillandsia usneoides). Most nonanthropogenic chloride is 24

attributable to marine sources. T. usneoides, an epiphytic plant, is

ideal as a subject for studies of atmospheric contaminants on plants. 

This study suggests that atmospheric and leaf chloride levels are

closely correlated, at least in certain situations.

In their discussion of past findings on chloride salt vegetation

toxicity, McWilliams and Sealy note that the “wind form” of vegetation

along the North Carolina coast was found to be the result of lethal

salt effects rather than wind. Atmospheric chloride can enter plants 24

via aerial organs and concentrate in leaf tissues, leading to foliar

damage.

H.2.8 HCl Terrestrial Animal Impacts (Wild and Domestic)

HCl may reach plants and soil via wet and dry deposition. Once

present it is available for intake by wild and domestic animals. 

Tukey et al. report that very little chloride leaches from leaves. 25

Therefore, animals foraging on HCl-enriched foliage will ingest

additional chloride.

Research on the effects of HCl on animals was reviewed by WHO. 22

Symptoms include eye and nasal irritation, but the mucous membranes

and respiratory tract are the primary targets. Edema is

characteristic of the initial symptom of hydrogen chloride toxicity,

proceeding to additional inflammation, degeneration, and necrosis when

in contact with tissue. The authors of the various studies noted that
H-14

chlorine does not appear to be teratogenic, mutagenic, carcinogenic,

or cocarcinogenic in animals, and that the carcinogenic potential of

hydrogen chloride could not be assessed due to a lack of adequate

studies.

H.2.9 HCl Aquatic Animal Impacts

This section addresses the impacts of HCl on aquatic animals. 

Sources of information on the ingestion, distribution, and

accumulation of HCl in aquatic animals, along with its toxicity, were

reviewed. Two sources on aquatic HCl toxicity were found and are

described below. No information pertaining to ingestion,

distribution, or accumulation was found for this review.

Stewart, et al. report that some chlorinated compounds, such as

OC1 , found in natural waters are quite toxic to freshwater biota, -

while others, such as Cl , are not. Chlorine is widely used in water - 17

treatment as a disinfectant for drinking water and wastewater prior to

discharge; Manning, Wilson, and Chapman found it to be toxic to

aquatic life. Many treatment plants use ammonia and chlorine 26

together for the treatment process in a ratio of 3:1 ammonia to

chlorine. This study obtained data on Australian organisms to support

development of appropriate treatment plant controls. It included an

evaluation of the toxicity of a mixture of chlorine and ammonia to

assist in evaluating the potential impacts of drinking water on

ecosystems. The water flea (Ceriodaphnia dubia) was used as the

freshwater organism, and the eastern king prawn (Penaeus plebejus) was

used as the saltwater organism.

There is a distinct difference between the American and

Australian C. dubias; the Australian water flea is not as sensitive to

chlorine as the North American variety. Based on results of 24-hour

LC obtained in these tests, the environmental concern level for 50

chlorine in marine situations is 0.0018 mg/L and the environmental

concern level for chlorine and ammonia in a freshwater situation is

0.003 mg/L. These figures are in line with the standards determined

in the northern hemisphere. The chlorine/ ammonia mixture is not seen

in salt water, so a test was not performed on the salt water organism. 

Environmental concern levels for chlorine and ammonia alone were not

reported for fresh water.

H.3 FINDINGS FOR HYDROGEN FLUORIDE

H.3.1 HF Emissions and Formation

Hydrogen fluoride (HF) can exist as either a colorless, corrosive

liquid or a gas at room temperature and is used in numerous production

processes, including those of high-octane gasoline, aluminum,

plastics, electrical components, fluorescent light bulbs, and

refrigerants. 27

The information on nationwide utility HF emissions was obtained

from Table 3-3 in Chapter 3 of this report. The 1990 HF emissions are

reported as 19,500 tons/year. Utility emissions are the most
H-15

significant anthropogenic source of atmospheric HF. Other important

sources in addition to utilities are phosphate manufacturing and

aluminum manufacturing.

Anthropogenic sources are responsible for a considerable amount

of atmospheric fluoride. Anthropogenic emissions of HF originate 27

from coal combustion and the aluminum, phosphate, and steel-making

industries. For example, fluoride in the atmosphere of the Colorado

Plateau region of the western United States is believed to originate

primarily from power plant emissions; this is a result of the high

fluoride content of western coals and the lack of aluminum

electrolysis plants in the region. Relatively small amounts are 28

emitted from the production of HF itself (2,400 metric tons per year). 

 Volcanoes are the primary natural sources of HF. Total annual

volcanic emissions of fluoride-containing compounds are estimated to

be between 1 and 7.3 million metric tons per year. Ocean spray,

fires, and dust from soil and rock weathering also contribute fluoride

to the atmosphere; however, annual emissions from these sources are

believed to be negligible.

Atmospheric concentrations of fluoride in remote rural areas are

reported to be approximately 0.1 g/m , which is at the limit of 3

detection. Urban fluoride concentrations ranging from 0.1 g/m to 29 3

1.0 g/m have been observed in British industrial cities. However, 3 29

HF accidentally released from chemical plants may produce

concentrations on the order of 50 g/m for several hours. As of 3 

August 1994, only one HF monitoring station had been operated in the

United States, with 20 samples measured. Measured concentrations of 7

HF ranged between 1 g/m and 8 g/m . 3 3

As in the case of HCl, ambient concentrations of HF will be of

interest as part of the effort to achieve compliance with the new

particulate matter standards. If a significant portion of ambient HF

is in the fine fraction, it could contribute to PM exceedances. 2.5

Information was found concerning the relative proportions of

gaseous to particulate fluoride emissions, fluoride particulate

composition and diameters, HF formation from volcanic emissions,

general formation material, and HF reactants. Hance, et al. report

that up to 40 percent of atmospheric emissions of industrial fluoride

are gaseous, with the remaining 60 percent emitted as particulate

matter. Slooff, et al. estimate that about 25 percent of atmospheric 27

fluoride is emitted as particulate, with the remaining 75 percent in

the gaseous state (mostly HF).30

Gaseous HF undergoes hydrolysis before dispersion throughout the

atmosphere. Fluoride particulates vary from distinct minerals to 27

alumina with HF adsorbed to its surface, and particle diameters range

from <0.1 m to approximately 10 m. Davison reports that fugitive 29

fluoridated alumina particle sizes near a dry-scrubbed aluminum

smelter can range from <1 m to about 5 m in diameter.29
H-16

Volcanic emissions are not usually predominately HF, but rather

other fluoride-containing compounds which react in the atmosphere to

form HF. These compounds include boron trifluoride, carbonyl

fluoride, phosphorus pentafluoride, silicon tetrafluoride, sulfur

tetrafluoride, and phosphorus trifluoride, all of which produce HF via

hydrolysis.

Both the conditions under which HF is released and the

atmospheric conditions can effect the behavior of HF upon release. 31

Superheated HF that is released under pressure will form a cloud of HF

vapor and aerosol, which will react with water vapor in the air. If

present in a concentration greater than 40 percent, both anhydrous and

aqueous HF will react with atmospheric moisture to produce white

fumes, which have a pungent odor and are extremely irritating if they

are inhaled or if they come in contact with an organism. HF molecules

form variable length chains up to (HF) at ambient temperatures through 8

hydrogen bonding. HF can also be present as single molecules at

higher temperatures. HF’s hydrogen bonding leads to the formation of

vapor clouds that can be either neutral or positively buoyant,

depending on atmospheric conditions.

Based on thermodynamic data, Slooff, et al. suggest that HF

reacts with other gaseous acidic species, including HNO , H SO and HCl, 3 24

present in the atmosphere. Therefore, it is possible that HF may 30

have some indirect, limited impact on acid rain formation, and may

possibly have an indirect, limited impact on visibility and fine-PM

issues. As is the case with HCl and chloride salt formation, Spicer,

et al. found that HF reaction products also include primarily fluoride

salts. The authors describe the identification of reaction products

as qualitative in most cases, because few studies reported mass

balance information.11

H.3.2 HF Atmospheric Processes

Stability and removal mechanisms of atmospheric HF determine its

persistence. Spicer, et al. describe HF as moderately persistent in

the atmosphere, with an estimated lifetime of between 1 and 5 days,

which places it in the 46 percent of HAPs with lifetimes in this

range. Slooff, et al. estimate the half-life of gaseous HF to be 11

0.54 days, while fluoride aerosol has a half-life of 2.1 days. 30

Spicer, et al. also found deposition, both wet and dry, to be the

primary atmospheric removal process for HF; thus, they do not expect

HF to undergo significant chemical transformation in the atmosphere. 11

Hance, et al. found precipitation to be the major global route of

atmospheric HF removal. Slooff, et al. also cite wet and dry 27

deposition as the primary routes of HF removal from the atmosphere. 30

Little information was available concerning atmospheric chemistry

or degradation of HF. The material that was reviewed indicates that

HF does not biodegrade, regardless of whether it is released to air,

water, or land.31
H-17

In summary, HF is described as moderately persistent in the

atmosphere, with an estimated lifetime of approximately 1 to 5 days. 

Wet and dry deposition are the primary routes of HF removal from the

atmosphere. 

H.3.3 HF Atmospheric Transport

Fluoride emissions from utilities are transported on a regional

scale. In fact, fluoride (F :SO ratio) was used as an atmospheric total x

tracer in one study reviewed because this ratio, which is relatively

constant for the coal-fired utilities examined, provides a

characteristic utility emissions fingerprint.

In a source-apportionment study of regional air pollution

transport to the Canyonlands National Park, Eatough, et al. found

evidence that coal-fired utility emissions, among other sources,

contribute to measured concentrations of atmospheric fluoride at

distances of up to 500 km. Slooff, et al. report that atmospheric 32

fluoride, especially aerosol fluoride, can be dispersed over great

distances by wind or atmospheric turbulence. Fine particles will 30

also be deposited over a greater area than large particles. Leece, et

al. found that elevated levels (18-21 ppm) of fluorides emitted from a

power station at Liddell, New South Wales, could be detected in grape

leaves at distances of up to 37 kilometers. Background foliar 33

fluoride concentrations were measured at less than 1 ppm.

The measurement of fluoride transport and deposition has been

reported to have problems of accuracy due to methods used. Davison

cites analytical problems and the lack of a suitable isotope as

adversely affecting the scope and accuracy of measurements of fluoride

deposition and transport. The lack of an ideal analysis technique 29

for biological materials has resulted in the absence of an absolute

laboratory standard. This has commonly resulted in a sampling error

of ±10 percent for repeated analysis of the same sample. The short

half-life of the F isotope (1.8 hours) has precluded its use for 18

measuring fluoride deposition and movement through trophic levels. 

Thus, fundamental problems with fluoride present limits to the study

of deposition, pathways, and accuracy of measurements.

To summarize, evidence was found that coal-fired utility

emissions, among other sources, contribute to measured concentrations

of atmospheric fluoride at distances of up to 500 km. The measurement

of fluoride transport and deposition has been reported to have

problems of accuracy due to methods. 

H.3.4 HF Terrestrial Processes

The stability and persistence of a chemical is linked to its

ability to accumulate in food chains. Information concerning HF

volatilization and the ability of soluble fluoride to alter physical

and chemical soil properties is reviewed in this section.

Davison’s synopsis of research shows that fluoride is lost from

the various surfaces on which it is deposited and leaves ecosystems at
H-18

a rapid rate. Some possible mechanisms have been investigated with 29

varying success. The pathway of volatilization to the air was

presumed to account for most of the lost fluoride. However, research

on this pathway yielded contradictory results. The hypothesis was

proposed that although fluoride ions are not significantly volatile, a

portion of fluoride forms complexes with hydrogen ions to form HF,

which is volatile at a pH of less than 5. Davison concludes that the

volatilization pathway as a route of fluoride export from ecosystems

needs further investigation.

Arocena, et al. investigated the effects of soluble fluoride in

the process water and leachate of phosphogypsum formed by the

production of phosphate fertilizer from phosphate rock. Analyses of 34

both calcarious and non-calcarious soil clay components showed a

mineral composition of smectite, kaolinite, mica, and trace amounts of

chlorite. The phosphogypsum process water was shown to dissolve much

of the fine clay fraction, as well as the smectite of the coarse clay

fraction. The observed changes in the clay fraction can significantly

alter physical and chemical properties of the soil, such as hydraulic

conductivity, absorption, and ion exchange capacities. Although the

HF described in this article originated in the production of

fertilizer, the effects of HF in acid rain could be similar. The

research in this study is significant, because the dissolution of clay

by fluoride ion could be an important impact on ecosystem soils, as

well as the performance of landfill clay liners.

Fluoride in non-saline soils is associated with the clay-sized

fraction because it reacts with aluminum compounds and various

minerals. Soil can be both a sink and source of fluoride. In 29

undisturbed soil profiles, the fluoride gradient usually increases in

concentration with depth as a result of leaching. However, because

soils can fix large amounts of fluoride, the rate of leaching is

usually slow. Soil fluoride fixation also results in retention in the

upper soil horizons. Fluoride is not usually available or labile in

soils. Soil acts as both a sink and resistance, slowing and

controlling fluoride transport to plant and aquatic systems. Ash

deposition after fire events, saline (low calcium, high pH) soil

conditions, and formation of aluminum-fluoride complexes as a result

of aluminum smelter fluoride particulates are possible exceptions that

can facilitate rapid water-soluble fluoride uptake and transport.

If sufficient calcium is present in soil or water, it will form

an insoluble solid with fluoride ion, which removes it as an immediate

environmental hazard. The natural buffering capacity of soils or

water, or dilution can potentially reduce acidity added by the

presence of HF.

Information on the ability of disturbances to cause HF mobility,

and HF’s ability to mobilize soil aluminum were reviewed. HF does not

biodegrade, regardless of whether it is released to air, water, or

land. Disturbance events can impact fluoride’s movement through an 31

ecosystem. For instance, Murray found that fire increases the water-
H-19

soluble fluoride fraction in standing soil by a factor of two, but

only for a period of a few months. 35

Bond, et al. conducted a study on the effects of sulfate and

fluoride on soil pH, soil aluminum (Al) concentrations, and Al

transport. The experiments were designed to test the hypothesis that 36

soil fluoride increases Al mobility in soils and raises soil pH, and

were performed in soil columns. The results were compared to the

findings of a similar set of batch studies that found that fluoride

increased Al soil transport. Overall, the authors determined that

under the more realistic soil:solution ratios experienced in soil

columns, fluoride did not increase Al mobility. In fact, both sulfate

and fluoride were found to slightly retard aluminum’s mobility through

soils. This experiment suggests that fluoride’s hypothesized ability

to raise soil pH by mobilizing Al was spurious, and that fluoride is

not a preferred agent for alleviating soil acidity.

To summarize, fluoride is believed to be lost from the various

surfaces on which it is deposited, and leaves ecosystems at a rapid

rate. Soluble fluoride-containing process water and leachate of

phosphogypsum were shown to dissolve much of the fine clay fraction,

as well as the smectite of the coarse clay fraction of soils. HF does

not biodegrade in air, on land, or in water. The natural buffering

capacity of soils, water, or dilution can potentially reduce acidity

added by the presence of HF. Fire was found to increase the watersoluble fluoride fraction in standing soil, and sulfate and fluoride

were found to slightly retard aluminum’s mobility through soils.

H.3.5 HF Aquatic Processes

Fluoride is known to be a major component of seawater, with a

residence time of approximately one million years. In fresh water 30

with pH greater than 5, fluoride is mainly present as fluoride ion,

based on stability diagrams derived from thermodynamic calculations.

In fresh water, fluorapatite, Ca(PO ) F or Ca (P 4 3 10 4 6 2 O ) F , is only

somewhat soluble. Slooff, et al. provide the following equilibrium 30

reactions:

 Ca(PO ) F + 6H 43 2 4  5Ca + 3H PO + F (8) + 2+ - -

log K = -1

and:

Ca (PO ) F 10 4 6 2 4  10Ca + 6PO + 2F (9) 2+ 3- -

log K = -119

which show that the solubility product is very low.

Elrashidi and Lindsay found that in waters with pH greater than

5.6, the fluoro-hydroxy-apatite system is dominant. The following

equation describes the fluoro-hydroxy-apatite system:37
H-20

Ca (PO ) F + H O 5 43 2 5 43  Ca (PO ) OH + F + H (10) - +

log K = -15.14

Dutch water quality officials regularly measure aquatic

concentrations of fluoride that are three orders of magnitude greater

than what is expected, based on the above calculations. Accordingly,

there must be a factor increasing fluoride concentrations in fresh

water beyond those accounted for in the theoretical calculations of

thermodynamic equilibrium. Researchers are not able to balance the

fluoride budget in the Netherlands. The addition of atmospheric HF to

the equilibrium in reaction (10) would be likely to generate more

fluorapatite.

Fluorapatite discharged from phosphate ore-processing plants or

precipitating in water may accumulate in water body sediments. 30

Slooff, et al. note the need for greater understanding of the

distribution of labile and insoluble fluoride to help resolve the

uncertainty about partitioning of fluoride in waterbodies and

sediments.

H.3.6 HF Human Health Impacts

Table 6-10 of this report lists the California Air Pollution

Control Officers Association (CAPCOA) acute Reference Exposure Level

(REL) for HF as 580 g/m . Additional literature reviewed on human 3

health impacts centered on ingestion and inhalation. Information on

HF acute toxicity was obtained from National Research Council Lab

Chemical Safety Sheets, which provided a TLV-TWA of 3 ppm (2.6

milligrams per cubic meter), and a PEL of 3 ppm for HF. Table H-2 23 

contains toxicity values for several fluoride species. 

Whitford reviewed fluoride ingestion findings in a number of

studies. The difficulty in estimating the lethal dose or potentially 38

toxic dose of fluoride in humans, particularly in children, can partly

be attributed to the problem of estimating the actual dose after a

poisoning event has taken place. Human studies of the relationship

between treatment with fluoride and bone strength have revealed a

conflicting picture. If there is either a beneficial or a detrimental

effect, Whitford claims it is subtle, and convincing evidence will

require large, carefully controlled, prospective research studies.

According to Davison, water sources in many cases account for

about half of human fluoride intake in non-fluoridated water

environments. Daily human fluoride intake varies as a result of age, 29

fluoride water content and consumption rate, and diet. Whether

ingested fluoride is available for uptake will depend on its chemical

state, which also affects retention time and pathway through the body. 

The chemical form of fluoride ingested will affect the pathway of

elimination from the body and also the chemical species eliminated.

A considerable amount of information was available on the topic

of HF’s toxicity to humans. This section discusses the wide range of

potential toxic effects of HF, as well as some possible benefits.
H-21

Table H-2. National Research Council Acute Toxicity

Values for Fluorides

Chemical TLV-TWA PEL

 Hydrogen fluoride 3 ppm 3 ppm

 Fluorine 1 ppm 0.1 ppm

 Fluorides 2.5 mg/m3 2.5 mg/m3

PEL = Permissible Exposure Limit

TLV = Threshold Limit Value

TWA = Time Weighted Average

Fluoride, the ionic form of fluorine, is a potent inhibitor of

many enzymes. Unlike iodide it does not accumulate in the thyroid. 38

The rate of elimination of fluoride from the kidneys is many times

greater than it is for other halogens. It stimulates new bone

formation and can inhibit, or even reverse the formation of dental

caries. Fluoride has other beneficial effects. It may reduce the

incidence and severity of osteoporosis, and increased consumption of

fluoride following the introduction of water fluoridation in the

United States may be related to the sharp decline in the death rate

due to heart disease.

Fluoride is a hazardous substance when taken acutely in large

doses, and numerous claims exist of harm arising from the chronic

ingestion of low doses. The effects range from dental fluorosis, 38

gastric disturbances, and reductions in urinary concentrating ability,

to skeletal fluorosis and death. Based on an earlier estimate by

Hodge and Smith, the certainly lethal dose (CLD) of sodium fluoride is

estimated to be between 70 and 140 mg/kg body weight for an adult. 39

Based on a report by Dukes, a dose of approximately 4 mg/kg may be

fatal for a young child. Insoluble forms, such as calcium fluoride 40

and cryolite (Na AlF ), are less toxic than sodium fluoride, as they 3 6

are less well absorbed. Monofluorophosphate (MFP) is approximately

half as acutely toxic as sodium fluoride.

Whitford summarizes the findings pertaining to humans of a recent

USPHS report on fluoride (one of three recent critical reviews) as

follows: “(1) there is no detectable risk of cancer in humans

associated with the consumption of optimally fluoridated water;

(2) there is no indication that organ systems are affected by chronic,

low-level fluoride exposure (although more research on human

reproduction, for which there is a paucity of data, was recommended);

(3) fluoride exposure is not associated with birth defects, including

Down’s syndrome; (4) genotoxicity studies, which are highly dependent

on the methods and models used, have yielded contradictory results so

that any possible effect of fluoride in humans and laboratory animals

remains unresolved; (5) the prevalence of dental fluorosis in the USA
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is higher now than in the 1940s but there is disagreement about

whether this condition is a toxic effect; (6) crippling skeletal

fluorosis has not been and is not a public health problem in the USA;

(7) the beneficial effect of high fluoride regimens in reducing

osteoporosis has not been demonstrated; and (8) further

epidemiological studies are required to determine whether or not an

association exists between various levels of fluoride in the drinking

water and bone fractures.”38

According to the WHO, excessive fluoride exposure may affect

human health, but it has been difficult to substantiate cases of human

fluorosis brought about solely by exposure to atmospheric fluorine,

even in old, poorly controlled industrial environments.41

In summary, fluoride is a potent inhibitor of many enzymes,

stimulates new bone formation, can reverse dental caries, and may

reduce the incidence and severity of osteoporosis. Fluoridation of

water in the United States may be related to the sharp decline in the

death rate due to heart disease. Adverse effects of fluoride on human

health include dental fluorosis, gastric disturbances, reductions in

urinary concentrating ability, skeletal fluorosis and even death. 

Whitford’s recent review of fluoride’s toxicity provides a

comprehensive synopsis of the subject.

H.3.7 HF Vegetation Impacts

This section covers available literature pertaining to fluoride

uptake and accumulation by vegetation, as well as injury to vegetation

by fluoride. Studies reviewed pertain to Brazilian rainforest trees,

grape leaves, lichens, and pine trees.

Klumpp, et al. investigated fluoride accumulation in three tree

species in a field study in the Atlantic Rainforest, near Cubatao,

Brazil, where a large fertilizer industry is a source of fluoride

emissions. All were pioneer tree species known to be ubiquitous in 42

the study area. The vegetation and soils within a 60 km radius of 2

Cubatao have experienced severe damage as a result of 30 years of high

pollution, resulting in a high frequency of landslides and the

replacement of primary vegetation with secondary vegetation types over

a vast area. The study assessed the spatial and temporal distribution

of fluoride’s effects on the natural vegetation in the region (passive

monitoring), as well as fluoride’s effects on exposed Tibouchina

seedlings being tested as a cumulative indicator species (active

monitoring). Foliar fluoride concentrations were measured at four

sites with varying pollutants and concentrations.

