r, EPA
United States
Environmental Protection
Agency
Office of Air Quality 453/R-98-004b
Planning and Standards February 1998
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
U.S. Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th Floor
Chicago, IL 60604-3590
-------
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-l
Appendix B Matrix of Electric Utility Steam-Generating Units
and Emission Test Sites B-l
Appendix C Listing of Emission Modification Factors for Trace
Elements Used In the Individual Boiler Analysis . . C-l
Appendix D Discussion of the Methodology Used to Develop
Nationwide Emission To.tals D-l
Appendix E Health Effects Summaries: Overview E-l
Appendix F Documentation of the Inhalation Human Exposure
Modeling for the Utility Study F-l
Appendix G Data Tables for Dioxin Multipathway Assessment . . . G-l
Appendix H Literature Review of the Potential Impacts of
Hydrogen Chloride and Hydrogen Fluoride H-l
Appendix I Mercury Control Technologies 1-1
-------
Appendix A - Median Emission Factors, Determined from Test Report
Data, and Total 1990, 1994, and 2010 Emissions, Projected with the
Emission Factor Program
-------
Table A-l. 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
HAPs
Antimony
Arsenic
Beryllium
Hydrogen chloride
Hydrogen cyanide (HCN) a
Hydrogen fluoride
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Phosphorus (P) b
Selenium
Estimated total 1990
emissions (tons)
7.95
60.93
7.13
143,000
240.66
19,500
3.33
73.27
21.21
75.47
163.97
45.80
58.05
270.74
153.83
Estimated total 1994
emissions (tons)
7.98
55.81
7.93
134,000
250.8
23,100
3.15
61.60
22.67
61.77
167.72
51.34
52.04
331.41
183.68
Estimated total 2010
emissions (tons)
9.93
70.61
8.20
155,000
318.31
25,700
3.82
87.43
27.08
86.89
219.02
59.74
68.65
358.09
213.21
Nationwide hydrogen cyanide emissions were determined from stack
emission factors and not from EMFs.
Nationwide phosphorous emissions were determined from stack emission
factors and not from EMFs.
A-l
-------
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:
inorganic HAPs
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Hydrogen chloride
Hydrogen fluoride
Lead
Manganese
Mercury
Nickel
Phosphorus (P) a
Selenium
Estimated total 1990
emissions (tons)
5.02
0.46
1.71
4.74
20.41
2860
143
10.58
9.28
0.25
392.83
67.25
1.65
Estimated total 1994
emissions (tons)
3.51
0.40
1.09
3.91
15.48
2100
284
8.92
7.30
0.19
322.37
50.89
1.42
Estimated total 2010
emissions (tons)
2.54
0.23
0.86
2.40
10.31
1450
73
5.35
4.70
0.13
198.17
34.10
0.84
Nationwide phosphorous emissions were determined from stack emission
factors and not from EMFs.
A-2
-------
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:
inorganic HAPs
Arsenic
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Phosphorus
Estimated total 1990
emissions (tons)
0.15
a
a
0.14
0.43
a
0.0015
2.19
5.65
Estimated total 1994
emissions (tons)
0.18
a
a
0.15
0.47
£
0.0017
2.42
6.23
Estimated total 2010
emissions (tons)
0.25
3
S
0.23
0.68
2
0.0243
3.49
8.98
The emission factors are not available for this compound, but the
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
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Appendix B - Matrix of Electric Utility Steam-Generating Units and
Emission Test Sites
Table B-l 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-l 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.
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B-5
-------
Appendix B: References
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 NOX control (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.
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. SRI-
ENV-93-1049-7960. December 1993.
B-6
-------
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-7
-------
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-8
-------
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 19927 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-9
-------
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 Institxite, 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-10
-------
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-l. 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.
B-ll
-------
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.