Inherent differences in the resistance of some tropical tree

species to fluoride may be related to their capacity to accumulate

aluminum. Although some plant species are tolerant of elevated

fluoride levels, the storage of large amounts of fluoride in plant

tissues may present a risk to the ecosystem. For instance, the

movement of fluoride from plants to soils may affect nutrient

turnover. Litter decomposition impacts may include fluoride
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accumulation in soil detritivores, increased humus layer, and reduced

microbial activities, all of which have been observed in the vicinity

of aluminum smelters. Foliar accumulation of fluoride may also impact

plant-insect relationships, which may have ramifications through the

food chain, affecting insects, vertebrate herbivores, or rodents.

A study of fluoride by Slooff, et al. conducted under the

auspices of the Dutch National Institute of Public Health and

Environmental Protection found that plant uptake of fluoride is

limited to the smaller, water-soluble and labile fractions. The 30

major pathway of fluoride to plants is atmospheric deposition both to

plant surfaces and intake through the stomata.

Leece, et al. found that in foliage, gaseous HF dissolves within

the leaf and is transported to the leaf margins, where it

accumulates. If the ionic concentrations are high enough, stable 33

salts may form, thereby rendering both cations and anions

physiologically inactive. Reactions are as follows:

HF (g) H + F (11) + -

 Ca + 2F CaF (12) 2+ - 2

The authors found evidence of fluoride transport of up to

37 kilometers. Although elevated foliar fluoride concentrations

attributable to power plant emissions were found, symptoms of fluoride

toxicity were not. Grape leaves were found to accumulate up to 40 ppm

fluoride without developing toxicity symptoms in drought conditions. 

Because foliar damage caused by environmental conditions and

physiological stresses (e.g., moisture stress, potassium deficiency,

and chloride toxicity) appears similar to fluoride contamination, the

authors urge chemical analysis of foliar tissue.

In an effort to obtain data on fluoride emissions from volcanos

over time, Davies and Notcutt sampled lichen species on and around Mt.

Etna, on the island of Sicily. Lichen fluoride concentrations were 43

measured in a preliminary survey in 1985 (77 sites) and then again in

1987 (56 sites). The fluoride accumulation pattern was affected by

the volcano, with concentrations highest downwind of the plume and

tapering off with increasing distance from the volcano, as mediated by

topography and prevailing winds. This paper indicates that lichens,

because of their ability to accumulate fluoride from both industrial

and natural emissions sources, can be a useful tool for measuring

utility and/or industrial fluoride emissions. The higher fluoride

concentrations (>100 ppm) found at five sites are similar to those

found at certain types of industrial sources. Noting studies that

found elevated fluoride levels in vegetation and small mammal bones

near an aluminum smelter, the authors suggest that gaseous fluorides

emitted from both industrial and natural sources may have effects

higher up the food chain.
H-24

Amundson, Belsky, and Dickie investigated foliar fluoride

concentrations, foliar decomposition, and upper soil horizon fluoride

accumulation in a coastal pine plantation in both close and distant

proximity to a new aluminum smelter located near Charleston, South

Carolina. Pine stands within 0.8 km of the smelter had foliar 44

fluoride concentrations significantly higher than one stand 1.8 km

from the source. Fluoride emitted from the newly operating smelter

was not observed to alter needlefall amount, temporal patterns, or

rates of needle decomposition over a six-month period, although it did

accumulate in needles. Over a seven-year period, soluble fluoride

concentrations in the upper 10 cm of soils increased significantly,

but only at sites nearest to the emissions source. However, these

findings on two species of Pinus may not apply to other arboreal

species or herbaceous plant species. It is possible that as needles

decompose in the absence of fire for longer than 6 months, the

concentration of fluoride may eventually increase to levels that

inhibit decomposition. Questions remain about whether deposited

fluoride is adsorbed to the needle surface or absorbed inside the

needle, and whether increased soluble fluoride in the soil contributes

to elevated levels in the needles.

Atmospheric fluoride is capable of injuring certain plant species

at lower concentrations than any other air pollutant studied. For 45

instance, a review conducted by Davison noted that ripening peach

fruit tissues can be visibly injured by fluoride concentrations of 0.3

g/m for 12 hours, 5 days a week. However, most plant species are 3 29

considered resistant to fluoride and can tolerate concentrations of up

to 30 g/m .3

In their study conducted on and around Sicily’s Mt. Etna, Davies

and Notcutt found no morphological damage to lichen species, although

lichen samples from five sites contained fluoride concentrations in

excess of 100 ppm.43

In the Brazilian Atlantic Rainforest study, Klumpp, et al. found

one of the three native plant species to exhibit marginal foliar

necroses and malformations at the most polluted site. Crown heights 42

of all three species were lower at the polluted sites when compared

with the reference area. Fluoride was found to be the most important

pollutant contributing to vegetation damage in one section of this

tropical rainforest, although ambient concentrations of other

pollutants, including ozone, NO , ammonia, and SO , are likely to be x x

affecting the vegetation as well. The study notes that the question

of whether Al-F complexes are phytotoxic has not been clarified to

date.

In summary, although some plant species are tolerant of elevated

fluoride levels, the storage of large amounts of fluoride in plant

tissues may present a risk to ecosystems. Inherent differences in the

resistance of some tropical tree species to fluoride may be related to

their capacity to accumulate aluminum. Plant uptake of fluoride was

found to be limited to the smaller, water-soluble and labile
H-25

fractions. The major pathway of fluoride to plants is atmospheric

deposition. Although elevated concentrations of fluoride were found

in lichens, grape leaves, and pine trees, foliar damage was not. In

one section of a Brazilian rainforest, however, fluoride was found to

damage vegetation.

H.3.8 HF Terrestrial Animal Impacts (Wild and Domestic)

The literature reviewed for this section pertained to small

mammal and raptor species. Information concerning fluoride’s

accumulation in food chains, distribution, toxicity, and soil

contamination as a route of fluoride transfer was also reviewed.

Fluoride can be ingested by small mammals as they consume

vegetation. However, few of the numerous studies on fluoride’s

effects on vegetation attempt to quantify cycling from vegetation to

soil and animals. Boulton, Cooke, and Johnson conducted a study on

the toxicity of fluoride on short-tailed field voles and found that 19

to 43 percent of daily uptake of fluoride was retained. Fluoride 46

concentrations were increased in femur and incisors, but significantly

decreased in molars. Incisors also had visible lesions. The

difference in results between study locations is attributed to the

chemical speciation of fluoride in the plant material and the variable

levels of other dietary components, including Ca and Al . The 2+ 3+

authors of this study suggest that for absorbed fluoride, kidney

excretion is likely a minimal removal pathway as compared to

deposition in bones and teeth.

In another study, Boulton, Cooke, and Johnson compare the toxic

effects of inorganic fluoride on four species of small mammal:

laboratory white mice (Mus musculus L.), wood mice (Apodemus

sylvaticus L.), short-tailed field voles (Microtus agrestis L.), and

bank voles (Clethrionomys glareolus L). The primary difference

between the species was the fluoride uptake rate, which was heavily

influenced by the administration in the drinking water. The order 47

of uptake among species was M. Agrestis > C. Glareolus > A.

Sylvaticus > M. Musculus. The observation of intake helps explain

premature mortalities observed in M. Agrestis and C. Glareolus. 

Differential absorption and retention were observed in metabolic cage

studies of M. Agrestis and M. Musculus. Interspecific variation

related to intake of fluoride was observed in assays for fluoride

concentration in femurs, molars, and incisors. Severe dental lesions

were observed in animals surviving the highest dose at the termination

of the experiment. The authors conclude that the interspecies effects

in this experiment were the consequence of different intake levels. 

This report may be of interest because it examines several wildlife

species, the use of which opens the possibility for the development of

biomarkers. The observation that effects were primarily due to intake

rates may also help identify susceptible species that have increased

rates of consumption of fluoride-contaminated media.

Seel, Thomson, and Bryant measured bone fluoride in four

predatory bird species in the British Isles: sparrow-hawk (Accipiter
H-26

nisus), kestrel (Falco tinnunculus), barn owl (Tyto alba), and tawny

owl (Strix aluco). Fluoride loads increased with age in all four 48

species. The data for all species were combined for comparison with a

known fluoride gradient on Anglesey, North Wales, and an estimate of

the regional occurrence of fluoride in predatory birds in the British

Isles. Lower-than- average fluoride loads occurred in most of the

study area, but higher-than-average loads were reported in more

industrial regions of England. Although females weighed more in all

species, fluoride bone concentrations were found to be greater in

males of all species examined. Seel, Thomson, and Bryant report that

fluoride does tend to bioaccumulate in birds, and therefore might have

secondary or tertiary food chain effects. 

Conflicting information was found concerning fluoride

accumulation in food chains. Slooff, et al. report that exposure to

high concentrations of fluoride causes terrestrial organisms to

accumulate more fluoride than when exposed to background levels. 30

Both invertebrate and vertebrate predators accumulate higher levels of

fluoride than herbivores. Thus, although the data are ambiguous,

predator concentrations of fluoride are higher, indicating moderate

biomagnification. However, Davison reports that although data are

limited, it is generally assumed that fluoride concentrations do not

increase with trophic level.29

Of the three main routes of fluoride transport to animals—

inhalation, ingestion, and deposition to outer surfaces— inhalation

is not usually considered an important route. This is because the 29

amount of fluoride taken up and retained by the lungs is small both in

magnitude and in comparison to the other two routes. Although Davison

reports that no analyses exist which allow the calculation of

deposition rates on skin, wings, shells, etc., these routes may be

important, especially because some particulate fluorides are known to

be toxic and have been used as insecticides. The preening of bird

coats, pollen collection from bee bodies, and coat-licking by cattle

are all examples of how fluoride deposition can be taken up by

animals. Variations in fluoride concentrations within plant parts can

result in animal species with differing feeding niches ingesting

different amounts of fluoride. For instance, animals feeding on

nectar (e.g., adult butterfly), phloem (e.g., aphid), or whole leaf

tissues (e.g., caterpillar) of the same plant are likely to ingest

varying amounts of fluoride. Invertebrates have several attributes

which render them more suitable for fluoride pathway analysis and

transfer rate analysis than vertebrates, such as smaller feeding areas

and smaller variation in longevity between trophic level.

 

Contamination of foliage with soil that has been treated with

phosphate fertilizer or exposed to substantial airborne deposition of

fluoride may constitute an important route of fluoride transfer to

large herbivores. Davison reported that, as of 1987, virtually 29

nothing was known about fluoride transport to invertebrates or

detritus chain organisms. However, one study on this topic was found
H-27

during this review and is discussed below in the section on toxicity

of fluoride.

As with human studies, animal studies of the relationship between

treatment with fluoride and bone strength have revealed a conflicting

picture. Whitford concludes that when considering the extensive

literature on the subject of fluoride and bone fractures, it appears

that fluoride has little or no effect on bone strength.38

A large amount of literature is dedicated to the subject of

fluoride toxicity to animals. Studies reviewed pertain to mice,

voles, isopods, rats, and rabbits. Several studies have been

conducted on the effects of fluoride, both alone and in conjunction

with other compounds/pollutants, on small mammals. Lead accumulation,

absorption, and assimilation can be affected by the presence of

various other dietary components, including fluoride. For instance,

in a study conducted by Cooke, Andrews, and Johnson, laboratory wood

mice (Apodemus sylvaticus) were fed solutions containing soluble salts

of lead, zinc, cadmium, and fluoride in their drinking water. 49

Fluoride concentrations showed their typical accumulation in femur

tissue. The highest mean Pb concentrations in liver, kidney, and

femur were found in the treatment comprised of all four elements. 

High levels of Pb and fluoride were found to thwart Zn’s antagonistic

effect on the accumulation of Cd in kidneys of wood mice. The authors

note that although other studies have found Pb to have a small but

significant antagonistic effect on bone fluoride accumulation, this

study did not. 

In the high Zn treatment, greater femur fluoride levels were

observed but not believed to be of any biological significance. The

treatment combining Pb, Zn, Cd, and fluoride reflected the dietary

intake of wild animals living in the grasslands established on

fluorspar wastes and yielded the highest Pb concentrations in all

three tissue types, the highest kidney and femur Cd levels, and the

highest kidney fluoride levels. The observed reversal of Zn

antagonism on Cd kidney levels in the treatment group exposed to all

four elements is important because of Cd’s renal toxicity. The

authors urge caution in extrapolating these lab results to the field

because the availability of soluble forms of these elements in field

situations is unknown. This research indicates the complexity of the

interactions among trace elements in small mammals.

Boulton, Cooke, and Johnson conducted an investigation into the

effects of fluoride toxicity on small mammals by examining populations

of two species of vole (Microtus agrestis L. and Clethrionomys

glareolus L.) at three different types of contaminated sites. 50

Vegetation was collected and analyzed at each site as a means of

estimating fluoride dose to resident voles. Fluoride uptake by new

vegetation was relatively low, even when hydrochloric acid was used to

mimic extraction by vole stomach acids. This suggests that because

fluoride solubility is moderate, so is assimilation by both small

mammals. Study results suggest that the order of assimilation of
H-28

dietary fluoride by location was: reference site < tailings dam <

smelter < chemical works. Fluoride concentrations in the teeth and

bones of both vole species were significantly higher at each of the

three contaminated sites than at the reference site. The chemical

plant and aluminum smelter both experienced a combination of hydrogen

fluoride gas and fluoride-laden dust deposition. Both of these sites

had considerably higher fluoride concentrations than the tailings dam. 

Fluorides taken up by leaf stomata as hydrogen fluoride gas

and/or from deposition on leaf surfaces clearly contributed to higher

water- and acid-soluble fluoride in vegetation at the atmospherically

contaminated sites. Both species of vole exhibited severe incisor and

molar dental lesions and had significantly higher tissue fluoride

concentrations at the chemical plant and aluminum smelter sites; the

damage at the mine tailings dam was less pronounced. The variation in

effects at different types of sites was believed to be the result of

differences in fluoride speciation and its impacts on fluoride

availability for bioassimilation. This study indicates that small

mammals can experience severe dental lesions at sites atmospherically

contaminated by fluoride and can be effective indicators of

environmental contamination.

Another possible impact that fluoride can have on terrestrial

animals is reduction of the vitality of the decomposer community and

its ability to carry out the decomposition function. Van Wensem and

Adema investigated fluoride’s impact on soil fauna-mediated

decomposition and survival and growth of isopods (as represented by

Porcellio scaber) and found neither to be adversely affected by soil

fluoride concentrations of up to 170 micro mol per gram. A previous 51

study by Buse found fluoride to accumulate in invertebrates. 52

However, in this study fluoride did have a significant adverse effect

on the extractable ammonium, nitrate, and phosphate concentrations of

a terrestrial micro-ecosystem and it was concluded that fluoride is

toxic to microbial processes at concentrations found at moderately

polluted sites. During weeks one through four of the study, the

respiration rate and total carbon dioxide production were

significantly increased, but only at the highest fluoride

concentration. The order of sensitivity of mineralization processes

to fluoride was found to be phosphate < ammonium < nitrate. The

no—observed—effects concentrations for net mineralization of ammonium,

nitrate, and phosphorus were 17, 5.3, and 53 micro mol fluoride per

gram dry weight of litter, respectively. Another study by Wang and

Bian found silkworms to be relatively sensitive to atmospheric

deposition of fluoride. High accumulation of substances in isopods 53

without concomitant toxicity, such as the authors found in this study,

has also been observed with heavy metals.

The combined effect of more than one pollutant on nutrient

bioavailability is another consideration when evaluating fluoride’s

impact on terrestrial animals. For example, Clerklewski and

Ridlington’s examination of the influence of dietary soluble (lead

acetate) and insoluble (lead carbonate) lead on fluoride
H-29

bioavailability in rats found depressed weight gain at both high and

low fluoride doses, and significant reductions in femur and second

molar fluoride at the higher fluoride dose. Fluoride failed to 54

influence the increased lead concentration in the plasma, femur,

liver, and kidney, as well as the increased excretion of deltaaminolevulinic acid. The overall conclusion is that dietary lead may

reduce fluoride bioavailability, but fluoride does not influence lead

utilization. Cerklewski and Ridlington were unable to determine if

the reduced fluoride bioavailability was due to decreased absorption

or retention. It is valuable to note this combined effect of dietary

fluoride and lead, as environmental pollutants are commonly found in

mixtures.

Susheela and Bhatnagar investigated fluoride’s toxicity to rabbit

teeth. The authors examined the effects of long-term fluoride 55

toxicity on rabbit teeth morphology and inorganic chemical (fluoride,

calcium, phosphorous) composition. Researchers found a significant

increase in fluoride and a significant decrease in calcium levels in

both the 18- and 23-month dose groups; however, the calcium/phosphate

ratio changed significantly only in the 23-month group. The scanning

electron microscope analysis showed hypoplastic, rough, uneven,

pitted, and cracked enamel covered with granular deposits. Overall,

the study shows that chronic fluoride ingestion causes changes in the

structural and biochemical constituents of rabbit tooth enamel. This

research is valuable because it considers the effects of chronic

exposure to a wildlife species since the exposure of wildlife to

environmental fluoride pollution may be chronic in nature.

Fluoride can sometimes ameliorate the toxicity of other

compounds. For instance, Pleasants, et al. explored the effects of

several vitamins and fluoride on cadmium (Cd ) toxicity in rats. 2+ 56

They found that fluoride reduced the weight depression associated with

cadmium toxicity and lowered the relative weights of rat testes. The

combined dose of fluoride and D hormone was found to significantly

increase the dry and mineral femur weights of cadmium-exposed rats,

which may be of interest in the treatment of osteoporosis. A group of

rats treated with fluoride in conjunction with vitamins A and 1,25-

dihydroxyvitamin D appeared to have experience reduced Cd 3 2+

hematoxicity. In another study, Yu, et al. examined whether high

dietary fluoride inhibits selenite toxicity in rats. The significant 57

effect shown in this experiment was the use of fluoride to prevent Se

liver pathology. However, other considerations such as Se

concentrations in plasma and kidney, depressed weight gain, and the

enzymatic activity of glutathione peroxidase and xanthine oxidase were

unaffected by fluoride. The authors suggest the most likely mechanism

for fluoride protection from Se liver pathology is the formation of an

insoluble Ca and F complex, which prevents excess intracellular Ca

(produced as an effect of Se toxicity) from activating phospholipases

and proteases. The possibility of fluoride use in the prevention of

Se toxicity may be of interest to EPA; however, the results of this

paper are not viewed as substantial enough to make the claim that

fluoride dosage is beneficial for this purpose. Further research
H-30

would be necessary to determine with certainty whether fluoride would

be useful for this purpose.

Zeiger, et al. conducted an extensive review of fluoride genetic

toxicity publications, presenting many articles both supporting and

denying genetic toxicity. The review covers microbial, in vitro 58

mammalian, and in vivo mammalian test systems. The review of in vitro

tests includes assays for mutation, chromosome aberrations, sister

chromatid exchanges, DNA damage/repair, and cell transformation for

both animal and human cell lines. The in vivo tests reviewed were

concerned with the somatic and germ cells of various rodents. The

article states that, based on the conflicting evidence from many of

the assays listed above, the mechanisms of fluoride toxicity are

purely speculative. In an attempt to summarize the conflicting

evidence, the authors find that the fluoride ion is not mutagenic in

standard bacterial systems but can cause some of the lesions examined

in vitro. Furthermore, the issue of in vivo genetic toxicity should

be considered unresolved. This review leaves the question of the

genetic toxicity of fluoride open for interpretation. The authors

provide many published assays containing conflicting evidence in

similar or identical test systems, which means that interpretation

must be based on weight of evidence considerations.

Davison reports that the first detectable signs of fluorosis

occur in animals consuming herbage exposed to airborne fluoride

concentrations in excess of approximately 0.3 g/m to 0.5 g/m for a 3 3

continuous period of several months. If higher concentrations are

present, more severe fluorosis symptoms will occur more rapidly.29

Whitford’s summary of a recent USPHS report on fluoride concludes

that data from animal studies have not established an association

between fluoride exposure, even extremely high and life-long exposure,

and cancer.38

To summarize, for absorbed fluoride, kidney excretion is likely a

minimal removal pathway as compared to deposition in bones and teeth. 

Differential absorption and retention observed in metabolic cage

studies of mice and voles was attributed to different intake levels. 

Bone fluoride concentrations were found to be greater in male than in

female raptors. Conflicting information was found concerning whether

fluoride accumulates in food chains. Variations in fluoride

concentrations within a given species of plant can result in animal

species with differing feeding niches ingesting different amounts of

fluoride. Contamination of foliage with soil may constitute an

important route of fluoride transfer to large herbivores in situations

where soil has been treated with phosphate fertilizer or exposed to

substantial airborne deposition of fluoride. It appears that fluoride

has little or no effect on bone strength. Voles living near a mine

tailings dam, an aluminum smelter, and a chemical plant were found to

exhibit severe incisor and molar dental lesions and significantly

higher tissue fluoride concentrations at the chemical plant and

aluminum smelter sites, than at the mine tailings dam. High soil
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fluoride concentrations were not found to adversely affect soil faunamediated decomposition or the survival and growth of isopods, but were

found to be toxic to microbial processes at concentrations found at

moderately polluted sites. Limited evidence was found that dietary

lead may reduce fluoride bioavailability, but fluoride does not

influence lead utilization. Fluoride can sometimes ameliorate the

toxicity of other compounds. Limited evidence also exists showing

that high dietary fluoride inhibits selenite toxicity in rats. The

subject of fluoride’s genetic toxicity remains controversial. Animal

consumption of foliage exposed to 0.3 g/m to 0.5 g/m concentrations 3 3

of fluoride for a continuous period of several months may cause

fluorosis. Finally, data from animal studies have not established an

association between fluoride exposure, even extremely high and lifelong exposure, and cancer.

H.3.9 HF Aquatic Animal Impacts

According to Slooff, et al., bioconcentration factors of less

than 10 have been calculated for crustaceans and fish exposed to

fluoride concentrations of up to 50 mg/L. Retained fluoride is 30

usually stored in skeletal structures (e.g., bones and shells), with

the lowest fluoride levels observed in muscle tissues. Based on

limited data, it was concluded that biomagnification is negligible to

very slight.

Fluoride aquatic toxicity literature pertaining to trout, benthic

macroinvertebrates, minnows, water fleas, and diatoms was reviewed for

this study. Literature on both hard and soft water environments, as

well as both acute and chronic toxicity, was investigated as

available. 

In the first of two studies by Camargo and Tarazona examining the

effects of fluoride ion (F ) on aquatic species, two species of trout -

exposed to NaF in soft water were found to have 120-, 144-, 168-, and

192-hour LC values of 92.4, 85.1, 73.4 and 64.1 ppm F (rainbow trout) 50 -

and 135.6, 118.5, 105.1, and 97.5 ppm F (brown trout), respectively. - 59

Subject fish in fluoride-containing aquaria exhibited

hypoexciteability, darkened backs, and a decrease in respiration

before death. Trout mortality increased as fluoride concentrations

and exposure times increased. Rainbow trout fingerlings were found to

be significantly more sensitive to F than brown trout fingerlings of -

similar age and weight. The higher resistance of brown trout to

fluoride may be due to their greater physiological ability to inhibit

the toxic action of fluoride on groups of enzymes within cells or

through removing or immobilizing fluoride ions more effectively. Both

trout species appear to be more resistant to fluoride than freshwater

benthic macroinvertebrates. The authors suggest the possibility that

fluoride ions form stable complexes with calcium in the blood and

bones of fish, but not in freshwater insect larvae, as one explanation

for the lesser sensitivity of fish to fluoride. Freshwater fish may

be more resistant to high fluoride concentrations in hard water than

in soft water. It is possible that the reservoir of calcium that
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surrounds fish in hard water may compensate for the loss of calcium,

thereby delaying fluoride’s toxic effects.

Fluoride concentrations in sea water typically range from 1.2 to

1.4 mg/L, while most fresh waters contain less than 0.2 mg/L of

fluoride. Freshwater fluoride concentrations are believed to be

rising due to industrial pollution. In a second study of the acute

toxicity of fluoride to aquatic species, Camargo and Tarazona found

that five species of freshwater benthic macroinvertebrates (Chimarra

marginata Linnaeus, Hydropsyche lobata MacLachlan, Hydropsyche

bulbifera MacLachlan, Hydropsyche exocellata Dufour, and Hydropsyche

pellucidula) were found to be more sensitive to fluoride than the

aquatic species previously investigated in other studies. These five 60

freshwater benthic macroinvertebrates, all native to the rivers of the

Iberian Peninsula, were found to have 96-hour LC values of 44.90, 50

48.20, 26.30, 26.50, and 38.50 mg/L F , respectively. -

Macroinvertebrate mortality was observed to increase as fluoride

concentrations increased. The sensitivity of the five species

investigated indicates their possible usefulness as aquatic fluoride

bioindicators in soft water environments. The more sensitive species

investigated might, therefore, be useful in setting fluoride water

quality criteria in fresh water.

Khan, et al. used a solution containing 1.5 percent

hydrofluozirconic acid and 1.0 percent of ammonium bifluoride to

determine the chronic toxicity of a fluoride mixture to the fathead

minnow (Pimephales promelas) and water flea (Ceriodaphnia dubia). 61

Three toxicity tests were conducted, two with the solution at a pH of

2.5 and a third with pH adjusted to 7.5. The endpoints used were

survival, reproduction, and/or growth. After a 7-day exposure period,

fathead fry had an LC of 2.26 percent, a survival NOEC of 1.25 50

percent, and a growth NOEC of 0.625 percent. After an 8-day exposure

period, water fleas had an LC of 3.11 percent, a survival NOEC of 2.5 50

percent, and a reproduction NOEC of 0.625 percent. When the

solution’s pH was adjusted to 7.5, fathead minnows had a 7-day LC , a 50

survival NOEC, and a growth NOEC all of <6.25 percent. The article

does not state why these particular solution components were chosen. 

Although the adjustment of pH from 2.5 to 7.5 effectively ruled out

acidity as the causal agent in the observed mortality, the presence of

zirconium ions in the solution precluded a definitive assessment of

whether the remaining mortality effect was due to fluoride or

zirconium ions.

Joy and Balakrishnan studied the effects of fluoride ion on

cultures of diatoms Nitzschia palea (freshwater) and Amphora

coffeaeformis (brackishwater) in a laboratory setting. The number of 62

cells of both organisms and the amount of chlorophyll a and c were

increased, with greater concentrations of fluoride. Statistical

analysis showed that fluoride concentrations above 10 mg/L resulted in

significant stimulation of growth in terms of cell number and

chlorophyll in all tested concentrations. The results indicate that

these diatoms can tolerate and are stimulated to grow by high fluoride
H-33

concentrations; however, the ecological significance of this would be

varied because of fluoride’s propensity to transfer easily to animal

populations. Therefore, the impact of fluoride pollution in

waterbodies is still to be viewed with caution.

Slooff, et al. report that field studies of the toxicity of

fluoride-containing effluent to aquatic animals resulted in effects on

both abundance and diversity of estuarine/marine organisms at

relatively low fluoride levels. However, it was unclear to what 30

extent other effluent parameters such as pH and other substances were

affecting the organisms. 

To summarize, biomagnification in aquatic animals is reported to

be negligible to very slight. Two trout species were demonstrated to

be more resistant to fluoride than freshwater benthic

macroinvertebrates. Diatoms appear to be tolerant of, and stimulated

to grow by, high fluoride concentrations, with uncertain ecological

significance. Finally, limited evidence exists for fluoridecontaining effluent effects on both abundance and diversity of

estuarine/marine organisms at relatively low fluoride levels.

H.4 FUTURE RESEARCH

The following topics, discovered in the process of this research,

were found either not to be covered in the available literature, or to

be unresolved to some degree. Further research on these topics can be

expected to yield information important to gaining a better

understanding of the impacts of HCl and HF emissions from utilities.

1. What are HCl’s impacts on the spatial and temporal deposition of

SO , a species implicated in both acid rain and PM ? x 2.5

SO and HCl interact in the atmosphere. For instance, chloride 2

ions can catalyze the oxidation of sulfite to sulfate when

present in concentrations typical of plumes containing HCl. The

net effect of this change is that less SO will be dry-deposited 2

close to the source, but more H SO will be generated in cloud and 2 4

rain water, subject to further transport. HCl might, therefore,

alter the spatial and temporal deposition of SO , which may be an 2

important way in which utility HCl emissions indirectly affect

acid rain and PM . 2.5

2. Do HCl and/or HF have an impact on the speciation and deposition

of other HAPs?

Atmospheric HCl can affect the atmospheric chemistry of other

HAPs, such as mercury. In a theoretical study based on their

chemical kinetic model, Selgneur, Wrobel, and Constantinou found

that HCl concentration, among other factors (e.g., atmospheric

liquid water content, pH, and SO concentration), had a strong 2

effect on both the Hg(0)/Hg(II) ratio and the gas/liquid

partitioning of Hg(0). 14
H-34

Although the effect of this change may be considered

positive—shifting more mercury to a less toxic form—this might

not always be the case with mercury or with other HAPs emitted by

utilities. 

3. What percentage of ambient HCl and HF particulate is in the fine

fraction?

EPA has promulgated a primary 24-hour PM standard of 65 2.5 g/m 3

and a primary annual PM standard of 15 2.5 g/m . As part of the 3

effort to meet these standards, it will be important to know what

the contribution of HCl and HF are to ambient PM2.5

concentrations.

The literature evaluated for this paper contained several

references to ambient HCl and HF concentrations, both rural and

urban. Urban HCl concentrations were found to reach as high as 4

g/m . Information on rural HF concentrations was more scarce, 3

with a reported high end concentration of 1 g/m . However, 3

accidental industrial releases of HF can produce short-term

concentrations of up to 50 g/m . Davison reports fluoride 3

particulates range in diameter from <0.1 m to approximately 10

m. If a significant portion of ambient HCl and HF are in the

fine fraction, they could conceivably contribute to PM2.5

exceedances.

4. Are HCl and HF capable of being transported large distances? How

far are they typically transported and in what regions, if any,

do they reach sensitive receptors?

Several references to HCl and HF transport were found. A key

question is whether HCl/HF-enriched air parcels reach acid rain

or PM -sensitive receptors and if so, how long this process 2.5

takes. The literature reports atmospheric HCl lifetimes of

between 1 and 5 days and HF lifetimes of 0.54 to 5 days. Future

research might be directed at clarifying the extent of

atmospheric transport of these two species.

5. What are the interactive effects of fluoride and other

pollutants?

Interactions between trace elements in small mammals can be

complex. For example, Cooke, Andrews, and Johnson observed that

high concentrations of one or more pollutants can suppress

another pollutant’s antagonism of a third pollutant’s toxicity,

such as in the case of fluoride, Pb, Zn, and Cd with wood mice. 49

And although the evidence is far from conclusive, fluoride has

been shown in at least one case to prevent Se toxicity.

The combined effect of more than one pollutant on nutrient

bioavailability is another consideration when evaluating

fluoride’s impact on terrestrial animals. Cerklewski and
H-35

Ridlington found that dietary Pb may reduce fluoride

bioavailability, but that fluoride does not influence Pb

utilization. Further examination of HF and HCl emissions may be 54

warranted to elucidate subtle ecosystem effects.

6. Are ambient concentrations of fluoride toxic to microbial

processes?

Van Wensem and Adema found that fluoride had a significant

adverse effect on the extractable ammonium, nitrate, and

phosphate concentrations of a terrestrial micro-ecosystem, and

concluded that fluoride is toxic for microbial processes at

concentrations found at moderately polluted sites. If other 51

research corroborates these findings, some elements of the

detritus cycle may be inhibited at such sites.

7. Are Al/F complexes phytotoxic?

The most recent information found (1996) states that the question

of whether Al/F complexes are phytotoxic has not been clarified. 

Bond, et al. found that fluoride forms strong complexes with Al

in soil. Klumpp, et al. found strong indications that the 36

resistance of some tropical tree species may be related to their

capacity to accumulate Al. 42

8. Does fluoride accumulate with trophic level?

Conflicting information was found concerning whether fluoride

accumulates in food chains. Slooff, et al. report (1989) that

exposure to high concentrations of fluoride causes terrestrial

organisms to accumulate more fluoride than when exposed to

background levels, with both invertebrate and vertebrate

predators accumulating higher levels of fluoride than

herbivores. Davison, however, reports (1987) that although data 30

are limited, it is generally assumed that fluoride concentrations

do not increase with trophic level. Resolving the question of 29

biomagnification is essential to evaluating the ecological

impacts of HF and HCl.

9. Information on the following HCl topics was not available or not

found. 

Information pertaining to several topics concerning HCl was

scarce. The following topics, in addition to those listed above,

would be good candidates for future research: 

1. Terrestrial stability and persistence of HCl and byproducts; 

2. terrestrial HCl transport, mobility, and partitioning; 

3. aquatic stability and persistence of HCl and by-products;
H-36

4. terrestrial animal HCl ingestion, inhalation, dermal

absorption, distribution, accumulation; and

5. aquatic animal HCl ingestion, distribution, and

accumulation.

H.5 GLOSSARY

Biomagnification - The tendency of a substance to accumulate

through a food chain. Occurs when substances

stored in the tissues of a large number of

organisms at lower trophic levels are taken in

and stored by a predator at a higher trophic

level.

Chlorosis - A symptom of disease or disorder in plants,

which involves a reduction in or loss of the

normal green coloration. Affected plant(s)

will be pale green or even yellow. Chlorosis

is caused by conditions that prevent the

formation of chlorophyll.

Fluorosis - A disease that results from the ingestion of

fluorine in amounts that substantially exceed

bodily requirements. In excess, fluorine leads

to the thickening of bones, sometimes to the

extent that joints stiffen and the increased

weight of the skeleton makes it difficult for

the head to be raised. Teeth may also be

softened and stained.

Hornblende - A mineral, CaNa(Mg,Fe) (Al,Fe,Ti) Si O (OH,F) , 4 3 6 22 2

commonly green to black in color, formed late

in the cooling of igneous rock.

Interspecific - Arising or occurring between species.

Solubility product - The product of the molar concentrations of the

constituent ions, each raised to the power of

its stoichiometric coefficient in the

equilibrium equation.
H-37

1. Ruprecht, H. and Sigg, L. Interactions of Aerosols (Ammonium

Sulfate, Ammonium Nitrate, and Ammonium Chloride) and of Gasses

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No. 3. 1990. p. 573-584.

2. Wegner, et al. Variations of tropospheric HCl amounts over

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8. 1997. p. 849-852.

3. Puxbaum, et al. Seasonal variation of HNO , HCl, SO , NH and PM 3 23

at a rural site in NE Austria. Atmospheric Environment. Volume

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4. Graedel, T. E. and W. C. Keene. Tropospheric Budget of Reactive

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5. Lightowlers, P. J. and J. N. Cape. Sources and Fate of

Atmospheric HCl in the U.K. and Western Europe. Atmospheric

Environment. Volume 22., No. 1. 1988. p. 7-15.

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Research Triangle Park, NC. July 1993.

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H-38

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19. Peters, N. E. Chloride Cycling in Two Forested Lake Watersheds

in the West-Central Adirondack Mountains, New York, U.S.A. 

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20. Norton, S. A., Handley, M. J., Kahl, J. S., and Peters, N. E. 

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DOC. Water, Air & Soil Pollution. Volume 91. 1996. p. 283-

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21. Zahn, M. T. and W. D. Grimm. Nitrate and Chloride Loadings as

Anthropogenic Indicators. Water, Air, & Soil Pollution. Volume

68. 1993. p. 468-483.

22. World Health Organization and United Nations Environment

Programme. Chlorine and Hydrogen Chloride. Environmental Health

Criteria. Volume 21. Geneva. 1982. p. 1-81.
H-39

23. National Research Council. Lab Chemical Safety Summaries:

Hydrochloric Acid and Hydrogen Chloride, and Hydrogen Fluoride

and Hydrofluoric Acid. Obtained on the Internet at

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24. McWilliams, E. L. and R. L. Sealy. Atmospheric Chloride: Its

Implications for Foliar Uptake and Damage. Atmospheric

Environment. Volume 21, No. 12. 1987. p. 2661-2665.

25. Tukey, H. G., H. G. Tukey, and S. H. Wittwer. Loss of Nutrients

by Foliar Leaching Determined by Radioisotopes. American Society

for Horticultural Science. Volume 71. 1958. p. 496-506.

26. Manning, T.M., S.P. Wilson, and J.C. Chapman. Toxicity of

Chlorine and Other Chlorinated Compounds to Some Australian

Aquatic Organisms. Bulletin of Environmental Contamination and

Toxicology. Volume 56. 1996. p. 971-976.

27. Hance, C. D., P. A. Solomon, L. G. Salmon, T. Fall, G. R. Cass. 

Hydrofluoric Acid in the Southern California Atmosphere. 

Environmental Science & Technology. Volume 31, No. 4. 1997. p.

956-959.

28. Winiwarter, E., et al. Tracers of Coal-Fired Power Plants in the

Lake Powell Region. In: Transactions: Visibility and Fine

Particles. Edited by Mathai, C. V. AWMA, Pittsburgh. 1989. p.

885-897.

29. Davison, A. W. Pathways of Fluoride Transfer in Terrestrial

Ecosystems. In: Pollutant Transport and Fate in Ecosystems. 

1987. Blackwell Scientific Publications, Boston, MA. p. 193-

210.

30. Slooff, W., et al. Integrated Criteria Document Fluorides.

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Biltoven, The Netherlands. Report No. 758474010. 1989.

31. U.S. Environmental Protection Agency. Hydrogen Fluoride Study,

Report to Congress, Section 112(n)(6) Clean Air Act As Amended,

Final Report. EPA/550/R-93-001. Chemical Emergency Preparedness

and Prevention Office. September 1993.

32. Eatough, et al. Apportionment of Sulfur Oxides at Canyonlands

During the Winter of 1990-1. Study Design and Particulate

Chemical Composition. Atmospheric Environment. Volume 30, No.

2. 1996. p. 269-281.
H-40

33. Leece, D. R. and J. H. Scheltema. The Use of Leaf Analysis to

Diagnose Fluoride Toxicity in Grapes. In: Fluoride Emissions: 

Their Monitoring and Effects on Vegetation and Ecosystems.

Edited by F. Murray. Academy Press Australia, New South Wales. 

1982. p. 101-123.

34. Arocena, et al. Weathering of Clay Minerals Induced by FluorideContaining Solutions from Phosphogypsum By-Product. Canadian

Journal of Soil Science. 1995. p. 219-226.

35. Murray, F. Ecosystems as Sinks for Atmospheric Fluoride. 

Fluoride Emissions: Their Monitoring and Effects on Vegetation

and Ecosystems. Edited by F. Murray. Academy Press Australia,

New South Wales. 1982. p. 191-205.

36. Bond, W. J., et al. The Effect of Sulfate and Fluoride on the

Mobility of Aluminum in Soil. Australian Journal of Soil

Research. Volume 33. 1995. p. 883-897.

37. Elrashidi, M. A., and W. L. Lindsay. Solubility of Aluminum

Fluoride, Fluorite, and Fluoriphlogopite Minerals in Soils. 

Journal of the Soil Science Society of America. Volume 50. 

1986. p. 594-598.

38. Whitford, G. M. The Metabolism and Toxicity of Fluoride (Revised

edition). Karger. 1996.

39. Hodge, H. C. and F. A. Smith. Fluorine Chemistry, Volume IV -

Biological Properties of Inorganic Fluorides. Edited by J. H.

Simons. Academic Press, NY, NY. 1965.

40. Dukes, M. N. G. Fluoride. Dukes Side Effects of Drugs, Annual

4. 1980.

41. World Health Organization. Environmental Health Criteria 36: 

Fluorine and Fluorides. Geneva. 1984.

42. Klumpp, et al. Fluoride Impact on Native Tree Species of the

Atlantic Forest Near Cubatao, Brazil. Water, Air, and Soil

Pollution. Volume 87. 1996. p. 57-71.

43. Davies, F. B. M. and G. Notcutt. Accumulation of Fluoride by

Lichens in the Vicinity of Etna Volcano. Water, Air and Soil

Pollution. Volume 42. 1988. p. 365-371.

44. Amundson, R. G., A. J. Belsky and R. C. Dickie. Fluoride

Deposition via Litterfall in a Coastal-Plain Pine Plantation in

South Carolina. Water, Air, and Soil Pollution. Volume 50. 

1990. p. 301-310.
H-41

45. Weinstein, L. H. and R. Alscher-Herman. Physiological Responses

of Plants to Fluoride. In: Effects of Gaseous Pollutants in

Agriculture and Horticulture. Edited by Unsworth, M. H. & D. P.

Ormrod. Butterworth Scientific, Boston, MA. 1982. p. 139-167.

46. Boulton, I. C., J. A. Cooke, and M. S. Johnson. Experimental

Fluoride Accumulation and Toxicity in the Short-Tailed Field Vole

(Microtus agrestis). Journal of Zoology. Volume 234. 1994. p.

409-421.

47. Boulton, I. C., J. A. Cooke, and M. S. Johnson. Fluoride

Accumulations and Toxicity in Laboratory Populations of Wild

Small Mammals and White Mice. Journal of Applied Toxicology.

Volume 15, No. 6. 1995. p. 423-431.

48. Seel, D. C., A. G. Thomson, and R. E. Bryant. Bone Fluoride in

Four Species of Predatory Bird in the British Isles. Pollutant

Transport and Fate in Ecosystems. Blackwell Scientific

Publications, Boston, MA. 1987. p. 211-221.

49. Cooke, J. A., S. M. Andrews and M. S. Johnson. The Accumulation

of Lead, Zinc, Cadmium and Fluoride in the Wood Mouse (Apodemus

sylvaticus L.) Water, Air, & Soil Pollution Volume 51. 1990. p.

55-63.

50. Boulton, Cooke, and Johnson. Lead, Zinc, Cadmium, and Fluoride

in Small Mammals from Contaminated Grassland Established on

Fluorspar Tailings. Water, Air, and Soil Pollution. Volume 51,

No. 1-2. 1990. p. 43-54.

51. Van Wensem, J., and T. Adema. Effects of Fluoride on Soil Fauna

Mediated Litter Decomposition. Environmental Pollution. Volume

72. 1991. p. 239-251.

52. Buse, A. Fluoride Accumulation in Invertebrates Near an Aluminum

Reduction Plant in Wales. Environmental Pollution. Volume 41A. 

1986. p. 199-217.

53. Wang, J., and Y. Bian. Fluoride Effects on the Mulberry-Silkworm

System. Environmental Pollution. Volume 52. 1988. p. 11-18.

54. Cerklewski, F. L., and J. W. Ridlington. Influence of Dietary

Lead on Fluoride Bioavailability in the Rat. Biological Trace

Element Research. Volume 14. 1987. p. 105-113.

55. Susheela, A. K., and M. Bhatnagar. Fluoride Toxicity: A

Biochemical and Scanning Electron Microscopic Study of Enamel

Surface of Rabbit Teeth. Archives of Toxicology. Volume 67. 

1993. p. 573-579.
H-42

56. Pleasants, E. W., C. Waslien, and B. A. Naughton. Dietary

Modulation of the Symptoms of Cadmium Toxicity in Rats: Effects

of Vitamins A, C, D, D Hormone, and Fluoride. Nutrition

Research. Volume 13. 1993. p. 839-850.

57. Yu, et al. Effect of Dietary Fluoride on Selenite Toxicity in

the Rat. Biological Trace Element Research. Volume 34. 1992. 

p. 265-278.

58. Zeiger, et al. Genetic Toxicity of Fluoride. Environmental and

Molecular Mutagenesis. Volume 21. 1993. p. 309-318.

59. Camargo, J. A. and J. V. Tarazona. Short-Term Toxicity of

Fluoride Ion (F ) in Soft Water to Rainbow Trout and Brown Trout. -

Chemosphere. Volume 22, Nos. 5-6. 1991. p. 605-611.

60. Camargo, J. A. and J. V. Tarazona. Acute Toxicity to Freshwater

Benthic Macroinvertebrates of Fluoride Ion in Soft Water. 

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45. 1990. p. 883-887.

61. Khan, A., D. Kent, J. Barbieri, and S. Khan. Chronic Toxicity of

a Fluoride Mixture to Freshwater Organisms. Water Science and

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1990. p. 241-249.
Appendix I - Mercury Control Technologies

This Appendix provides information extracted from the Mercury

Study Report to Congress, Volume VIII: An Evaluation of Mercury

Control Technologies and Costs. EPA-452/R-97-010. December 1997 and

is provided for the readers information. Appendix I contains text

elements provided by the Department of Energy regarding resent

research on mercury controls for electric utility steam generating

units.
x
I-1

I.1 INTRODUCTION

I.1.1 Coal Cleaning

Coal cleaning is an option for removing mercury from the fuel

prior to combustion. In some States, certain kinds of coal are

commonly cleaned to increase their quality and heating value. 

Approximately 77 percent of the Eastern and Midwestern bituminous coal

shipments are cleaned in order to meet customer specifications for

heating value, ash content, and sulfur content.

There are many types of cleaning processes, all based on the

principle that coal is less dense than the pyritic sulfur, rock, clay,

or other ash-producing impurities that are mixed or embedded in it. 