-------
Table C-l. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Circulating Fluidized Bed Furnaces
(Coal-Fired)
Unit Name
Arsenic
Beryllium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRISitelO
1.00
0.77
0.40
1.00
0.49
0.59
1.00
1.00
1.00
NSP - Black Dog #2
0.59
0.41
0.54
0.36
0.68
1.00
0.45
0.71
EMF (Geometric
mean)
0.77
0.56
0.46
1.00
0.42
0.63
1.00
0.67
0.84
Geometric standard
deviation
1.44
1.56
1.25
N/A
1.24 •
1.11
1.00
1.76
1.27
Table C-2. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Tangentially-fired, Dry-bottom Furnace with
NOX Control (Coal-Fired)
Unit Name
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRISite11
0.92
0.79
0.35
0.72
0.92
1.00
0.98
0.25
1.00
EPRI Site 110
w/NOx control
0.66
0.39
0.43
0.70
1.00
1.00
0.36
0.76
0.97
0.82
DOE - Coal
Creek
0.01
0.69
0.35
0.11
1.00
0.61
0.29
0.59
0.85
1.00
0.38
DOE-
Springerville
0.03
0.29
0.87
1.00
0.68
0.73
0.19
0.72
1.00
0.70
0.93
EMF
(Geometric
mean)
0.06
0.52
0.57
0.41
0.84
0.80
0.37
0.75
0.92
0.64
0.74
Geometric
standard
deviation
4.74
1.11
1.08
1.42
1.02
1.02
1.21
1.02
1.00
1.18
1.08
C-l
-------
Table C-3. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Tangentially-fired, Dry-bottom Furnace Without
NOX Control (Coal-fired)
Unit Name
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site 15
0.60
0.54
0.01
0.58
0.94
1.00
0.81
0.43
0.70
EPRI Site 110
0.55
0.89
0.93
1.00
1.00
0.93
1.00
0.71
0.66
0.91
0.58
DOE - Yates
0.67
1.00
1.00
1.00
1.00
0.97
1.00
1.00
1.00
0.84
0.70
EMF (Geometric
mean)
0.61
0.81
0.80
0.23
0.84
0.95
1.00
0.83
0.81
0.69
0.66
Geometric
standard
deviation
1.01
1.02
1.04
8.72
1.03
1.00
1.00
1.01
1.02
1.06
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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI
Site 12
1.00
1.00
0.14
1.00
1.00
1.00
1.00
0.74
(Notel)
0.29
1.00
EPRI
Site 14
0.50
0.92
0.02
0.67
1.00
0.79
0.93
0.74
0.37
0.05
NSP-
Sherburne
#3
0.79
0.58
0.99
0.49
0.49
1.00
0.67
0.21
EPRI
Site 111
0.11
0.05
0.20
0.02
EPRI Site 16
w/OFA and
LNOX Burners
0.80
1.00
0.82
0.11
0.69
0.66
0.66
0.88
1.00
0.54
0.37
EPRI Site
16 W/OFA
1.00
1.00
1.00
1.00
0.58
0.61
1.00
0.60
0.64
0.33
1.00
EMF
(Geometric
mean)
0.90
0.59
0.85
0.16
0.55
0.80
0.76
0.84
0.81
0.24
0.33
Geometric
standard
deviation
1.17
2.41
1.25
4.68
1.72
1.31
1.35
1.26
1.22
2.25
3.51
Note 1 - This EMF was obtained from the mercury retest.