Mechanical devices using pulsating water or air currents can

physically stratify and remove impurities. Centrifugal force is

sometimes combined with water and air currents to aid in further

separation of coal from impurities. Another method is dense media

washing, which uses heavy liquid solutions usually consisting of

magnetite (finely ground particles of iron oxide) to separate coal

from impurities. Smaller sized coal is sometimes cleaned using froth

flotation. This technique differs from the others because it focuses

less on gravity and more on chemical separation.

Some of the mercury contained in coal may be removed by coal

cleaning processes. Volume II of the Mercury Report (An Inventory of

Anthropogenic Mercury Emissions in the United States) presents

available data on the mercury concentrations in raw coal, cleaned coal

and the percent reduction achieved by cleaning. These data, which

cover a number of different coal seams in four States (Illinois,

Pennsylvania, Kentucky, and Alabama), indicate that mercury reductions

range from 0 to 64 percent, with an overall average reduction of 21

percent. This variation may be explained by several factors,

including different cleaning techniques, different mercury

concentrations in the raw coal, and different mercury analytical

techniques. 

It is expected that significantly higher mercury reductions can

be achieved with the application of emerging coal preparation

processes. For example, in one bench-scale study, five types of raw

coal were washed by conventional cleaning methods followed by column

froth floatation or selective agglomeration. Conventional cleaning

and column froth flotation reduced mercury concentrations from the raw

coals by 40 to greater than 57 percent, with an average of 55 percent. 

Conventional cleaning and selective agglomeration reduced mercury

concentrations from the raw coals by greater than 63 percent to 82

percent, with an average of 68 percent. In a second bench-scale study

in which three types of coals were cleaned with a heavy-media-cyclone

(a conventional cleaning method) followed by a water-only-cyclone and

a column froth flotation system, mercury concentrations in the raw

coal were reduced by as much as 63 to 65 percent. Bench-scale testing

is also being carried out by DOE to investigate the use of naturally
I-2

occurring microbes to reduce mercury (and other trace elements) from

coal.

Any reduction in mercury content achieved by coal cleaning

results in a direct decrease in mercury emissions from boilers firing

cleaned coals. The mercury removed by cleaning processes is

transferred to coal-cleaning wastes, which are commonly in the form of

slurries. No data are available to assess the emissions of mercury

from coal-cleaning slurries. 

I.1.2 Flue Gas Treatment Technologies

Most metals have sufficiently low vapor pressures at typical air

pollution control device operating temperatures that condensation onto

particulate matter is possible. Mercury, on the other hand, has a

high vapor pressure at typical control device operating temperatures,

and collection by particulate matter control devices is highly

variable. Factors that enhance mercury control are low temperature in

the control device system (less than 150C [300 to 400F]), the

presence of an effective mercury sorbent, and a method to collect the

sorbent. In general, high levels of carbon in the fly ash enhance

mercury sorption onto particulate matter which is subsequently removed

by the particulate matter control device. Additionally, the presence

of hydrogen chloride (HCl) in the flue gas stream can result in the

formation of mercuric chloride (HgCl ), which is readily adsorbed onto 2

carbon-containing particulate matter. Conversely, sulfur dioxide

(SO ) in flue gas can act as a reducing agent to convert oxidized 2

mercury to elemental mercury, which is more difficult to collect.

Add-on controls to reduce mercury emissions are described in

detail in this appendix, including information on commercial status,

performance, applicability to the specified mercury emission sources,

and secondary impacts and benefits. The controls described are:

Wet scrubbing;

Treated activated carbon adsorption; and

Activated carbon injection.

The most important conclusions from the assessment of flue gas

treatment technologies include:

Factors that enhance mercury control are low temperature in

the control device system (less than 150C [300 to 400F]),

the presence of an effective mercury sorbent and a method to

collect the sorbent. In general, high levels of carbon in

the fly ash enhance mercury sorption onto particulate matter

which is subsequently removed by the particulate matter

control device. Additionally, the presence of HCl in the

flue gas stream can result in the formation of HgCl , which 2

is readily adsorbed onto carbon-containing particulate

matter, so it can be efficiently scrubbed by a wet FGD

system. Conversely, sulfur dioxide (SO ) in flue gas can 2
I-3

act as a reducing agent to convert oxidized mercury to

elemental mercury, which is more difficult to collect.

Control technologies designed for control of pollutants

other than mercury (e.g., acid gases and particulate matter)

vary in their mercury-removal capability, but in general

achieve reductions no greater than 50 percent (except for

high removal efficiencies for HgCl by wet scrubbers). 2

Selenium filters are a demonstrated technology in Sweden for

control of mercury emissions from lead smelters. Carbon

filter beds have been used successfully in Germany for

mercury control on utility boilers and MWCs. These

technologies have not been demonstrated in the U.S. for any

of these source types.

Injection of activated carbon into the flue gas of MWCs and

MWIs can achieve mercury reductions of at least 85 percent. 

The addition of activated carbon to the flue gas of these

source types would not have a significant impact on the

amount of particulate matter requiring disposal.

No full-scale demonstrations of mercury controls have been

conducted in the U.S. for utility boilers. Based on limited

pilot-scale testing, activated carbon injection provides

variable control of mercury for utility boilers (e.g., the

same technology might capture 20 percent of the mercury at

one plant and 80 percent at another). The most important

factors affecting mercury control on utility boilers include

the flue gas volume, the flue gas temperature and chloride

content, the mercury concentration, and the chemical form of

mercury being emitted. 

The chemical species of mercury emitted from utility boilers

vary significantly from one plant to another. Removal

effectiveness depends on the species of mercury present. To

date, no single control technology has been identified that

removes all forms of mercury.

The addition of activated carbon to utility flue gas for

mercury control would significantly increase the amount of

particulate matter requiring disposal.

I.2 MERCURY CONTROLS

This section provides information on mercury controls that

provide opportunities for significant further reductions of mercury

emissions. Two major types of control techniques are described:

Coal cleaning; and

Flue gas treatment technologies.
I-4

I.2.1 Coal Cleaning

Approximately 77 percent of the Eastern and Midwestern bituminous

coal shipments are cleaned to meet customer specifications for heating

value, ash content and sulfur content (Akers et al, 1993). Along with

removing ash-forming and sulfur-bearing minerals, coal cleaning can

also reduce the concentration of many trace elements, including

mercury. 

Conventional coal cleaning methods are based on the principle

that coal is lighter than the pyritic sulfur, rock, clay, or other

ash-producing impurities that are mixed or embedded in it. Mechanical

devices using pulsating water or air currents can physically stratify

and remove impurities. Centrifugal force is sometimes combined with

water and air currents to aid in further separation of coal from

impurities. Another method, dense media washing, uses heavy liquid

solutions usually consisting of magnetite (finely ground particles of

iron oxide) to separate coal from impurities. 

Volume II of the Mercury Study Report to Congress (An Inventory

of Anthropogenic Mercury Emissions in the United States) presents

available data on the mercury concentrations in raw coal and cleaned

coal, as well as the percent reduction achieved by conventional coal

cleaning methods. These data, which cover a number of different coal

seams in four states (Illinois, Pennsylvania, Kentucky, and Alabama),

indicate that mercury reductions range from 0 to 64 percent, with an

overall average reduction of 21 percent. This variation may be

explained by several factors, including different cleaning techniques,

different mercury concentrations in the raw coal and different mercury

analytical techniques.

I.2.1.1 Advanced Coal Cleaning. Advanced coal cleaning methods

such as selective agglomeration and column froth flotation have the

potential to increase the amount of mercury removed by conventional

cleaning alone. In one bench-scale study, five types of raw coal were

washed by conventional cleaning methods followed by column froth

flotation or selective agglomeration. Conventional cleaning and

column froth flotation reduced mercury concentrations from the raw

coals by 40 to greater than 57 percent, with an average of 55 percent

(Smit, 1996). Column froth flotation reduced mercury concentrations

remaining in the washed coals by 1 to greater than 51 percent, with an

average of 26 percent (Smit, 1996). Conventional cleaning and

selective agglomeration reduced mercury concentrations from the raw

coals by greater than 63 percent to 82 percent, with an average of 68

percent (Smit, 1996). Selective agglomeration reduced mercury

concentrations remaining in the washed coals by greater than 8 percent

to 38 percent, with an average of 16 percent (Smit, 1996). 

In a second bench-scale study, three types of coals were cleaned

by a heavy-media-cyclone (a conventional cleaning method) followed by

a water-only-cyclone and a column froth flotation system. The heavymedia-cyclone reduced mercury concentrations in the raw coal by 42 to

45 percent (ICF Kaiser Engineers, 1995). The water-only-cyclone and
I-5

column froth flotation system reduced the concentrations of mercury

remaining in the cleaned coals by 21 to 23 percent (ICF Kaiser

Engineers, 1995). The combined reduction in mercury concentrations

from the coals ranged from 63 to 65 percent (ICF Kaiser Engineers,

1995). 

Bench-scale testing is also being carried out by DOE to

investigate the use of naturally occurring microbes to reduce the

mercury (and other trace elements) from coal.

Any reduction in mercury content achieved by coal cleaning

results in a direct decrease in mercury emissions from utility boilers

firing cleaned coals. The mercury removed by cleaning processes is

transferred to coal-cleaning wastes, which are commonly in the form of

slurries. No data are available to assess the emissions of mercury

from coal-cleaning slurries.

While advanced cleaning technologies can reduce mercury from the

coal (30 to greater than 60 percent) the potential impact on postcombustion form and control of the remaining mercury has not been

thoroughly investigated. Mercury mass transfer limitations are

encountered in emissions control systems on furnaces firing raw or

conventionally cleaned coals. Advanced coal-cleaning may exacerbate

this problem. In addition, chemical cleaning techniques being

considered may provide a coal that yields a different form of mercury

under combustion and post-combustion conditions. This could

adversely impact the natural mercury capture of the fly ash and across

wet/dry flue gas desulfurization (FGD) systems. There needs to be

more laboratory, bench-, and pilot-scale combustion and subsequent

post-combustion studies to evaluate these potential impacts. In

addition, the added costs for advanced coal cleaning separately and in

combination with post-combustion controls for mercury have not been

fully developed.

I.2.1.2 Commercial Status. As mentioned above, approximately

77 percent of the Eastern and Midwestern bituminous coal is cleaned to

meet customer specifications for heating value, ash content, and

sulfur content. While most of this coal is cleaned by conventional

cleaning methods, advanced cleaning methods, such as column froth

flotation, are starting to emerge. MicrocelTM is a type of column

froth flotation available through ICF Kaiser Engineers and Control

International. The company is the exclusive licensee for the

technology in the coal fields east of the Mississippi River and has

sold units for commercial operation in Virginia, West Virginia, and

Kentucky, as well as in Australia under sub-license to Bulk Materials

Coal Handling Ltd. Ken-FloteTM is another type of column froth

flotation cell. 

I.2.2 Fuel Switching

Fuel switching refers to switching from one fuel to another

(e.g., high-sulfur coal to low-sulfur coal, or coal to natural gas) to

achieve required emission reductions in a more flexible or cost-
I-6

effective way. For example, coal-fired utilities might switch to

natural gas during the high ozone season in the Northeast, or to

achieve reductions in greenhouse gases. This would also lower their

mercury emissions. In addition, installing pollution control

equipment may not be cost-effective for sources that are not affected

by Title IV regulations, which are generally smaller than affected

utilities. Given the economic benefits of the opt-in program, fuel

switching can be more cost-effective for such smaller sources. 

I.2.3 Flue Gas Treatment for Utility Boilers

This section discusses three types of flue gas treatment which

have been evaluated to some extent for their effectiveness in removing

mercury from utility boiler flue gases. The three technologies are

activated carbon injection, wet flue gas desulfurization (FGD), or wet

scrubbers, and FGD spray dryers. The effectiveness of these

technologies for mercury control vary widely depending on a number of

factors. These factors are described in the sections that follow. 

Current research into the improvement of mercury capture efficiency of

these, and other, approaches is described in section I.2.4 below.

I.2.3.1 Activated Carbon Injection for Utility Boilers. The

effectiveness of activated carbon injection in controlling mercury

emissions from MWCs has been demonstrated (U.S. EPA, 1989a; U.S. EPA,

1989b). The application of activated carbon injection to utility flue

gas, however, cannot be directly scaled from the application at MWCs

due to differences in the amount and composition of flue gas at

utility plants and MWCs. At utility plants, small concentrations of

mercury are contained in a large volume of flue gas, and large amounts

of activated carbon are needed to provide adequate contact between the

carbon particles and mercury. The differences in flue gas

characteristics at MWCs and utility plants must be carefully examined

before considering any technology transfer assumptions.

The level of mercury control achieved in utility flue gas may

depend upon flue gas characteristics such as volume, temperature, fly

ash, and chloride and mercury content. These properties are

distinctly different from those in MWC flue gas.

As shown in Table I-1, typical MWC flue gas is hotter than

utility flue gas after leaving an air preheater. The air preheater

cools the utility flue gas by transferring heat to the incoming

combustion air. Moreover, the mercury concentration of the two gas

streams differs significantly. Mercury concentrations in MWC flue gas

streams may be up to several orders of magnitude greater than those

seen in utility flue gas streams. Likewise, the chloride content of

MWC flue gas may be from 1.4 to 400 times greater than the content

seen in utility flue gas. Finally, with regard to the volume of flue

gas, a utility boiler may have flow rates up to 30 times that of an

MWC.

Because of differences in the amount and composition of flue gas

at utility plants and MWCs, pilot-scale studies of activated carbon
I-7

Table I-1. Comparison of Typical Uncontrolled Flue Gas Parameters at

Utilities and MWCsa,b

Uncontrolled Flue Gas Coal-Fired Utility Oil-Fired

Parameters Boilers Utility Boilers MWC c,d d,e,f g,h

Temperature (C) 121 - 177 121 - 177 177 - 299

Mercury Content (g/dscm) 1 - 25 0.2 - 2 400 - 1,400 i

Chloride Content (g/dscm) 1,000 - 140,000 1,000 - 3,000 200,000 - 400,000

Flow Rate (dscm/min) 11,000 - 4,000,000 10,000 - 2,000,000 80,000 - 200,000 j

Standard conditions are 0C and 1 atmosphere. a

Moisture content in the MWC flue gas was assumed to be 13.2 percent. b

Radian Corporation, 1993a, UNDEERC, 1996, CONSOL INC, 1997. c

Heath, 1994. d

Radian Corporation, 1994. e

Radian Corporation, 1993b. f

Brown and Felsvang. g

Nebel and White, 1991. h

It is not known if oil-fired utility boilers release less mercury overall than coal-fired boilers because the mercury release during oil refining is i

essentially unstudied.

Min = minute j

injection were conducted on utility flue gas where the nominal

concentration of mercury is one part per billion and may have a wide

range of distribution between the different forms of mercury. 

Preliminary results from a limited number of pilot-scale tests on

utility flue gas are summarized in Figure I-1 and presented in greater

detail in section I.2.3.2. These data indicate that the effectiveness

of activated carbon injection varies with several factors. The

mercury removal efficiency for fabric filter and activated carbon

systems ranged from a low of 14 to 47 percent with a median of 29

percent (107-121C, low carbon injection) to a high of 95 to 99 percent

with a median of 98 percent (88-107C, high carbon injection). When

activated carbon injection was used ahead of a spray dryer absorber,

mercury removal efficiency ranged from 50 to 99 percent with a median

of 60 percent when a fabric filter was used for particulate control,

and from 75 to 91 percent with a median of 86 percent when an ESP was

used for particulate control.

Recent results from a few pilot-scale studies under different

flue gas conditions and APCD configurations are also summarized in

this section of the report.

I.2.3.1.1 Utility Flue Gas Factors Affecting Mercury Removal by

Activated Carbon Injection. The level of mercury control achieved in

utility flue gas depends on the temperatures upstream and within the

existing APCDs, residence time (e.g., extent of contact between the

carbon and flue gas mercury) upstream and within the APCDs, volume of

flue gas, flue gas vapor and particulate phase constituents (i.e.,

chlorine as HCl, nitrogen oxides, sulfur oxides, metal oxides on the

I-9

surfaces of particulate matter, fly ash composition, percent carbon in

fly ash, etc.), their interactions with the various types of

carbon(s)/sorbent(s), and the mercury concentration and chemical

species being formed.

Recent studies indicate mercury capture is mass transfer limited

in utility flue gas streams and can be enhanced or suppressed

depending on the temperature, flue gas composition and residence time

within the flue gas. The reasons for this limitation are the low

concentrations of mercury present (one ppb) in the relatively high 

volumes of flue gas (11,000 - 4,000,000 dscm/min). There are higher

concentrations of competing species occupying the active sites of the

carbon. In addition, the flue gas residence time upstream of an ESP

is nominally one second or less with flue gas velocities in the range

of 50 to 60 ft/sec at 149 C (300 F). Compounding the mass transfer o o

limitations is the decrease in the carbon reactivity and capacity at

this nominal, but high temperature. Particle size of the activated

carbon can also impact mercury mass transfer (Vidic et al, 1996;

Flora, et al, 1997; Korpiel et al, 1997; Liu et al, 1997; Rostam-Abadi

et al, 1997; PSCO/ADA et al, 1997; Radian et al, 1997; Carey et al,

June and August, 1997; Waugh et al, August and December, 1997;

PSCO/ADA Technologies, Inc. et al, 1997; and Haythornthwaite et al,

1997). These factors are reviewed below.

Temperature. Mercury is found predominantly in the vapor phase

in utility flue gas (Clarke and Sloss, 1992). If the vapor-phase

mercury were condensed onto PM, the PM could be removed with existing

particulate control devices. Theoretically, cooler temperatures will

increase mercury condensation onto PM (Clarke and Sloss, 1992) and,

subsequently, increase mercury removal with existing PM control

devices.

Earlier data provide some evidence for the temperature dependence

of mercury removal in a pilot-scale FF study. The pilot study

suggests that mercury removal efficiencies apparently increase as the

temperature of the flue gas decreases. Specifically, as the flue gas

temperature decreased from 107 to 99 to 96C (225 to 210 to 205F), the

mercury removal efficiency percentages for a pilot-scale FF

correspondingly increased from 27 to 33 to 51 percent (Chang et al,

1993).

These studies indicate mercury removal efficiencies and the

required amount of activated carbon injection were apparently

temperature dependent within a range of 88 to 121C (190 to 250F) in a

pilot-scale study on the effect of reducing mercury levels in utility

flue gas through activated carbon injection upstream of a FF (Chang et

al, 1993). At the lower temperatures within this range (88 to 96C

[190 to 205F]), mercury concentrations were reduced by 97.7 percent

with an activated carbon injection rate of approximately 155 g

carbon/g of inlet mercury, while at higher temperatures (110 to 121C

[230 to 250F]) mercury concentrations were reduced by only 75 to 87
I-10

percent with an activated carbon injection rate of approximately 3,500

g carbon/g of inlet mercury.

Recent data collected from some coal-fired facilities utilizing

either pilot-scale FFs or ESPs further indicate an apparent

temperature dependence on mercury removal. The FF and ESP pilot-scale

studies indicate an increase of mercury removal with the native fly

ash without carbon injection. Further increases of mercury removal

with carbon injection during lower temperature operation were also

indicated. The studies without carbon injection showed measured

elemental mercury removals across a pilot-scale pulse-jet filter (air

to cloth ratio of 4 ft/min) of 10 and 17 percent at 135 C (275 F) and o o

65 percent at 121 C (250 F); 67 percent at 93 C (200 F), across a pilot- oo oo

scale reverse-gas baghouse of less than 20 percent for an average

temperature of 143 C (289 F), and upstream of a pilot-scale ESP of mean o o

average of 30 percent at 93 - 109 C (200 - 228 F) for the native fly o o

ash (nominal <0.5 percent carbon in ash) from the combustion of a PRB

Belle Ayre coal (PSCO/ADA Technologies, Inc.,et al, 1997; Sjostrum et

al, 1997; and Haythornthwaite 1997).

In contrast to the higher mercury removals at lower temperatures

are data collected from a full-scale utility boiler without carbon

injection. The testing was conducted on a 70 MWe unit firing a Powder

River Basin coal from the Montana area in a Riley Stoker front-fired

boiler. The only APCD is a reverse-gas baghouse for particulate

control. Mercury measurements were taken at the inlet and outlet of

the baghouse with triplicate samples being collected and analyzed for

total mercury, including speciation. Draft U.S. EPA Method 29 and the

Bloom or MESA method were utilized simultaneously at each location. 

Both methods measured total inlet mercury concentrations (three data

points for each method) at the 6.4 and 6.5 g/m levels, respectively, 3

with approximately 60 percent of the total being measured as elemental

mercury for each method. The elemental mercury was essentially

removed across the baghouse due to the native fly ash (during the

three test periods the percent carbon was 3.5, 2.9, and 2.9 with an

average of 3.1 percent) with the outlet concentrations being 2.6 and

3.1 g/m of the ionic form as measured by the respective methods. The 3

removals indicated by the two methods were 60 and 52 percent of the

total, respectively, at average temperatures (three data points each)

at the air heater outlet of 189 C (372 F), baghouse inlet of 174 C oo o

(346 F), and baghouse outlet of 166 C (330 F). Approximately 40 percent o oo

of the total mercury was indicated on the filter catch of the Method

29 train (filter at 121 C + 8 C [250 F+ 15 F] which could capture the oo oo

mercury as it comes in contact with the filtered fly ash) and the

hopper ash samples indicated a high level of mercury comparable to the

removals. The mercury capture during this testing was indicative of

removals across the baghouse and not in-flight capture upstream of the

baghouse (Jackson et al, 1994).

 As indicated, the mercury removals of the native fly ash at

these conditions are not typical of the past and more recent field

characterizations and pilot-scale mercury technology investigations. 
I-11

This utility site is proposed to be further characterized in mid-1998

with the more precise Ontario Hydro mercury speciation method. In

addition, in-flight capture of mercury will be investigated upstream

of the baghouse along with the baghouse removals (DOE/FETC et al,

Phase II, 1997). Currently, laboratory tests are being conducted on

the fly ash under simulated flue gas conditions to provide some

insight into the factors influencing high elemental mercury capture at

nominal flue gas temperatures of 149 C (300 F) (U.S. DOE/FETC R&D, o o

1997).

Typical removals of mercury by the fly ash for low-sulfur and

medium- to high-sulfur bituminous coals under the above conditions are

approximately 10 percent or less and can be influenced by the sampling

method. The fly ash is captured on a filter of the sampling train at

121 C ([250 F] which is lower than the flue gas) before the chilled o o

impinger based solutions being utilized for the collection of the

vapor phase mercury. The passing of the flue gas through the captured

fly ash on the filter can provide false indications of in-flight

capture of mercury. As indicated, the removals of mercury assumed

from the fly ash in-flight can be inflated due to the sampling method,

but still in most cases are below 10 percent (Miller 1994 and 1995;

EPRI, 1994; U.S. DOE Report, 1996; Laudal et al, 1996 and 1997; Hargis

et al, 1996; Redinger et al, 1997; Holmes et al, 1997; Waugh et al,

1997; and DeVito et al, 1997).

The pilot-scale activated carbon injection studies indicated that

more mercury was removed and less carbon was needed at lower flue gas

temperatures; in other words, the ideal use of activated carbon may be

at lower flue gas temperatures. It may not be possible, however, to

lower the flue gas temperature at a given utility plant because

utility plants typically operate with a stack gas temperature between

121 and 177C (250 to 351F) upstream of any particulate control device

to avoid acid condensation and, consequently, equipment corrosion. 

The stack gas temperature may be lowered below 96C (205F) and acid

condensation may be avoided provided low-sulfur coals (less than about

1 weight percent sulfur) are burned, but it may depend on whether the

coal is a subbituminous or a bituminous coal (McKenna and Turner,

1989; ABB et al, 1996 and 1997; PSCO/ADA Technologies, Inc. et al,

1996 and 1997; Sjostrum et al, 1997; Haythornthwaite 1997; Radian et

al, 1997; Carey et al, 1996 and 1997; Hargrove et al, 1997; Waugh et

al, 1997). If a utility burns low-sulfur coal and uses an ESP for

particulate control, however, the flue gas will probably require

conditioning to reduce the high resistivity of the fly ash because

high resistivity makes the fly ash difficult to collect with an ESP,

but again, it is dependent on coal type.

Further research is needed to evaluate humidification in flue gas

ducts while firing other low-sulfur coals and most importantly mediumto high-sulfur coals in the furnace. This is extremely important for

the approximately 65 percent of the utility industry utilizing an ESP

as the only APCD. Subsequent sulfuric acid mist formed from the

condensation of sulfur trioxide below the acid dew point(s) can be
I-12

extremely detrimental to ESP- and FF-equipped utilities, duct work,

all downstream equipment, compliance for opacity, and plume effects

(i.e., visibility - blue plume). In addition, it is desirable for

utilities to minimize the amount of sulfuric acid being emitted as

these emissions must be reported annually to the Toxics Release

Inventory.

In some cases, lower temperatures do have an influence on the

amount of mercury removed by certain native fly ashes alone and in

combination with activated carbon, but this not typical of the utility

population (e.g., majority of low- and medium-sulfur bituminous

coals). The factors or mechanisms influencing the ability of the

small percentage of coals and subsequent fly ash to adsorb mercury

and/or convert mercury from one form to another in-flight and across

fabric filters need to be further investigated in order to effectively

capture the different forms of mercury. These mechanisms can be

associated with the type of activated carbon, the fly ash components,

the vapor phase chemical species of the flue gas, and all the possible

interactions, along with the control device being augmented to remove

mercury. These factors are not fully understood at this time, but

many research organizations are performing fundamental and applied

research studies to investigate and subsequently understand them. 

Based upon the preliminary pilot-scale studies conducted at

temperatures below 121C (250F), the least efficient and most costly

use of carbon injection for mercury control is at higher temperatures

with greater injection rates.

Volume. At utility plants, mercury control techniques must

adequately treat the entire volume of gas in order to remove

relatively small concentrations of mercury (0.2 to 21 μg/dscm, at

7 percent O ). High mass carbon-to-mercury ratios will be required due 2

to a nominal one ppb of mercury being in different forms and being in

the high flue gas volumes with competing vapor phase compounds at many

orders of magnitude higher. Currently, mercury mass transfer

limitations are encountered regardless of the type of coal, operating

conditions, and APCD.

Mercury Speciation and Type of Activated Carbon. With a few

exceptions, the total mercury concentration in coal is relatively

constant across the United States (20 ppb to 120 ppb). However, when

the different coals are fired in a combustor there are substantial

variations in the concentrations of elemental versus ionic mercury. 

The percentage of Hg is from near zero percent to >70 percent. The 0

speciation then is very dependent on coal type. The chemical species

of mercury formed during the combustion process and post-combustion

conditions vary significantly from one plant to another. While

combustion conditions vary, the subsequent fly ash, carbon in the ash,

and vapor phase constituents may play a major role in the percentage

of the chemical species of mercury formed. Understanding the rate

controlling mechanisms (i.e. transport, equilibrium, and kinetics)

will aid in predicting the species formed and eventually will aid in
I-13

optimizing existing APCDs for mercury removal. Kinetics may play more

of a role on the form of mercury than anticipated. Depending on the

type of coal utilized, effective removal may be dependent on the

species of mercury present in the flue gas (Senior et al, June and

November, 1997; PSI et al, 1997). For example, the ionic mercury form

(i.e., Hg ) is water soluble and is less volatile than elemental ++

mercury (i.e., Hg ). Thus, reducing the temperature of the flue gas 0

and wet scrubbing of the flue gas may result in increased ionic

mercury removal.

 In the early 1990s EPRI and DOE initiated very extensive electric

utility air toxics characterization programs. As part of these

programs, measurement of speciated mercury emissions from each plant

was attempted. Because there was no validated mercury speciation

sampling method, U.S. EPA Method 29 and the Bloom or Brooks Rand

(referred to as the MESA) methods were used. The results from these

characterizations strongly suggested that U.S. EPA Method 29 does not

properly speciate mercury under certain conditions. In addition,

there were questions as to the ability of the MESA method to speciate

mercury in flue gas from coal combustion. Results from the MESA

sampling method and unique analytical technique(s) are summarized in

Table I-2 for coal- and oil-fired utility flue gas (Bloom et al,

1993).

As shown in Table I-2, the distribution of ionic mercury, most

likely HgCl in coal-fired utility flue gas, ranged from 12 to 99 2

percent of the total mercury content and averaged 79 percent; the

distribution of elemental mercury in coal-fired utility flue gas

ranged from 0.8 to 87.5 percent of the total mercury content and

averaged 21 percent. Analysis of two samples of flue gas taken from

oil-fired boilers, however, suggests that mercury in oil-fired boiler

flue gas is predominantly in the elemental form (see Table I-2). The

variability in the speciation of vapor-phase mercury in coal-fired

flue gas may explain the variation in mercury removal that is seen

with existing control devices (DeVito et al, 1993).

Since that time a substantial amount of work has been done to

develop sampling and analytical methods for determining mercury

speciation in flue gas from fossil fuel combustion. In 1994 EPRI and

DOE contracted with the University of North Dakota Energy &

Environmental Research Center (UNDEERC) to complete a series of benchand pilot-scale evaluations on mercury speciation measurement methods.

Concurrently, work was also being conducted by CONSOL, Inc., Radian

International, Advanced Technology Systems, and Babcock & Wilcox at

the bench- and pilot-scales, along with full-scale coal-fired power

plant studies and characterizations.

In the pilot-scale work conducted at EPRI’s ECTC by Radian

International and the pilot-scale work conducted by the UNDEERC for

both EPRI and DOE, it was proven that U.S. EPA Method 29 does not

properly speciate mercury under certain conditions (Hargrove et al,

1995; Laudal et al, 1996; Stouffer et al, 1996; Khosah et al, 1996; 
I-14

Table I-2. Mercury Speciation in 33 Samples of Coal-Fired and 3 Samples of Utility Flue Gas

μg/Nm Percent 3a

Totalb

(μg/Nm ) Ionic Other Hg Species Elemental Total Ionic Elemental 3 

c d

COAL-FIRED BOILERS

5.17 0.29 1.6 7.06 77.34 22.66 4.24 0.41 0.59 5.24 88.74 11.26 2.96 0.41 0.31 3.68 91.57 8.42 2.84 0.56 0.15 3.55 95.77 4.23 8.7 --- 0.29 8.99 96.77 3.23 8.6 --- 0.2 8.8 97.73 2.27 

10.8 --- 0.22 11 98.00 2.00 

7.86 --- 0.06 7.92 99.24 0.76 7.7 --- 0.07 7.77 99.1 0.90 9.4 --- 0.1 9.5 98.95 1.05 7.7 --- 0.08 7.7 98.97 1.03 6.03 --- 1.94 7.97 75.66 24.34 6.46 --- 1.29 7.75 83.35 16.65 3.03 --- 1.52 4.55 66.59 33.41 5.36 --- 1.54 6.9 77.68 22.32 5.98 --- 2.54 8.52 70.19 29.81 3.5 --- 1.67 5.17 67.7 32.30 3.1 --- 1.72 4.82 64.32 35.68 
Table I-2. (continued)

μg/Nm Percent 3a

Totalb

(μg/Nm ) Ionic Other Hg Species Elemental Total Ionic Elemental 3 

c d

I-15

2.99 --- 1.51 4.5 66.44 33.56 

2.83 --- 0.88 3.71 76.28 23.72 3.33 --- 1.26 4.59 72.55 27.45 3.83 --- 1.52 5.35 71.59 28.41 2.2 --- 1.57 3.77 58.36 41.64 

--- --- --- 12.5 87.5 

7.37 0.0021 1.85 9.22 79.92 20.07 5.01h --- 1.26 6.28 79.90 20.10 5.72 1.39 <0.1 7.10 98.61 1.39 4.49 0.40 0.89 5.78 84.60 15.40 2.93 0.07 0.16 3.15 94.94 5.06 2.79 0.32 1.02 4.13 75.30 24.70 5.71 1.37 2.99 10.22 70.31 29.69 7.01 1.02 0.11 8.13 98.65 1.35 4.11 0.62 2.32 7.05 67.09 32.91 1.14 0.30 3.12 4.56 31.58 68.42 

COAL-FIRED BOILER DATA RANGES

1.1 - 10.8 0.002 - 1.4 0.06 - 3.1 3.1 - 11 12 - 99 1.8 - 88 
Table I-2. (continued)

μg/Nm Percent 3a

Totalb

(μg/Nm ) Ionic Other Hg Species Elemental Total Ionic Elemental 3 

c d

I-16

OIL-FIRED BOILERS

0.03 0.01 0.15 0.19 21.05 78.95 

<0.02 0.01 <0.02 --- --- --- 

0.02 0.01 0.09 0.12 25.00 75.00

OIL-FIRED BOILER DATA RANGES

0.02 - 0.03 --- 0.02 - 0.15 0.12 - 0.19 23 77 

μg/Nm3 = micrograms per normal cubic meter (at 0 C). a oValue is a combination of ionic and other Hg species. bFor other Hg species, the references list this value as an organic mercury species. According to Chang (1994), however, the organic mercury species values are invalid 

c

species concentrations to give a total ionic mercury value.The total ionic percentage represents the ionic and other Hg species concentrations. dChang et al, 1993. Flue gas sampled at FF inlet. eNoblett et al, 1993. Flue gas sampled at wet scrubber inlet. fFelsvang et al, 1993. Flue gas sampled at inlet to pilot-scale spray dryer absorption (SDA)/ESP system. gFelsvang et al, 1993. Flue gas sampled at inlet to full- and pilot-scale SDA/FF system. hDismukes et al, 1993. iBloom et al, 1993. j
I-17

and Laudal et al, December, 1997). Further studies at UNDEERC

indicated this finding is related to high SO concentrations with the 2

method overestimating the ionic mercury up to 50 percent. Therefore,

tests were conducted to evaluate a number of alternative sampling

methods. Mercury speciation sampling methods that have been tested

include the following:

 U.S. EPA Method 29

 Mercury Speciation Adsorption Method (Frontier Geosciences

and Brooks Rand - the Bloom method)

 Ontario Hydro method (Ontario Hydro)

 Tris-buffer method (Radian International)

 U.S. EPA Draft Method 101B (Research Triangle Institute)

Bench- and pilot-scale studies also showed that the MESA method

did not speciate mercury correctly when tested with coal-fired flue

gas. The method is greatly affected by an interaction between SO and 2

NO in the flue gas. When SO is present in concentrations >500 ppm x 2

and NO is present at >250 ppm, the MESA method can overestimate the x

ionic mercury fraction up to 75 percent (Laudal et al, 1996). Based

on the exploratory pilot-scale tests, the Ontario Hydro method and

U.S. EPA Draft Method 101B were selected to be more formally evaluated

using the protocol established in U.S. EPA Method 301. However,

because there is no reference method to compare to U.S. EPA Method

301, the method only provides the precision and bias associated with

the sampling procedures. To obtain the accuracy of the speciated

mercury measurement methods, it was necessary to do dynamic spiking of

the flue gas stream. Spiking was done first with elemental mercury,

then with HgCl . Results showed that both the Ontario Hydro and U.S. 2

EPA Draft Method 101B passed the U.S. EPA Method 301 criteria;

however, the Ontario Hydro method showed much less variability than

Method 101B. Therefore, the Ontario Hydro method is being recommended

by DOE as the best method to speciate mercury in coal-fired systems. 

The method is being submitted to the American Society for Testing and

Materials and U.S. EPA for approval.

Field tests comparing U.S. EPA Method 29 and/or the MESA method,

with either or both the Ontario Hydro method and the tris-buffer

method have been completed during 1995 through 1997. Results showed

that U.S. EPA Method 29 and the MESA method gave a high bias for the

ionic form of mercury compared to the Ontario Hydro and tris-buffer

methods, which is in agreement with the Radian International and

UNDEERC pilot-scale studies. DOE and EPRI are planning field studies

and characterizations on mercury speciation with the Ontario Hydro

method.

 

The variability in the distribution of vapor-phase mercury

species in coal-fired flue gas may depend upon the chloride

concentration in coal. Using the analytical techniques developed by

Bloom et al, (1993), it has been observed that higher concentrations

of ionic mercury are obtained in utility flue gas when the combusted

coal has a high chloride concentration (0.1 to 0.3 weight percent)
I-18

(Felsvang et al, 1993; Noblett et al, 1993), but more data are needed

to verify this association. The distribution of mercury species in

coal-fired flue gas also appears to vary with the type of coal (e.g.,

bituminous, subbituminous, or lignite) (Chang, 1994; Boyce, 1994;

Laudal et al, 1996 and 1997; Redinger et al, 1996 and 1997; and 

DeVito et al, 1997).

 Low-sulfur bituminous coals and other subbituminous coals with

low-sulfur content are very different regarding the mercury

distribution between the elemental and oxidized forms in the flue gas.

(Bloom et al, 1993; DeVito et al, 1993; EPRI, 1994; Prestbo et al,

1995; U.S. DOE Report, 1996; Laudal et al, 1996 and 1997; Pavlish et

al, 1997; Hargrove et al, 1997; Senior et al, June and November 1997;

PSI et al, 1997; and Devito et al, 1997). The fly ash characteristics

are extremely different and some of the subbituminous coals produce

fly ash that is more reactive and adsorbs mercury at higher rates than

fly ash from the bituminous coals. In addition, the bituminous coals

convert the elemental mercury at higher rates and levels as compared

to the fly ash from subbituminous coals. The adsorption and/or

conversion is impacted by temperature, but the composition of the fly

ash and vapor phase compounds also play a major role in these effects 

(Miller, 1994 and 1995; Laudal et al, 1996 and 1997; Carey et al, 1996

and 1997; Radian International et al, 1997; Senior et al, June and

November 1997; and DeVito et al, 1997). 

Radian International conducted both laboratory and field studies

to investigate catalytic oxidation of vapor-phase elemental mercury in

coal-fired utility flue gas streams. Catalytic oxidation of vaporphase elemental mercury can potentially increase the total mercury

removal in the two technologies with the most potential for removing

mercury from flue gas: wet scrubbing and sorbent injection. To

investigate this process, potential catalyst materials were tested

using three different test configurations. These configurations

included laboratory fixed beds tests, pilot-scale fabric filter tests,

and sample filter tests using flue gas from a full-scale utility.

Oxidation of elemental mercury using catalyst materials was

successfully demonstrated using each of the test configurations

mentioned above. In the laboratory fixed bed tests, the effects of

temperature and flue gas composition were investigated. In general,

oxidation of elemental mercury decreased as the temperature increased. 

Flue gas composition also appears to be important to oxidation, with

HCl and possibly NO affecting oxidation. x

Based on the laboratory and pilot-scale tests, the most

successful catalyst was a carbon-based material. After injecting

about 20 pounds of this material into a pilot-scale fabric filter,

greater than 75 percent of the inlet vapor-phase elemental mercury was

oxidized across the fabric filter for 10 consecutive days. Similar

results were obtained at a full-scale facility by measuring oxidation

across a sample filter. These results confirmed the ability of the

carbon-based material to oxidize elemental mercury under different
I-19

flue gas conditions (with and without HCl and various levels of NO ). x 

Other catalyst materials that were identified and warrant further

investigation included several iron-based materials, a conventional

SCR catalyst, and some fly ash samples (Carey et al, 1996 and 1997;

Radian International et al, 1997).

The speciation of mercury is extremely important in planning

control strategies, but it is still in the early stages of

investigation. Preliminary laboratory- and field pilot-scale studies

indicate the form of mercury being removed is impacted by the type of

carbon being injected. Both physical and chemical adsorption of the

mercury can be achieved, but is dependent on the concentration and

most importantly the form of mercury (elemental or ionic/oxidized). 

Limited studies have indicated simultaneous removal of both forms of

mercury with one activated carbon, but at very low levels. A further

complication is that some activated and chemically impregnated

activated carbon can, under certain conditions, convert the elemental

mercury to an ionic form with either a net increase or decrease in

mercury capture (Miller, 1994 and 1995; PSCO/ADA Technologies, Inc.,

1997; and Radian et al, 1997).

Earlier studies with activated and chemically impregnated

activated carbon utilized either U.S. EPA Method 101A (only total

mercury) and either U.S. EPA Method 29 or the MESA method (both for

speciated mercury as well as total) for the mercury measurements. As

indicated from the studies conducted at the UNDEERC, these two

speciated methods have overestimated the ionic form of mercury up to

50 percent and 75 percent, respectively. The interactions of these

carbons with the fly ash and vapor phase species in the flue gas can

dramatically increase or decrease mercury capture of the carbon, and

measuring the impacts are difficult and sometimes impossible to do. 

In addition, controlled laboratory studies were conducted with the

injection of activated carbon(s) and elemental mercury or HgCl in 2

either nitrogen or simulated flue gas streams. The results indicated

different and varying levels of mercury capture between the nitrogen

and simulated flue gas streams. Promising results from these tests,

in most cases, have not been repeated on actual flue gas streams of

the pilot-scale and slipstream studies at the various coal-fired

facilities.

More recent tests have been conducted on flue gas streams

containing primarily elemental mercury that was often supplemented 

with additional elemental mercury during testing. The tests were

designed to investigate elemental mercury capture with commercially

available activated carbons. Limited studies have been conducted on

chemically impregnated carbons, but they are being considered for

future testing on both simulated and actual flue gas.

Several types of novel activated carbons for gas phase elemental

mercury removal that have orders of magnitude higher saturation

capacities when compared to virgin activated carbons are also

available. These activated carbons are typically impregnated with
I-20

sulfur or iodine lending to the enhanced capacity for mercury uptake

due to the chemical reaction between the impregnated material and

elemental mercury. However, many of the sorbents exhibited

deteriorated performance at temperatures typical of coal-fired power

plant operations. 

Recently, researchers at the University of Pittsburgh developed a

series of sulfur-impregnated carbons that exhibited high elemental

mercury uptake efficiency at 140 C (284 F) when compared to o o

commercially available activated carbons. Dynamic adsorption capacity

of these carbons as high as 4,000 μg Hg/g was measured using a fixedbed absorber with an empty bed contact time of 0.011 second and

influent mercury concentration of 55 μg/m . This capacity is almost 3

three orders of magnitude greater than the capacity of virgin

activated carbon and an order of magnitude greater than the capacity

of commercially available impregnated activated carbon. The

comparisons were conducted at identical operating conditions using

nitrogen as a carrier gas.

The increased performance is attributed to the impregnation of

the carbon(s) with sulfur at elevated temperatures of 400 - 600 C (752 o

- 1112 F). This promoted a more uniform distribution of short linear o

chains of sulfur allotropes (S and S ) on the carbon surface as 2 6

opposed to having predominately S rings condensed in the macropore 8

region of commercially available sulfur impregnated carbons. In

addition, the sulfur impregnated carbons prepared at elevated

temperatures exhibited significantly better thermal stability since no

sulfur loss was observed even after exposure at 400 C (752 F) (Vidic et o o

al, 1996; Korpiel et al, 1997; Flora et al, 1997; and Liu et al,

1997). 

These impregnated activated carbons exhibited orders of magnitude

higher dynamic capacity as compared to virgin activated carbons. 

However, the key question remains as to whether this capacity can be

utilized in a flue gas stream where residence times of one second or

less are available for injection upstream of the ESP- equipped

facility. These high capacity carbons may be limited to use on FFequipped facilities or control strategies employing devices for higher

flue gas and carbon contact or residence times. The costs associated

with impregnated activated carbons may also limit their use to FFequipped facilities. 

Further investigation, development, and enhancement of activated

carbons and chemically impregnated carbons for mercury capture in flue

gas from coal-fired facilities is needed. The conditions of the

chemical impregnation may be critical and commercially available

impregnated activated carbons may not be highly effective in all the

various flue gas produced from the combustion of coal. New virgin and

chemically impregnated activated carbons may need to be developed for

the highly variable and complex flue gas streams encountered in the

utility industry and the extreme mercury mass transfer limitation(s).
I-21

The association between chloride content of the fuel and the

concentration of ionic mercury in the flue gas also may apply to fuel

oil. This association, however, has not been examined.

Studies of a pilot-scale wet FGD system treating coal-fired flue

gas indicate that more than 90 percent of the ionic mercury was

removed while hardly any of the elemental mercury was removed (Noblett

et al, 1993; Redinger et al, 1996 and 1997; Carey et al, June and July

1996; Evans et al, 1996; and Hargrove et al, 1995 and 1997). 

Similarly, studies at a pilot-scale SDA/ESP system treating coal-fired

flue gas suggest that 95 percent of the ionic mercury and essentially

none of the elemental mercury were removed (Felsvang et al, 1993). 

The effectiveness of activated carbon injection in recovering

different forms of mercury is still being studied. Preliminary

results are available from the studies described in Section 2.3.1.2,

Current Research on Activated Carbon Injection for Utilities. 

Flue Gas Composition. The temperature, volume of the flue gas,

and type of activated carbon can have an impact on the form and

subsequent capture of mercury in coal-fired produced flue gas streams. 

These factors are not independent of one another, but are synergistic

with one another and are very dependent on the composition of flue

gas. This includes both the vapor and particulate phases of the flue

gas. As previously indicated, hydrogen chloride, sulfur and nitrogen

oxides, oxygen, water, fly ash and its composition, and even carbon

monoxide in the flue gas can either impede or enhance the form and

subsequent capture of the mercury with fly ash and injected carbon. 

There are other flue gas constituents that could also impact mercury

collection, but research is needed to determine what other

constituents do and why. 

A recent bench-scale study investigated the effects of SO and HCl 2

on the adsorption of elemental mercury and mercuric chloride (HgCl ) by a 2

lignite-based activated carbon (Carey et al, 1997). Equilibrium

adsorption capacities were determined for fixed beds of the carbon at

275F and three flue gas compositions: one containing 1,600 ppm SO and 2

50 ppm HCl (the baseline composition); a second containing no SO and 50 2

ppm HCl; and a third containing 1,600 ppm SO and no HCl. (All three 2

compositions of flue gas had the same concentration of elemental mercury,

mercuric chloride, CO , water, and O ). 2 2

Figure I-2 illustrates the effect of SO and HCl on the 2

equilibrium adsorption capacity of the lignite-based activated carbon

for elemental mercury and mercuric chloride. Removing SO from the 2

flue gas increased the equilibrium adsorption capacities for both

kinds of mercury (compared to the baseline capacities). The increase

was particularly notable for the adsorption of elemental mercury. For

example, after removing SO from the flue gas, the equilibrium 2

adsorption capacity for elemental mercury increased by a factor of

about 5.5 compared to 3.5 for mercuric chloride.
1200

4300

1200

2800

15400

0

2000

4000

6000

8000

10000

12000

14000

16000

HgCl2

Baseline

HgCl2,

No SO2

HgCl2

No HCl

Hg(0)

Baseline

Hg(0)

No SO2

Hg(0)

No HCl

I-22

Equilibrium Adsorption Capacity,

Figure I-2

Equilibrium Adsorption Capacity of Elemental Mercury

(Hg(0)) and Mercuric Chloride (HgC12) by

a Lignite-Based Activated Carbon
I-23

Removing HCl from the flue gas did not affect the equilibrium

adsorption capacity of the carbon for mercuric chloride; however, it

did prevent the carbon from adsorbing elemental mercury. The latter

result suggests that HCl participates in the adsorption mechanism of

elemental mercury when using a lignite-based activated carbon and that

the adsorption mechanism is not purely physical, i.e., interactions

between elemental mercury and HCl on the carbon surface may be

important.

The results from Figure I-2 indicate that flue gas composition

affects carbon performance. With no HCl in the gas, the carbon

adsorption capacity for mercuric chloride was larger than that for

elemental mercury. This result is opposite to that observed at

baseline conditions where the carbon adsorption capacity for elemental

mercury was larger than that for mercuric chloride. The results from

Figure I-2 also indicate that performing carbon adsorption tests under

realistic operating conditions is important. Many bench-scale carbon

tests in the past have been conducted using nitrogen as the carrier

gas. Tests conducted in nitrogen could produce different results than

tests conducted in simulated flue gas; however, the effect of SO and 2

HCl on adsorption capacity could also be sorbent dependent. Other

carbons may not be affected by the presence of HCl and SO if the 2

mercury adsorption mechanism is different.

Further details on the effects of flue gas components, including

the interactions with fly ash, can be obtained from two reports by

Laudal et al (November, 1996 and December, 1997). The flue gas and

mercury chemistries and their subsequent interactions need to be fully

understood at the various flue gas conditions encountered across the

utility industry for effective low cost mercury strategies to be

universally realized.

I.2.3.1.2 Current Research on Activated Carbon Injection for

Utilities. Previously, research was conducted on activated carbon

injection at a facility with a pilot-scale SDA/ESP system in Denmark

(Felsvang et al, 1993); at a facility with both a pilot- and fullscale SDA/FF system by Joy/Niro and Northern States Power (Felsvang et

al, 1993); at a pilot-scale coal combustor and FF by Miller et al

(1994 and 1995); and at a pilot-scale pulse-jet FF system at a utility

power plant by EPRI (Chang et al, 1993). These results are presented

in detail in section I.2.3.2. Preliminary results are available from

the first three studies as described below.

In testing at the first facility, a pilot-scale SDA/ESP system in

Denmark (Felsvang et al, 1993), the flue gas contained from 66.6 to

83.4 percent ionic mercury, with an average of 75.2 percent ionic

mercury, and elemental mercury comprised the remainder of the total

mercury concentration in the flue gas. Without activated carbon

injection, the pilot-scale SDA/ESP system removed 96.8 percent of the

ionic mercury and essentially none of the elemental mercury from coalfired flue gas or, in other words, the system removed 72.5 percent of

the total mercury. During testing with activated carbon injection,
I-24

the flue gas contained from 58.4 to 77.7 percent ionic mercury, with

an average of 69.5 percent ionic mercury, and elemental mercury

comprised the remainder of the total mercury concentration in the flue

gas. Activated carbon injection ahead of the SDA/ESP system removed

46.4 percent of the elemental mercury and 84.3 percent of the total

mercury (Felsvang et al, 1993).

In testing by Joy/Niro and Northern States Power at the second

facility that had a full- and pilot-scale SDA/FF system, the flue gas

contained 85 to 90 percent elemental mercury. Without activated

carbon injection, the full- and pilot-scale SDA/FF systems removed 10

to 20 percent of the total mercury from the coal-fired flue gas

(Felsvang et al, 1993), and the low removal of total mercury may be

attributed to essentially complete removal of the ionic mercury and

poor removal of the elemental mercury. Activated carbon injection

ahead of the pilot-scale SDA/FF system increased the removal of total

mercury to approximately 55 percent, and injection of iodide- and

sulfur-impregnated activated carbon increased the removal of total

mercury to approximately 90 percent (Felsvang et al, 1993). Thus, the

studies at this SDA/FF system suggest that sulfur- and iodideimpregnated carbons are needed for total mercury removals of 90

percent, when elemental mercury is the predominant mercury species. 

Furthermore, the studies suggest that total mercury removal

efficiencies are dependent upon mercury speciation.

Finally, laboratory-scale tests at the UNDEERC found that for

some conditions iodine-impregnated carbon is much more effective than

lignite-based activated carbon in removing elemental mercury (Miller

et al, 1994). Sorbent injection tests were conducted at flue gas

temperatures ranging from 125 to 200C (257 to 392F). Iodineimpregnated carbon had a high removal efficiency of elemental mercury

(greater than 95 percent removal) across the entire range of

temperatures for one subbituminous coal. However, for a second

subbituminous coal the iodine-impregnated carbon appeared to convert

the elemental mercury to ionic mercury with little net total mercury

removal. A reason for the difference is not obvious, but may be the

result of differing concentrations of SO , HCl, NO , HF, and possibly 2 x

CO. Lignite-based activated carbon removed approximately 50 percent

of elemental mercury at 130C; however, it’s removal efficiency for

elemental mercury dropped dramatically as temperature increased. For

both carbons, the removal efficiency of oxidized mercury was highly

temperature dependent. At 125C, the iodine-impregnated carbon was

somewhat effective at removing oxidized mercury, while it removed no

oxidized mercury at 175C. The lignite-activated carbon showed a

similar trend (Miller et al, 1994 and 1995).

The most recent studies have utilized American Norit Companies’

commercially available Darco FGD activated carbon developed from a

lignite coal. This carbon has been extensively utilized more than any

other commercial activated carbon for the DOE and EPRI-funded mercury

control studies investigating sorbent injection (Miller et al, 1994

and 1995; Chen et al, 1996; Hunt, 1996; ABB et al, 1997; Carey et al,
I-25

July, 1996 and June, 1997; Radian International et al, 1997; Sjostrum 

et al, 1997; Haythornthwaite et al, 1997; PSCO/ADA, et al, 1997;

Rostam-Abadi et al, 1997; Waugh et al, August and December, 1997; and

Brown, 1997.) The activated carbon typically has a mass mean diameter

of 15 microns, a BET surface area of 600 m /g and a nominal 2

equilibrium adsorption capacity of 500 μg Hg/g C. These parameters

have been repeated by many research institutions and are in agreement

with Norits’ specifications (Carey et al, 1997; Radian International

et al, 1997; Haythornthwaite et al, 1997; Waugh et al, 1997; and

Rostam-Abadi et al, 1997).

The equilibrium adsorption capacity of the activated carbon is

important for fabric filter systems. For flue gas residence times of

less than one second, typical upstream conditions prior to the inlet

of an ESP, the equilibrium adsorption capacity of 500 μg Hg/g C may

not be the most critical parameter. Reactivity may need to dominate,

but can be suppressed at the nominal temperature of 149 C (300 F) of o o

the flue gas upstream of utility ESPs. Chemically impregnated carbons

may increase the reactivity and subsequent capture of mercury, but

very few studies have indicated the effectiveness of chemically

impregnated carbons for in-flight capture of mercury (especially at

one second or less residence time) (Vidic et al, 1996; Korpiel et al,

1997; and Liu et al, 1997). 

The chemically impregnated carbons may be cost prohibited and may

be better suited for high mercury adsorption capacities corresponding

to longer contact times (carbon and novel fluid beds or fabric filters

- reverse-gas and pulse-jet with the pulse-jet also being downstream

of an existing ESP). Examples of this technology are EPRI’s COmpact

Hybrid Particulate Collector (COHPAC) or TOXICON (a pulse-jet baghouse

operating at a high air-to-cloth ratio downstream of the primary

particulate control device with sorbent injection upstream of the

baghouse for air toxics or in these cases mercury).

Recent studies further support the mercury mass transfer

limitations since the removal of mercury above 50 percent to the 90

percent level for in-flight capture and above 75 percent to 90 percent

for extended contact times (>one half hour across a fabric filter) is

dependent on near exponential increases in the carbon injection or

carbon to mercury ratios (Vidic et al, 1996; Flora et al, 1997;

PSCO/ADA et al, 1997; Carey et al, June and August, 1997; Korpiel et

al, 1997; Liu et al, 1997; Rostam-Abadi et al, 1997; and Waugh et al,

August and December, 1997). The PSCO/ADA studies indicate a nominal

5000:1 carbon-(Norit or Darco FGD)to-mercury mass ratio at 106 Co

(222 F) upstream of an pilot-scale ESP with a residence time ranging o

between 0.75 and 1.5 seconds to remove the mercury at a level of 48

percent. This 48 percent includes 30 percent of the mercury being

removed by the native fly ash. Studies have indicated the fly ash

from this PRB coal (Comanche or Belle Arye coal from Wyoming) has a

high equilibrium adsorption capacity for mercury even at <0.5 percent

carbon levels in the fly ash (Miller et al, 1994 and 1995; Laudal et

al, 1996 and 1997; Haythornthwaite et al, 1997; and PSCO/ADA et al,
I-26

1997). This mercury removal in-flight is high compared to other PRB

and subbituminous coals. The overall mercury adsorption can be higher

than bituminous coals for the same amount of carbon in the fly ash. 

The adsorption capacity or reactivity for both ranks of coal does

increase with a decrease in temperature, but not at the same rate or

level. In addition, tests were conducted with the re-injection of the

Commanche fly ash upstream of the ESP configuration and indicated on

average less than 10 percent mercury capture.

The pulse-jet pilot-scale FF tests at the PSCO facility also

indicated a substantial increase in carbon injection or mass carbonto-mercury ratio from 76 percent mercury removal at a ratio of

>20,300:1 (C/Hg) to >90 percent mercury removal at a ratio of

>36,600:1. Mercury concentrations were not constant at these ratios

with nearly 18 percent mercury reductions being attributed to residual

fly ash on the bags. These tests were conducted as “clean” tests,

that is, no fly ash was in the flue gas stream (the flue gas was drawn

downstream of the facility’s existing fabric filter). During the

testing with fly ash present, different results were indicated. The

mercury removal “by the fly ash” was dramatically impacted by

temperature. At temperatures between 93 C (200 F) and 121 C (250 F) oo oo

mercury removals due to the fly ash were at 66 percent while an

increase to 135 C (275 F) indicated removals in the range of only 10 o o

percent to 17 percent. In addition to the fly ash removals, the

amount of carbon needed at even small increases in temperature was

noticeable. Carbon to mercury ratios of 3400:1 were needed for

mercury removals of 74 percent at only 109 C (228 F) while ratios of o o

>8700:1 were needed to remove mercury at 87 percent for a temperature

of 113 C (236 ). The mercury concentrations were steady during these o o

tests. 

These data were collected at the same contact times (carbon

exposed to flue gas across the fabric filter) and the QA/QC on the

mercury sampling methods were indicative of the close mercury

concentrations for all the tests at the close, but different

temperatures. The adsorption of the mercury appears to be mass

transfer limited even at high residence or contact times. In

addition, the high mercury removals include the 66 percent mercury

removed by the fly ash (Sjostrum et al, 1997; Haythornthwaite et al,

1997; and PSCO/ADA et al, 1997). If this type of fly ash was not

present, the mass carbon-to-mercury ratios could be much higher as

indicated at the tests at the Public Service Electric and Gas

Company’s Hudson station (Waugh et al, August and December, 1997).

These data indicate mercury removals at greater than 90 percent,

but the mass of carbon-to-mercury was still between 20,000:1 and

50,000:1 (116 C or 240 F) for a pulse-jet at an air-to-cloth ratio of o o

approximately 12 ft/min (in this case EPRI’s COHPAC or TOXICON). ESP

pilot-scale tests indicated mercury removals of 83 percent at 105 Co

(221 F) and a mercury removal of 35 percent at 133 C (272 F) at the same o o o

mass carbon-to-mercury ratio of 45,000:1. Low-sulfur Eastern

bituminous coal was fired at the utility and the fly ash mercury
I-27

removals across the range of temperatures was a nominal 15 percent

(Waugh et al, August and December, 1997). 

Mercury mass transfer limitation(s) may be dominant under these

most recent field pilot-scale studies. Small deviations in the

temperature indicate an increase in carbon needed to maintain even low

levels of removal with fabric filters and most indicative, upstream of

an ESP with or without flue gas cooling. Optimizing is not the issue

at this time. Research is needed and the high mass carbon-to-mercury

ratios may not be cost effective, based on the recent data on carbon

injection for mercury removal. The data presented in 1993 by EPRI

(Chang et al, 1993) were extremely innovative, but since then many

improvements have been made to aid in the collection and

interpretation of the data. The methods to measure mercury were not

at the level of today’s standards and the fly ash, based on the recent

tests at the Comanche Station, can account for close to 65 percent of

the mercury removal. Data have been presented that the fly ash alone

can remove >90 percent of the mercury across the Station’s existing

reverse-gas baghouse. This is not typical of the majority of the fly

ashes collected in the utility industry. The recent PSCO data is

collected at the same facility as the 1993 data. The mass carbon-tomercury ratios are higher than indicated in the 1993 work.

Mass carbon-to-mercury ratios of >100,000:1 may be required at

one second or less residence time upstream of an ESP at 149 C (300 F) o o

in order to achieve 90 percent mercury removal. The scenarios for the

ESPs may require fabric filters downstream. The fabric filter of

choice would probably be a pulse-jet filter operating at a high airto-cloth ratio. 

A reverse gas fabric filter is an option in the cost of control

models in Appendix B of the Mercury Study Report being utilized

downstream of an ESP for mercury capturing the injected carbon being

used for mercury removal. A more compact pulse-jet filter could be

utilized for mercury removal and this option would also be effective

for collecting the fine particulate escaping the upstream ESP (e.g.,

EPRI’s COHPAC or TOXICON). Further research is needed to verify this. 

If the ESP is 98.5 to 99 percent efficient (greater than the 0.03

lb/MMbtu NSPS limit), then a considerable amount of particulate (less

than 5 microns) will accumulate or be collected with the injected

activated carbon. This is a benefit, but it could have an impact on

pressure drop and cleaning frequency of the pulse-jet. This could

limit the utilization of the carbon for mercury capture and the

increase of pressure drop would require additional fan power. If the

size of the pulse-jet is at the levels requiring higher air to cloth

ratios between 6 and 8 ft/min or higher, the pressure drop would

increase in a shorter period of time requiring more frequent cleaning

and subsequently the mercury capture would decrease per unit mass of

carbon injected due to less contact time. There are currently

problems with pulse-jet filters as a polishing device while cleaning

on line for the fine particulate (reentrainment of the fine fly ash)

since there is not an adequate dust cake formed. Humidification may
I-28

help, but it has just been tested under this type of application

(Waugh et al, December 1997). The reentrainment issue could further

complicate the problem and demand additional costs for taking the

filter off-line. A design could be provided to recirculate the underutilized carbon and fly ash mixture, which would require an additional

cost of handling of the solids and re-injection. If there is no

recirculation of the carbon collected in the hoppers, then more carbon

would be needed than anticipated. These concepts or designs are in

their infancy and data still need to be collected and carefully

interpreted.

The Department of Energy Federal Energy Technology Center and the

Electric Power Research Institute are planning to conduct several

pilot-scale field studies at different utility sites, with possible

full-scale demonstrations. Before the use of activated carbon for

mercury removal is cost effective in the coal-fired electric utility,

a large collaborative effort, the collection of the data and its

interpretation from all the fundamental, laboratory-, bench-, and

pilot-scale tests being performed must be realized.

I.2.3.2 Test Data on the Effectiveness of Activated Carbon

Injection for Utility Boilers.

Limited test data indicate that activated carbon (AC) injection

effectively reduces mercury emissions when used in conjunction with

existing control devices, such as fabric filters (FFs) and spray dryer

absorbers (SDAs).

Table I-3 presents pilot-scale test data on the mercury removal

efficiency of AC injection when used ahead of FFs. Such a

configuration, with no prior PM control, has a median mercury removal

efficiency that varies with temperature and AC injection rate. With a

low AC injection rate (<1,000 wt C/wt inlet Hg) and an average flue

gas temperature between 107C (225F) and 121C (250F), a median

mercury removal efficiency of 29 percent was found, with a range from

14 percent to 47 percent removal. With a low AC injection rate (same

as above) and an average flue gas temperature between 88C and 107C, a

median mercury removal efficiency of 97 percent was found, with a

range from 76 percent to 99 percent removal. A high AC injection rate

(>1,000 wt C/wt inlet Hg) and an average flue gas temperature between

107C (225F) and 121C (250F) produced a median mercury removal

efficiency of 82 percent, with a range from 69 percent to 91 percent

removal. A high AC injection rate (same as above) and an average flue

gas temperature between 88C (190F) and 107C (225F) produced a

median mercury removal efficiency of 98 percent, with a range from

95 percent to 99 percent removal (Chang et al., 1993).

Table I-4 presents test data for AC injection when used before

SDA systems. Tested SDA/ESP systems with AC injection had a median

mercury removal efficiency of 85.9 percent, with a range from 74.5

percent to 90.9 percent removal (Felsvang, 1993). Pilot-scale testing

of a SDA/FF system with AC injection had a median mercury removal 
I-29

Table I-3. Activated Carbon Injection Before Fabric Filter Dataa

Unit Control Device Hg removal %

Low temperature + low carbon injection rate (< 1000 wt C/wt Hg) based on inlet Hg

Test #4, Run #1 AC + FF (88C (190F) and 216 wt C/wt inlet Hg; inlet Hg 97

concentration of 5.35 g/dscm)

Test #4, Run #2 AC + FF (88C (190F) and 126 wt C/wt inlet Hg; inlet Hg 99

concentration of 8.19 g/dscm)

Test #4, Run #3 AC + FF (91C (196F) and 123 wt C/wt inlet Hg; inlet Hg 97

concentration of 8.62 g/dscm)

Test #6, Run #3 AC + FF (102C (216F) and 727 wt C/wt inlet Hg; inlet Hg 76

concentration of 1.94 g/dscm)

High temperature + low carbon injection rate (< 1000 wt C/wt Hg) based on inlet Hg

Test #5, Run #1 AC + FF (107C (225F) and 362 wt C/wt inlet Hg; inlet Hg 14

concentration of 5.53 g/dscm)

Test #5, Run #2 AC + FF (110C (230F) and 373 wt C/wt inlet Hg; inlet Hg 28

concentration of 4.45 g/dscm)

Test #5, Run #3 AC + FF (116C (241F) and 457 wt C/wt inlet Hg; inlet Hg 47

concentration of 3.47 g/dscm)

Test #6, Run #1 AC + FF (121C (250F) and 286 wt C/wt inlet Hg; inlet Hg 29

concentration of 5.04 g/dscm)

Test #6, Run #2 AC + FF (118C (244F) and 367 wt C/wt inlet Hg; inlet Hg 35

concentration of 4.22 g/dscm)

Low temperature + high carbon injection rate (> 1000 wt C/wt Hg) based on inlet Hg

Test #2, Run #1 AC + FF (91C (196F) and 2843 wt C/wt inlet Hg; inlet Hg 95

concentration not measured but assumed to be 7.00

g/dscm)

Test #2, Run #2 AC + FF (96C (205F) and 3132 wt C/wt inlet Hg; inlet Hg 98

concentration not measured but assumed to be 7.00

g/dscm)

Test #2, Run #3 AC + FF (93C (200F) and 3121 wt C/wt inlet Hg; inlet Hg 98

concentration not measured but assumed to be 7.00

g/dscm)

Test #3, Run #2 AC + FF (93C (200F) and 4361 wt C/wt inlet Hg; inlet Hg 99

concentration of 6.23 g/dscm)

Test #3, Run #3 AC + FF (96C (205F) and 3850 wt C/wt inlet Hg; inlet Hg 99

concentration of 6.91 g/dscm)

(continued)
Table I-3. (Continued)

Unit Control Device Hg removal %

I-30

High temperature + high carbon injection rate (> 1000 wt C/wt Hg) based on inlet Hg

Test #3, Run #1 AC + FF (110C (230F) and 3332 wt C/wt inlet Hg; inlet Hg 91

concentration of 7.95 g/dscm)

Test #7, Run #1 AC + FF (121C (250F) and 1296 wt C/wt inlet Hg; inlet Hg 69

concentration of 4.66 g/dscm)

Test #7, Run #2 AC + FF (121C (250F) and 1954 wt C/wt inlet Hg; inlet Hg 76

concentration of 4.30 g/dscm)

Test #7, Run #3 AC + FF (116C (241F) and 3649 wt C/wt inlet Hg; inlet Hg 87

concentration of 2.09 g/dscm)

 Source: Chang et al., 1993 a

Table I-4. Activated Carbon Injection Before Spray Dryer

Absorption Dataa

Unit Control Device Hg Removal %

SDA/ESP

Denmark AC + SDA/ESP (inlet Hg concentration ranges 80.3, 85.8, 75.8, 74.5, 90.9,

from 3.5 - 7.9 g/dscm) 89.5, 89.3, 86.7, 85.9

SDA/FF

NSP Sherco 3 (pilot AC + SDA/FF (inlet Hg concentration unknown) 50-60

unit)

Plant D2 AC + SDA/FF (inlet Hg concentration of 3.9 >99

g/dscm)

 Source: Felsvang, 1993 a
I-31

efficiency of 60 percent, with a range from 50 percent to 99 percent

removal (Felsvang, 1993).

I.2.3.3 Flue Gas Desulfurization (FGD) Scrubbers. Wet FGD

systems are currently installed on about 25 percent of the coal-fired

utility generating capacity in the United States (Redinger et al,

1997). Although their primary function is to remove SO emissions from 2

boiler flue gas, wet FGD systems can also be effective in removing

mercury emissions from boiler flue gas. The mercury removal

efficiencies of wet FGD systems can vary widely depending on the

mercury species in the incoming flue gas, the design and operation of

the wet FGD system, and reactions of mercury species in the scrubbing

solution.

Mercury Speciation of Incoming Flue Gas. The mercury removal

efficiency of a wet FGD system varies depending on the form or species

of mercury vapor in the incoming flue gas. Mercury in flue gas is

either associated with particulate matter or in the gas phase. In the

United States, most commercial wet FGD systems are used downstream of

ESPs (Redinger et al, 1997). An ESP removes most of the particulatebound mercury from the boiler flue gas before it reaches the wet FGD

system; thus, most of the mercury that enters a wet FGD system is in

the gas/vapor phase. The vapor phase mercury in boiler flue gas is

generally present as elemental mercury (Hg ) or oxidized mercury 0

(HgCl ) (Redinger et al, 1997). The proportion of elemental mercury to 2

oxidized mercury in the flue gas is influenced by a number of factors

such as the type of coal fired in the boiler, fly ash composition,

flue gas temperature, and the presence of other compounds in the flue

gas such as HCl, SO , and NO . Because oxidized mercury is much more 2 x

soluble in the aqueous solution present in a wet FGD system than

elemental mercury, it is more likely to be removed from the flue gas.

Recent studies indicate fly ash and its subsequent interaction(s)

with the vapor phase compounds in the post-combustion zone can

influence a higher proportion of oxidized mercury as compared the

elemental mercury (Carey et al, 1996 and 1997; Hargrove et al, 1997;

Laudal et al, 1996 and 1997; and Senior et al, June and November

1997). The fly ash from the combustion of certain Northern

Appalachian bituminous coals can have a significant impact, resulting

in high levels of the oxidized form of mercury entering the wet FGD

systems. A high conversion (>75 percent) of spiked elemental mercury

into a particle laden flue gas upstream of highly efficient pilotscale pulse-jet FFs was observed at two coal-fired facilities. The

conversion was measured with the Tris-Buffer and Ontario Hydro

speciation measurement methods. There was no apparent conversion of

the spiked elemental mercury measured in the particle free flue gas at

the outlet of the pulse-jet FFs (the FFs particulate control

efficiencies were measured at 99.99 percent) by the Tris-Buffer and

Ontario Hydro methods. 

The coals fired during the separate tests were both N.

Appalachian coals (Pittsburgh Seam/Blacksville and a blend of Ohio No.
I-32

5 and No. 6) that provide a high percentage of natural occurring

oxidized mercury. Bench-scale tests conducted by Radian International

and UNDEERC have indicated that the fly ash from the combustion of

Blacksville coal has the ability to convert elemental mercury to an

oxidized form. The exact vapor phase compounds and subsequent

mechanisms responsible for the conversion are being investigated with

this and other fly ashes. The conversion is less pronounced or not

indicated with PRB and other subbituminous coal fly ashes (Carey et

al, 1996 and 1997; Hargrove et al, 1997; Laudal et al, 1996 and 1997

&12/97; and Senior et al, June and November 1997).

EPRI has reported pilot-scale experience showing significant

capture of oxidized mercury in an ESP/wet FGD system (Chow and Owens,

1994). Approximately 60 percent of the total 10 μg/m of mercury in 3

the flue gas was in the oxidized form. The ESP/wet FGD system

captured all of the oxidized mercury while allowing the elemental

mercury to pass through the scrubber.

Radian conducted a series of pilot scale tests that showed

significant capture of oxidized mercury by a wet FGD system (Noblett,

1993). In these tests, more than 95 percent of the mercury in the

inlet flue gas to the scrubber was in the oxidized form. The scrubber

system removed over 90 percent of the oxidized mercury from the flue

gas while removing little elemental mercury.

FGD pilot testing by Babcock & Wilcox (B&W) with three Eastern

bituminous coals has demonstrated a range of total mercury emissions

reductions across the scrubber with the scrubber operating at constant

conditions (Redinger et al, 1997). With a baghouse/FGD emissions

control configuration, total FGD system mercury emissions control

ranged from 88 percent to 92 percent for the three coals. For the

same coals, with an ESP/FGD system configuration, mercury emissions

reduction across the FGD ranged from 23 percent to 80 percent.

Coal Type. EPRI has published data which show distinct

differences between the forms of mercury in the vapor phase and the

distribution of mercury between the particulate and vapor phases for

bituminous and sub-bituminous coals (Chang, 1994). In general, a

higher level of elemental mercury was observed for sub-bituminous coal

versus bituminous coal at typical wet FGD system inlet temperatures. 

The EPRI data indicated that at 300 F, 68 percent of the total vapor o

phase mercury was present as elemental mercury for the sub-bituminous

coal compared to 6 percent as elemental mercury for the bituminous

coal. This difference in mercury speciation suggests that a wet FGD

system will have a low mercury removal efficiency if it treats flue

gas from a boiler that fires sub-bituminous coal and a high mercury

removal efficiency if it treats flue gas from a boiler that fires

bituminous coal.

Design and Operation of the Wet FGD System. The liquid-to-gas

(L/G) ratio of a wet FGD system impacts the removal efficiency of

oxidized mercury. The L/G ratio of a wet limestone FGD system is
I-33

dictated by the desired removal efficiency of SO . In general, high 2

efficiency (95 percent SO removal) systems are designed with L/G 2

ratios of 120 gal/1000 acf to 150 gal/1000 acf. In an EPRI pilot

study, increasing the L/G ratio from 45 gal/1000 acf to 133 gal/1000

acf increased the removal efficiency of oxidized mercury from 90

percent to 99 percent (EPRI, 1994). In another pilot study by B&W,

increasing the L/G ratio from 37 gal/1000 acf to 121 gal/1000 acf

increased the removal efficiency of oxidized mercury from 91 to 98

percent; increasing the L/G ratio did not affect the removal

efficiency of elemental mercury, which was close to zero percent

(Redinger et al, 1997).

Configuration of the Wet FGD System. Most of the existing U.S.

wet FGD systems have open spray tower or tray tower designs (Redinger

et al, 1997). Recent research has shown that tray tower designs are

more effective in removing oxidized mercury from boiler flue gas than

open spray tower designs at the same operating conditions. In one

study where the composition of the flue gas was mostly oxidized

mercury, total mercury removal efficiencies from a wet FGD system with

a tray tower design ranged from 85 to 95 percent, whereas total

mercury removal efficiencies from a wet FGD system with an open spray

tower design ranged from 70 to 85 percent (removal efficiencies for

both systems increased as their L/G ratios increased from 39 to 122

gal/1000 acf) (Redinger et al, 1997).

Measurement Limitations and Reduction of Oxidized Mercury. A

high proportion of oxidized mercury in the inlet flue gas to a wet FGD

system does not guarantee that the scrubber will have a high total

mercury removal efficiency. Evidence exists that elemental mercury

can be generated in a wet FGD system by reduction of a portion of the

oxidized mercury absorbed in the scrubbing solution. Radian evaluated

mercury removal across a wet FGD system, in which 67 to 95 percent of

the inlet mercury to the scrubber was present in the oxidized form

(Hargrove, 1994). Despite these relatively high levels of oxidized

mercury, the average removal efficiency of total mercury from the

scrubber was only 50 percent. Radian noted possible generation of

elemental mercury across the scrubber. Recent tests by B&W using the

Ontario Hydro method have also noted higher concentrations of

elemental mercury in the outlet of a wet FGD system compared to the

inlet concentrations of elemental mercury. Pilot-scale testing using

the Ontario Hydro method to measure mercury upstream and downstream of

the scrubber has demonstrated the conversion of oxidized mercury

species at the scrubber inlet to elemental mercury across the scrubber

can be minimized by control of the dissolved species in the scrubbing

system slurry (Redinger et al, 1997).

Previous field studies conducted by EPRI and DOE did indicate

higher levels of elemental mercury (Hg ) at the outlet of wet FGD 0

scrubbers relative to the inlet. In addition, the removals indicated

higher than 95 percent of the reported oxidized mercury at the inlet. 

These measurements were reported from separate U.S. EPA Draft Method

29 (M29) samples and in combination with the MESA Method samples. Two
I-34

questions were raised: “Was the U.S. EPA M29 capable of accurately

measuring the oxidized form of mercury?” or “Was the oxidized form of

mercury being captured in the wet FGD scrubber solutions being

released as an “alternate” form not capable of being collected in the

appropriate impinger solutions?” 

Innovative pilot-scale studies were conducted by Radian

International at the EPRI ECTC to address these two questions. 

Extensive flue gas and intra-train mercury spiking tests were

conducted to investigate the acidified peroxide solutions of M29

(solutions for collecting the oxidized form of mercury). The first

series of tests had Hg and HgCl injected separately into the flue gas 0 2

stream at the inlet of the wet FGD. Results indicated 96 percent of

the HgCl (naturally occurring and spiked) was collected across the wet 2

FGD and the increase in Hg across the FGD was from 0.66 to 0.96 μg/m . 0 3

The results for the Hg spiking indicated 37 percent of spike was 0

measured in the acidified peroxide solutions and the total Hg removal

was only 29 percent. These results indicated the injected HgCl was 2

being effectively collected in the scrubber solutions and not being

reduced and subsequently re-emitted as Hg . In addition, M29 was not 0

effective in speciating the mercury at the inlet of this wet FGD

system when Hg was spiked. 0

The intra-train-spiking of either form of mercury into the flue

gas further indicated the inability of M29 to accurately measure the

distribution of the speciated and elemental mercury in the flue gas at

typical conditions upstream of a wet FGD. Radian conducted all of

these initial tests in 1994 and repeated them in 1995; they are

summarized in an EPRI and DOE report (Laudal et al, 1996).

Studies at the UNDEERC have duplicated the results of Radian. 

Recent studies at the UNDEERC indicated an overestimation of the

oxidized mercury of up to 50 percent for M29 and up to 70 percent for

the MESA method. The UNDEERC work has indicated the conditions at the

inlet of wet FGD systems (e.g., high SO concentrations and moderate to 2

high concentrations of NO ) have an impact on the overestimation of the x

oxidized form of mercury - SO for the U.S. EPA M29 and the combination 2

of SO and NO for the MESA. These findings are also detailed in the 2 x

EPRI and DOE report (Laudal et al, 1996).

After two years of evaluating and developing mercury speciation

measurement methods, the UNDEERC has identified the Ontario Hydro

Method as one of the most promising mercury speciation measurement

methods. To obtain the accuracy of the speciated mercury measurement

method, it was necessary to perform U.S. EPA Method 301 validation

procedures with dynamic spiking of mercury in the flue gas stream. 

Spiking was done first with elemental mercury, then with HgCl . 2

Results showed the Ontario Hydro method passed the U.S. EPA Method 301

criteria and was able to collect the form(s) of mercury correctly from

the flue gas. The testing was conducted at the same and higher

levels of SO in the flue gas as compared to the previous validation 2

studies for M29. The Ontario Hydro method was not impacted by the SO
I-35

concentrations as indicated for M29 and the MESA method. The Ontario

Hydro method is being recommended as the best method to measure

mercury speciation in coal-fired systems. The method is being

submitted to the American Society for Testing and Materials (ASTM) and

U.S. EPA for approval (Laudal et al, December, 1997).

The recent pilot-scale speciation measurement evaluation and

development studies and field results with the promising methods

indicate less of an increase in the apparent re-emission of the

captured oxidized mercury. Under certain conditions there has been an

increase of the outlet elemental mercury as compared to the inlet of a

wet FGD system (possible re-emission of the captured oxidized mercury)

while utilizing the Ontario Hydro method (Redinger et al, 1997). 

Further testing at the McDermott facility will be conducted to

determine at what wet FGD conditions the possible re-emission occurs.

I.2.3.4 Spray Dryer FGD Systems. In 1990, spray dryer FGD

systems were installed on approximately one percent of coal-fired

units in the United States (UDI, 1992). The primary function of spray

dryer FGD systems is to remove SO emissions from boiler flue gas; 2

however, they can also be effective in removing mercury emissions from

boiler flue gas.

The effectiveness of a spray dryer FGD system to remove mercury

emissions from boiler flue gas depends on the form or species of

mercury vapor present in the incoming flue gas. In one study, the

removal efficiencies of SO and total mercury from a spray dryer FGD 2

system were 82 percent and 63 percent, respectively; oxidized mercury

represented 73 percent of the total mercury at the scrubber inlet. In

another study, the removal efficiencies of SO and total mercury from a 2

spray dryer FGD system were 68 percent and 64 percent, respectively;

oxidized mercury represented 68 percent of the total mercury at the

scrubber inlet (Redinger et al, 1997). 

I.2.4 Research and Emerging Technologies for Controlling Mercury

Emissions from Utilities

Considerable research continues to develop efficient and costeffective technologies for mercury emission reductions from utility

plants. This section describes ongoing research and summarizes the

results to date. Much of the research is being sponsored by three

organizations: U.S. EPA, DOE and EPRI. Table I-5 lists the areas of

research currently being funded by these groups. 

Eleven Phase I mercury control projects have been completed as

part of DOE’s Advanced Emissions Control Technology “MegaPRDA

Program.” These Phase I efforts began in October 1995 and encompassed

two years of laboratory and bench scale testing and evaluation of

several approaches for controlling the emission of mercury from coalfired utility boilers. The approaches included those listed in Table

I-5. DOE has selected six Phase II proposals (two to three year

efforts) to further investigate and develop fine particulate and 
I-36

Table I-5. Current Mercury Control Research for Utility Boilers

Sponsor Research Area

U.S. Environmental Protection Agency Mercury speciation/High temperature control

Fundamental reactions/Low temperature control

Combined SO /Mercury control 2

U.S. Department of Energy Fundamental and bench-scale investigation of

adsorption and conversion of mercury by fly ash

Fundamental studies & model development to predict

mercury speciation, partitioning, and fate in coal-based

power systems

Fundamental and bench-scale studies on enhanced

sorbents for mercury adsorption

Pilot-scale field studies on sorbent injection for

conventional APCDs

Enhanced removal of oxidized and elemental mercury

in wet FGD systems

Capture of total mercury with regenerable sorbents

Coal cleaning (physical, biological, mild chemical)

Electric Power Research Institute Bench-scale: adsorption of mercury onto fly ash 

Fundamental studies & model development to predict

mercury speciation, partitioning, and fate in coal-based

power systems

Field scale: pilot tests (two sites) of sorbent injection

with ESP’s and fabric filters

Bench scale studies of mass transfer

Wet scrubber controls for mercury

Absorption of mercury in aqueous solution

mercury control technologies and concepts. Given the relative low

maturity level of these technologies, commercial deployment is still

at least several years away, and will be strongly dependent on the

results of the Phase II efforts.

Research continues on developing potential technologies for

mercury emission reduction from utility plants. This research is

aimed at either the addition of some type of sorbent technology to

adsorb the mercury, improving the mercury capture effectiveness of

existing pollution control technology, or using new technology for

mercury control. Before any of the technologies are fully realized
I-37

for utility application, the fundamental mechanisms of the flue gas

and mercury chemistries during the combustion and post-combustion

conditions, along with the various interactions with the different

types of fly ash must be understood (Brown, T.D., 1997).

Research at the fundamental level is being conducted by Physical

Sciences, Inc., to determine the mechanisms involved with both gasphase mercury transformations and the gas-solid interactions.

Attempts have been made to use thermochemical equilibrium

calculations to predict the mercury species in coal combustion flue

gas by using equilibrium calculations (see, for example, the review by

Galbreath and Zygarlicke, 1996). The results of equilibrium

calculations for mercury speciation in flue gas as a function of

temperature can be summarized briefly. Above about 975 K (700 C) 99 o 

percent of the Hg is predicted to exist as gaseous Hg. The rest (1

percent) is predicted to be gaseous HgO. Below 725 K (450 C) all the o 

Hg is predicted to exist as HgCl . Between 725 and 975 K, the split 2

between HgCl and Hg is determined by the chlorine content of the coal 2

(via the HCl content of the gas). HCl concentrations in flue gas from

U.S. coals are typically in the range of 1 to 100 ppm. Even at these

low concentrations, the reaction between Hg and HCl dominates the

equilibrium chemistry. At temperatures representative of the inlet to

the APCD, therefore, all the mercury should exist in the gas phase as

HgCl , if equilibrium is attained in the flue gas. 2(g)

However, there are strong arguments against the existence of

chemical equilibrium in the flue gas of a coal-fired power plant. The

flue gas cools rapidly as heat is transferred to water and steam;

typical cooling rates are on the order of 500 K/s. Minor species in

the flue gas such as CO and SO do not have time to equilibrate as the 2

gas cools. For example, the oxidation of SO to SO in coal combustion 2 3

flue gas does not proceed at a fast rate below about 1500 K (Flagan

and Seinfeld, 1988) and thus the SO concentration is effectively 3

frozen below this temperature in

the flue gas. Similarly for trace species, present in ppm or ppb

amounts, equilibrium may not be attained as the flue gas cools.

Recent kinetic calculations also indicate that the conversion of

another trace species, HCl, to Cl is frozen as the flue gas cools 2

(Senior et al, 1997).

The evidence from pilot-scale and full-scale combustion systems

is not consistent with the assumption of equilibrium for mercury

species in flue gas at the temperatures corresponding to the location

of the air pollution control devices (APCD). At the inlet to the

APCD, measurements in large scale combustion systems indicate that

only about 75 percent of the gas-phase mercury is found as Hg+2

(Prestbo and Bloom, 1990; Fahlke and Bursik, 1995; Meij, 1994). The

range of observed values is broad: one study consisting of mercury

speciation measurements from fourteen different coal combustion

systems reported anywhere from 30 percent Hg to 95 percent Hg +2 +2

upstream of the APCD (Prestbo and Bloom, 1990). There is some
I-38

evidence from laboratory and pilot data that the kinetics of Hg

oxidation are slow at low temperatures. Based on pilot data, the

addition of HCl at temperatures below 450K (180 C) did not increase o

the amount of HgCl in coal combustion flue gas, indicating no reaction 2

at those temperatures (Galbreath and Zygarlicke, 1996). In laboratory

experiments (Nordin et al, 1990) using simulated flue gas (in the

presence of activated carbon), equilibrium was not attained for Hg at

temperatures below 473 K (200 C). o 

The assumption of gas-phase equilibrium for mercury-containing

species in coal-fired power plant exhaust is not valid. Preliminary

evidence suggests that the oxidation of elemental mercury to mercury

chloride in the gas is frozen when the gas cools below 750-900K. 

Kinetic calculations on the formation of Cl , which is highly reactive 2

with elemental mercury, indicate that the conversion of HCl to Cl does 2

not attain equilibrium given the time temperature-history in a power

plant which lends support to the conclusion of frozen equilibrium for

mercury oxidation.

Understanding gas-phase speciation of mercury in coal fired power

plant flue gas is not sufficient to describe the transformations of

mercury in the combustion system. In order to understand the capture

of mercury in APCDs and the effectiveness of sorbents for mercury

capture, better understanding of the gas-to-particle conversion is

also needed, particularly the relationship between fly ash properties

and oxidation and/or adsorption of mercury.

Two key questions can be posed: first, what is the process by

which fly ash (and certain other solids) seem to catalyze the

transformation of gaseous elemental mercury to oxidized forms; second,

what are the mercury species adsorbed on fly ash? Answering both

these questions will require a detailed look at the constituents of

the fly ash and how they interact with mercury at temperatures

characteristic of the flue gas (400-600K) as it enters the APCD.

Gas-phase oxidized mercury is readily captured by activated

carbon, while elemental mercury has a much lower affinity for carbon. 

The surface of the carbon is crucial to mercury sorption; adding

sulfur or iodine can dramatically increase the capacity of activated

carbon for elemental mercury (Dunham and Miller, 1996; Krishnan et

al, 1994; Vidic and McLaughlin, 1996). Residual carbon from coal

combustion is not the same as activated carbon. The pore structure,

surface properties, and inorganic content may be strikingly different. 

Nonetheless, coal char does have some capacity for adsorbing mercury. 

Based on the recent experimental work (Senior et al, 1997), it can be

concluded that the mechanisms for adsorption of elemental and oxidized

mercury on coal char are very different. Properties of the coal char

(surface area, sulfur content, and forms of sulfur) have been shown to

determine the amount of mercury adsorption. In addition to carbon,

there is evidence for the adsorption of mercury on coal fly ash (Carey

et al, 1996) although the specific species which are adsorbed is not

known. 
I-39

In addition to adsorption, laboratory and pilot scale evidence

suggest that solids such as activated carbon and fly ash can act as

catalysts for oxidation of elemental mercury. Kinetic experiments in

a continuous flow reactor showed that the oxidation of elemental

mercury by oxygen only occurred in the presence of activated carbon

(Hall et al, 1991). A series of bench-scale experiments explored the

catalytic effect of solids, including traditional metal catalysts,

activated carbon, and coal fly ash, on the oxidation of elemental

mercury in simulated flue gas in a fixed bed reactor (Carey et al,

1996). The results showed that coal fly ash converted gaseous

elemental mercury to a mixture of gaseous oxidized mercury and

adsorbed mercury at temperatures from 420K to 640K (300 F to 700 F). o o

Fly ash from five different coals was tested. At 420K, 20-50 percent

of the elemental mercury was converted to a gaseous oxidized form,

probably HgCl based on equilibrium considerations, while 0-80 percent 2

was converted to an adsorbed form on the solids. The adsorbed species

might be HgCl , HgO, or HgSO . There was a wide variation in the 2 4

amount of adsorbed mercury depending on coal type. At 640K, less

elemental mercury was typically converted.

Information on the reactions of mercury species with fly ash can

be obtained by identifying specific mercury species on the surface of

char or carbon and then inferring the reaction pathway. Preliminary

analysis of the forms of mercury on four carbon-based sorbents as

described in PSI et al (1997) was recently completed (Huggins et al,

1997). These samples were treated with a simulated flue gas

containing N , O , CO , SO , H O, HCl, and elemental mercury. In order 22 2 22

to better understand the forms of adsorbed mercury, X-ray absorption

fine structure (XAFS) spectra were collected at the mercury L edge at III

approximately 12,284 eV at the Stanford Synchrotron Radiation

Laboratory. By combining both the XANES and EXAFS evidence, one could

speculate that the Hg bonding in the three different mercury sorbents

is different. In the iodine-impregnated activated carbon, the mercury

bonding appears consistent with Hg-I. In the sulfur-impregnated

carbon and the lignite-based activated carbon, the bonding is more

consistent with Hg-Cl or Hg-S. Further study, particularly of the Cledge XAFS spectra in the SAC and LAC samples is required.

Thus, particulate matter can promote oxidation of elemental

mercury and can collect a significant amount of mercury in flue gas. 

The amount retained in the particulate matter seems to depend on the

following factors:

 carbon content

 properties of the carbon surface

 inorganic constituents in carbon particles

 Hg speciation in the flue gas.

I.2.4.1 Sorbent Technology. Research continues on developing

potential technologies for mercury emission reduction from utility

plants. Although sorbent injection with activated carbon has been

shown to be a promising technology, even greater mercury removal may
I-40

be possible with impregnated activated carbons, sodium sulfide, and

other types of sorbents. The application of an activated carbon

circulating fluidized bed (CFB) also shows promise in removing

mercury.

With sulfur-impregnated activated carbon injection, the

carbon-bound sulfur reacts with mercury to form mercuric sulfide (HgS)

on the carbon, which is then removed by a particulate control device. 

In a pilot-scale study, sulfur-impregnated carbon increased mercury

removal to 80 percent, an increase of 25 percent over results achieved

with an equal amount of nonimpregnated activated carbon (Felsvang et

al, 1993).

Sulfur-impregnated carbons can potentially be enhanced for

mercury sorption by the impregnation of the carbon(s) with sulfur at

elevated temperatures of 400 - 600 C (752 - 1112 F). This has promoted o o

a more uniform distribution of short linear chains of sulfur

allotropes (S and S ) on the carbon surface as opposed to having 2 6

predominately S rings condensed in the macropore region of 8

commercially available sulfur impregnated carbons. In addition, the

sulfur impregnated carbons prepared at elevated temperatures have

exhibited significantly better thermal stability since no sulfur loss

was observed even after exposure at 400 C (752 F). The sulfur o o

impregnated carbons exhibited high elemental mercury uptake efficiency

at 140 C (284 F) when compared to commercially available activated o o

carbons. Dynamic adsorption capacity of these carbons were measure as

high as 4,000 μg Hg/g C. This capacity is almost three orders of

magnitude greater then the capacity of virgin activated carbon and an

order of magnitude greater than the capacity of commercially available

impregnated activated carbon (Vidic et al, 1996; Korpiel et al, 1997;

and Liu et al, 1997).

With iodide-impregnated activated carbon injection, the carbonbound iodide reacts with mercury to form mercuric iodide (HgI ) on the 2

carbon, which is then removed by a particulate control device. In a

pilot-scale study, iodide-impregnated carbon increased mercury removal

to nearly 100 percent, an increase of 45 percent over results achieved

with an equal amount of non-impregnated activated carbon (Felsvang et

al, 1993).

A study by the UNDEERC, as part of a Cooperative Agreement with

the DOE-FETC, found that iodide-impregnated activated carbon was

effective at removing mercury in a test combustor. Removal

effectiveness using the iodide-impregnated activated carbon exceeded

99 percent. Other sorbents tested were steam-activated lignite,

thermal-activated bituminous coal, chemical-activated hardwood, iodine

impregnated, steam-activated coconut shell, and sulfur-impregnated

steam-activated bituminous coal (UNDEERC, 1995).

Chloride-impregnated activated carbon injection has only been

tested on MWCs in Europe. The chloride reacts with mercury to form

HgCl on the carbon, and the carbon is removed by a particulate control 2
I-41

device. Experiments have shown that impregnating activated carbon

with chloride salts increases adsorptive capacity of the activated

carbon by a factor of 300 (Teller and Quimby, 1991).

Public Services Company of Colorado (PSCo) has investigated the

application of dry-sorbent injection for controlling mercury emitted

from coal-fired boilers. A number of sorbents, including activated

carbon, sulfur- and iodine-impregnated carbons, several proprietary

sorbents, and high-carbon fly ash, were screened in the laboratory

prior to pilot-scale testing. Two activated carbons have been tested

on a pilot-scale facility drawing flue gas from PSCo's Comanche

Station in Pueblo, Colorado under pulse-jet and reverse-gas FF-, and

ESP-configurations. American Norit Companies’ Darco FGD, an activated

carbon derived from lignite which has been utilized in the control of

mercury from municipal solid waste combustors, was tested. The second

sorbent is an activated carbon prepared from a bituminous coal

(Feeley, 1997). 

Parameters of flue gas temperature and carbon residence time were

varied to cover a wide range of utility conditions. The effects of

fly ash were also evaluated by pulling flue gas from the upstream and

downstream side of the existing reverse gas baghouse with carbon

injected in the slipstream prior to the inlet of the pilot-scale

configuration being tested. Elemental mercury had to be spiked

upstream of the pilot-scale unit due to low mercury concentrations of

the native flue gas stream.

The results indicate a high level of carbon is needed to remove

the mercury, but deceasing the temperature (either by heat exchangers

or spray cooling with water) had a net increase of the mercury

captured by both the injected carbon and the native fly ash. The

fabric filter configurations had the greatest removals up to 90

percent, but at high carbon injection rates. The ESP results indicate

removals of 50 percent with approximately 30 percent of the total

removal due to the native fly ash with the mass carbon-to-mercury

ratios greater than 5000:1. The test results for all the

configurations are summarized under Section 2.3.1.2, “Current

Research on Activated Carbon Injection for Utilities” (Sjostrum et al,

1996; Haythornthwaite et al, 1997; and PSCO/ADA et al, 1997).

Other innovative activated carbon injection studies have been

conducted by ADA Technologies for EPRI at Public Service Electric and

Gas Company’s (PSE&G) Hudson Unit 2 located in Jersey City, New

Jersey. The results also indicate a high level of carbon is needed to

remove the mercury, but decreasing the temperature caused a net

increase in the mercury captured by the injected carbon, but not for

the native fly ash. EPRI’s COHPAC or TOXICON configurations and a

pilot-scale ESP were tested with the Darco FGD activated carbon. The

test results for the different configurations are also summarized

under Section I.2.3.1.2, “Current Research on Activated Carbon

Injection for Utilities” (Waugh et al, 1997).
I-42

All the current work indicates the removal of mercury is mass

transfer limited in the various flue gases produced from the

combustion of coal. The reasons for this limitation are the low

concentrations of mercury present in the relatively high volumes of

flue gas. There are higher concentrations of other species competing

and occupying the active sites of the carbon. In addition, the flue

gas residence time upstream of an ESP is nominally one second or less

with flue gas velocities in the range of 50 to 60 ft/sec at 149 Co

(300 F). Compounding the mercury mass transfer limitation(s) is the o

decrease in the carbon reactivity and capacity at this nominal, but,

high temperature. Fundamental studies have been performed in the past

two years designed to understand the mechanisms impacting the mercury

mass transfer limitation(s) (Carey et al, 1996 and 1997; Vidic et al,

1996; Rostam-Abadi et al, 1997; Korpiel et al, 1997; and Liu et al,

1997).

Another technology with potential for improving mercury

collection efficiency combines calcium hydroxide (Ca(OH) ) with 2

activated carbon. This reagent, consisting of approximately 95 to 97

percent lime and 3 to 5 percent activated carbon, is known under the

product name Sorbalit (Nebel and White, 1991). Sorbalit has only ® ®

been tested on European MWCs and MWIs.

While sulfur-, iodide-, chloride salt- and Ca(OH) -impregnated 2

activated carbons show promise for increasing the mercury removal

efficiency, the cost of these modified carbons can be as much as 20

times higher than that of unmodified activated carbon (Maxwell, 1993). 

In addition, chemically impregnated carbons may increase the

reactivity and subsequent capture of mercury, but very few studies

have indicated the effectiveness of chemically impregnated carbons for

in-flight capture of mercury (especially at one second or less

residence time) (Vidic et al, 1996; Korpiel et al, 1997; and Liu et

al, 1997). These carbons, while being cost prohibited for in-flight

mercury removal, can possibly be designed for high mercury adsorption

capacities indicative of long contact times (carbon beds or fabric

filters - pulse-jet, if installed downstream of an existing ESP). 

The effectiveness of FF-configurations downstream of an ESP must be

further investigated.

 Argonne National Laboratory is investigating potentially lowcost, chemically treated, solid sorbents, such as volcanic pumice, as

an economical alternative to activated-carbon injection. In addition,

Argonne is planning to assess several key, ancillary issues that may

impact the potential use of these sorbents to control mercury,

including the effect of the sorbents on particulate control equipment

performance, fly-ash marketability, and by-product disposal (Feeley,

1997).

Mercury reduction has been achieved at MWCs through the injection

of Na S solution into the flue gas prior to the acid gas control 2

device. The specific reactions of Na S and Hg are not totally clear 2

but appear to be (Nebel and White, 1991):
I-43

Hg (gas) + Na S + 2H O 22 2  HgS (Solid) + 2NaOH + H and

HgCl (gas) + Na S 2 2  HgS (Solid) + 2NaCl.

The resulting solid, HgS, can be collected by a FF.

There are several potential limitations to Na S injection. These 2

include reaction of Na S with calcium (Ca) in the sorbent (as found in 2

Sorbalit ) to form calcium sulfide (CaS), reduction of the amount of ®

sulfur available to react with mercury (CaS can also cause scaling of

the sorbent feed line), corrosion of ductwork (Na S is a corrosive 2

material), clogging and plugging of the screw conveyor due to

solidification of Na S, and sludge formation due to the presence of 2

inorganic salts in the mixing water (Nebel and White, 1991).

At present, full-scale operational injection of Na S has been 2

done only in MWCs. No plans have been announced to test this

technology on fossil fuel-fired electric steam-generating units.

Sorbent Technologies is marketing a sorbent called Mercsorbent

(Nelson et al, 1997). The company claims that the sorbent is

effective in removing elemental mercury at high temperatures typical

of utility flue gas, and is unaffected by common co-existing flue

gases, such as SO , HCl, and H O. Mercsorbent can be used for sorbent 2 2

injection or it can be used as a coating on a FF. A bench-scale ductinjection system at Sorbent Technologies facilities is now being used

to test Mersorbents with this approach. The company is also scheduled

to demonstrate the sorbent at the refuse incinerator in Fort Dix, New

Jersey, in 1997; prior compliance sampling at this facility suggests

that a significant amount of its mercury is in the elemental form. A

coal-fired boiler or slipstream is also being sought for a test of the

new sorbent material. 

Another potential process for the reduction of mercury emissions

is the use of activated carbon in a CFB (Clarke and Sloss, 1992). In

a CFB, the activated carbon is continuously fed to the reactor where

it is mixed with the flue gas at a relatively high velocity, separated

in the subsequent FF and recycled to the reactor. A small part of the

used activated carbon is withdrawn from the process and replaced by

fresh material (Riley, 1991). The main advantages to CFB's over fixed

carbon beds are the increased flue gas-to-carbon contact area and the

smaller overall pressure drop. This system has been used in Germany

for MWC operation.

In the United States, Environmental Elements Corporation has been

developing and testing a CFB promoting agglomeration of fine

particulate matter, allowing for their capture in an ESP. In

addition, a single injection of iodine-impregnated activated carbon

was added to the fluid bed to adsorb mercury vapor. High residence

time, due to the recirculation of the particles, allows for effective

utilization of the carbon and high collection of the fine particles. 

Results from the laboratory-scale testing indicate spiked elemental
I-44

mercury was significantly reduced when passed through the fluidized

bed of fly ash (50 percent mercury removed) and further reduced to

essentially zero when the activated carbon was injected into the bed

(25 μg/m to zero) at 110 C (230 F). The iodine-impregnated 3 o o

activated carbon was fully utilized after >2 hours within the bed. An

adsorption capacity was calculated to be 770 gm/gm for the carbon and

480 gm/gm for the bed of ash. The ash still was able to remove 30

μg/m after 100 percent breakthrough (carbon fully utilized) was 3

indicated for the carbon. The unit needs to be tested on actual flue

gas from coal combustion, and there are plans to install a pilot unit

and conduct testing at Public Service Electric and Gas’s Mercer

Station (Feeley, 1997). 

I.2.4.2 Improving the Mercury Capture Efficiency of Existing

Pollution Control Technology. Research on improving the mercury

capture efficiency of existing pollution control technology can be

categorized as an investigation of either mercury removal with wet FGD

systems or particulate control technology for capturing mercury.

Enhancing Mercury Removal by Wet FGD Systems. Argonne National

Laboratory is investigating several additives that combine strong

oxidizing properties with relatively high vapor pressures to enhance

the capture of mercury in a wet scrubber. Due to a much higher

solubility compared to elemental mercury, oxidized mercury is readily

removed in a wet scrubber. Experimentation is continuing on the

effect of solutions of chlorine, bromine, and iodine on the conversion

and removal of elemental mercury in a laboratory-scale reactor. Of

the three halogen species tested to date, the chlorine solution

appears to remove the most elemental mercury in the presence of SO and 2

NO. Further testing of these and possibly other oxidizing reagents is

planned (Feeley, 1997).

Radian International LLC has also investigated the conversion of

vapor-phase elemental mercury to more soluble Hg at the bench- and ++

pilot-scales. Radian screened a number of catalysts and coal-based

fly ashes for their ability to oxidize elemental mercury, including

the effect of flue gas temperature, flue gas vapor phase compounds,

and residence time on the oxidation potential of the materials. 

Bench- and pilot-scale testing of iron-based catalysts, various

carbons, bituminous, subbituminous, and lignite fly ash have been

performed on a slipstream of flue gas at the EPRI Environmental

Control Test Center (ECTC) in Barker, New York. In addition, benchscale testing has been conducted at an utility firing a coal producing

a higher percentage of elemental mercury in the flue gas as compared

to the ECTC.

To date, the pilot-scale tests have shown the carbon-based

catalyst to be the most effective in converting elemental mercury to

Hg . Further testing of the carbon catalysts is being planned at ++

three utility sites at the bench-scale. Flue gas composition,

interaction with the fly ash, and temperature will be the variables. 

Deactivation of the catalysts will be investigated with reactivation
I-45

concepts being initiated. The tests will be designed to determine the

long-term capabilities of the catalysts, with testing being conducted

over a six month period of performance for all the catalysts. The

influence of sulfur and nitrogen oxides, HCL, and other vapor phase

compounds will be investigated. Converting elemental mercury into an

oxidized form could be advantageous in reducing mercury emissions with

existing technologies (Carey et al, 1996 and 1997; Hargrove et al,

1997; and Radian International et al, 1997).

Improving Particulate Control Technology for Capturing Mercury. 

Research into improving the existing particulate control technology

for capturing mercury is being conducted by several companies.

ABB Power Plant Laboratories is developing retrofittable

enhancements to existing ESPs to increase their efficiency in

capturing fine particles and air toxics. Several approaches to

improving the capture of fine particulates have been investigated. 

The most significant results were through flue gas cooling

(humidification and heat exchange) and in combination with pulsed

energization. The pulsed energization was accomplished through an ABB

proprietary transformer rectifier set - Switched Integrated Rectifier

(SIR). Flue gas cooling in combination with the SIR provided particle

reductions from 45 mg/m to less than 5 mg/m (<0.005 lbs/MMBtu) at a 3 3

gas temperature of 150 C (300 F). The particles in the 2.5 micron o o

range and less were effectively reduced by a factor of 10 to 20. 

Preliminary tests indicated a reduction between 40 and 50 percent of

the mercury in the flue gas by the native fly ash, which is

encouraging for both the low-sulfur bituminous and subbituminous

coals. This approach shows promise in improving the collection of

particulate-bound mercury, and may also cause vapor-phase mercury to

condense on particulate matter and be captured in the ESP. Future

work entails scaling the technology and testing under a variety of

coals and further investigating activated carbon injection with flue

gas cooling. Potential impacts on fine particle collection will be

monitored during all phases of testing (Feeley, 1997; Srinivasachar

and Porle, 1997; and ABB et al, 1997).

The performance of conventional control technology in reducing

the emissions of mercury from coal-fired boilers is being evaluated in

pilot-scale studies as part of Babcock & Wilcox's Advanced Emissions

Control Development Program (AECDP). Phase I of the AECDP involved

benchmarking the mercury capture performance of an ESP, a baghouse,

and a wet scrubber installed at B&W’s Clean Environment Development

Facility (CEDF). The focus of Phase II was to optimize the mercury

removal capability of the conventional pollution control technologies. 

The results of the work conducted in 1996 and 1997 were detailed in

the sections under “2.3.2, Flue Gas Desulfurization (FGD) Scrubbers”

(Feeley, 1997; Redinger et al, 1997; and Holmes et al, 1997).

Phase III of the program will be directed at the development of

new air toxics emissions control strategies and devices. Further

testing at the McDermott facility will be conducted to determine at
I-46

what wet FGD conditions the possible re-emission of captured oxidized

mercury occurs.

Under DOE funding, the Energy and Environmental Research Center

together with W.L. Gore and Associates is developing a new technology

for ultrahigh collection of fine particles, including the difficultto-collect trace element enriched submicron fraction. The concept

utilizes electrostatics and filtration in a unique manner that

provides over 99.99 percent fine particle collection in a device that

is up to 75 percent smaller than conventional technologies. The

approach also shows promise for collecting vapor-phase trace elements

such as mercury and selenium when combined with an effective sorbent. 

The concept will be scaled up for testing on a variety of coals under

various operating conditions (Miller et al, 1997; and UNDEERC et al,

1997).

I.2.4.3 New Technology for Controlling Mercury. A new

technology for controlling all forms of mercury from coal-fired

electric utility units has been investigated at the laboratory- and

bench-scales on simulated and on actual flue from coal combustion. 

ADA Technologies has been developing a technology utilizing a

regenerable sorbent allowing for the recovery of liquid elemental

mercury from the flue gas and appropriately called the Mercu-RE

process. The process takes mercury from flue gases and produces

liquid, elemental mercury with no secondary wastes. Noble metals are

used to adsorb mercury at typical flue gas temperatures. The mercury

is then thermally desorbed. 

Results from laboratory tests indicate that a gold-coated

monolith captured virtually all of the elemental mercury injected into

a simulated flue gas. Bench-scale tests on actual flue gas from the

combustion of four different coals showed the regenerable sorbent is

capable of removing 95 percent of both elemental and oxidized forms of

the mercury at temperatures between 150 C (300 F) and 204 C (400 F). oo oo

The unit ran for more than 700 hours and consistently reduced the

mercury (both forms) in the flue gas from inlet concentrations

averaging 10 g/m to less then 1 g/m at the outlet after more than 3 3

20 sorption-desorption cycles at Consol’s research facility in

Library, Pennsylvania. Further testing of the gold monoliths will

include repeated sorption and desorption cycles over longer-term

testing periods at different operating conditions and at a larger

scale (Feeley, 1997; Roberts and Stewart, 1996; Roberts and Stewart,

1997; ADA Technologies, Inc., et al, 1997).

Based on condensing heat exchanger technology, Babcock & Wilcox

is developing an integrated flue gas treatment system for recovering

waste heat and removing SO , SO , particulates, and trace elements from 2 3

coal combustion flue gas. The condensing heat exchanger is a twopass, counter-flow shell and tube heat exchanger. The hot flue gas

enters the top and flows downward through the first cooling stage,

across a horizontal transition region, and then upward through the

second cooling stage. An alkali reagent is sprayed from the top of
I-47

the second stage to aid in the removal of SO . Testing of the 2

technology was conducted at B&W’s research facility in Alliance, Ohio. 

Preliminary results indicate that total mercury removal across both

stages of the condensing heat exchanger is about 62 percent when

firing a blend of Ohio coals. Testing has been conducted on two other

bituminous coals with similar or higher mercury removals (Feeley,

1997).

The Enhanced Limestone Injection Dry Scrubbing (E-LIDSTM) process

combines furnace limestone injection with dry scrubbing to achieve

high efficiency SO particulate, and trace element emissions control. 2

Dry, pulverized limestone is injected into the upper furnace region of

the boiler. The limestone is calcined to lime and a portion of the

sorbent reacts with SO in the flue gas. The flue gas passes through a 2

particulate collector ahead of the dry scrubber to remove some of the

solids from the gas stream. The solids are mixed with material

collected in the baghouse to produce the SO scrubbing reagent for the 2

spray dryer.

Application of the E-LIDSTM system when firing an Ohio bituminous

coal in the Clean Environment Development Facility (CEDF) at the

Alliance Research Center of McDermott Technology, Incorporated, has

shown efficient emissions control performance. Sulfur dioxide

emissions generated from firing the nominal 3 percent sulfur coal were

reduced by more than 99 percent to less than 0.10 lbs SO /10 Btu. 2 6

Total mercury emissions were reduced from an uncontrolled level of

17.6 g/dscm to less than 0.2 g/dscm for an average total removal

efficiency of greater than 95 percent from the as-fired coal mercury. 

The measured performance confirmed earlier results obtained in the

5 x 10 Btu/hr small boiler simulator (SBS) facility. Mercury 6

measurements upstream of the dry scrubber indicated that both the

limestone injection and operation of the spray dryer/baghouse system

at close to the saturation temperature contributed to the observed

total mercury emissions reduction. The furnace limestone injection

alone reduced mercury emissions to an average of 3.1 g/dscm (Redinger

et al, 1997).

Environmental Elements Corporation is developing a process for

mercury control through DOE's Small Business Innovative Research

program. The first concept utilizes an intense corona discharge to

convert Hg to mercuric oxide. The process also produces SO to serve 0 3

as a conditioner for high-resistivity fly ash. A corona discharge in

coal combustion flue gas will produce oxidizing radicals, such as OH

and atomic oxygen. Bench-scale results indicate that the corona

reactor, operating at relatively low power levels and short residence

time, yielded high elemental mercury vapor oxidation. The mercuric

oxide, in the form of a solid particle, was removed using conventional

particulate control technology. The corona reactor may also convert

mercuric chloride to mercuric oxide, allowing for its capture as well. 

The system is currently being tested on a slipstream at Alabama

Power's Plant Miller (Feeley, 1997).
I-48

The capture of mercury on solid surfaces such as fly ash is being

studied by UNDEERC and DOE-FETC. Data have shown wide variation in

the amount of mercury that can be collected on fly ash associated with

particulate control devices. On occasion, very high levels of capture

have been observed in the presence of HCl separately and in

combination with nitrogen oxides. A number of possible interactions

between vapor-phase mercury and solid surfaces can occur, including

chemical adsorption, physical adsorption, and condensation. However,

the exact mechanisms of capture remain unknown. Research is being

conducted by UNDEERC to elucidate these mechanisms in order to better

define control strategies for mercury in coal combustion flue gases

(Brown, 1997).

There are plans to investigate the interaction of mercury with

metals such as zinc, silver, tin, and cadmium. Mercury has been shown

to amalgamate, rather than adsorb, when in contact with certain

metals. Both experimental and modeling efforts are planned to

determine the suitability of metals for the capture of mercury

(Feeley, 1997).
I-49

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Armstrong, M., 1994. March 1994 notes from Mike Armstrong, Arkansas

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Attari, A. and S. Chao, 1993. "Quality Survey of Natural Gas in the

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I-50

Carey, T. R., Hargrove, O. W., Brown, T. D., and Rhudy, R. G.

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Chen, S., M. Rostam-Abadi, and R. Chang (1996). “Mercury Removal from

Combustion Flue Gas by Activated Carbon Injection: Mass transfer

Effects,” in Proc. Am. Chem. Soc., New Orleans, LA, March 23-28.

Chow, W., et. al, 1994. "Pathways of Trace Elements in Power Plants:

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