C-2
-------
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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
NSP-
Riverside
#6-7
0.20
0.99
0.40
0.25
1.00
0.19
0.77
1.00
0.78
1.00
DOE-
Boswell
0.59
0.23
0.60
1.00
1.00
0.98
0.42
0.57
0.87
1.00
0.14
1990 EMF
(Geometric
mean)
0.34
0.48
0.49
0.50
1.00
0.98
0.28
0.66
0.93
0.88
0.37
1990
Geometric
standard
deviation
1.16
1.30
1.02
1.27
1.00
1.00
1.08
1.01
1.00
1.00
1.63
EPRI Site
116
0.12
0.7
0.35
0.12
0.27
0.3
0.26
0.21
0.97
0.23
0.47
1994
Geometric
Mean
0.24
0.54
0.44
0.31
0.65
0.54
0.28
0.45
0.94
0.56
0.40
1994
Geometric
Standard
Deviation
1.93
1.76
1.08
3.18
1.77
1.42
1.16
1.58
1.01
1.86
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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site
102
0.48
0.04
0.02
0.25
0.21
0.61
0.33
1.00
0.30
0.65
NSP-
Riverside
#8
0.61
0.51
0.08
0.16
0.22
0.38
0.13
1.00
0.12
1.00
EPRI Site
114
0.15
0.15
0.01
0.30
0.50
0.20
0.73
0.72
0.26
DOE-
Niles #2
1.00
0.58
0.26
0.11
0.28
0.20
0.56
0.15
1.00
0.29
1.00
DOE-
Niles #2
w/NOx
Control
0.92
0.85
0.30
0.20
0.35
0.33
0.60
0.19
1.00
0.39
0.92
DOE-
Baldwin
0.32
1.00
0.44
1.00
0.42
0.34
0.53
0.20
0.92
0.43
0.04
EMF
(Geometric
mean)
0.65
0.51
0.16
0.10
0.29
0.26
0.52
0.19
0.93
0.33
0.43
Geometric
standard
deviation
1.02
1.20
1.38
3.79
1.02
1.03
1.01
1.04
1.01
1.15
1.96
C-3
-------
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
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site 115
0.61
0.52
0.58
0.57
1.00
0.38
0.58
0.78
0.64
0.34
EMF (Geometric mean)
0.61
0.52
0.58
0.57
1.00
0.38
0.58
0.78
0.64
0.34
Geometric standard
deviation
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site 11 4, NOX
0.25
0.15
0.01
0.23
0.84
0.18
0.54
0.31
1.00
EMF (Geometric mean)
0.25
0.15
0.01
0.23
0.84
0.18
0.54
0.31
1.00
Geometric standard
deviation
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
C-4
-------
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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
DOE - Cardinal
0.08
0.91
0.96
1.00
0.61
0.96
1.00
0.27
0.41
0.76
0.07
EMF (Geometric mean)
0.08
0.91
0.96
1.00
0.61
0.96
1.00
0.27
0.41
0.76
0.07
Geometric standard
deviation
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
C-5
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C-6
-------
Table C-ll. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Opposed-fired, Dry-bottom Furnace Without Nox
Control (Oil-fired)
Unit Name
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site 106
0.45
0.02
0.10
0.32
1.00
0.46
1.00
1.00
0.10
EPRI Site 109
0.01
0.39
1.00
1.00
0.26
0.80
0.04
0.79
0.02
EMF (Geometric
mean)
0.08
0.02
0.20
0.57
1.00
0.35
0.89
0.04
0.89
0.04
Geometric standard
deviation
11.80
N/A
2.61
2.23
1.00
1.50
1.17
N/A
1.18
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
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI
Site 13
0.64
0.47
0.19
0.62
0.08
1.00
1.00
0.71
0.58
EPRI
Site 118
0.14
1.00
0.78
0.29
0.57
1.00
1.00
0.64
0.46
EPRI
Site 117
1.00
1.00
1.00
1.00
0.98
0.97
1.00
1.00
1.00
1.00
1990 EMF
(Geometric
mean)
0.44
1.00
0.69
0.53
0.56
0.35
1.00
1.00
0.77
0.64
1990
Geometric
standard
deviation
2.83
1.00
1.71
2.42
1.83
3.73
1.00
1.00
1.26
1.50
EPRI
Site 113
0.21
1.00
1.00
0.40
0.37
0.93
1.00
0.40
1.00
1994 EMF
(Geometric
mean)
0.37
1.00
0.78
0.62
0.52
0.36
0.98
1.00
0.65
0.72
1994
Geometric
standard
deviation
1.95
1.00
1.27
1.68
1.58
2.18
1.02
1.00
1.32
1.30
C-7
-------
Table C-13. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Tangentially-fired, Dry-bottom Furnace Without
Nox Control (Oil-fired)
Unit Name
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site 112
1.00
0.79
0.67
0.66
0.38
0.26
0.80
1.00
0.53
1.00
EMF (Geometric mean)
1.00
0.79
0.67
0.66
0.38
0.26
0.80
1.00
0.53
1.00
Geometric standard deviation
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
Table C-14. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Tangentially-fired, Dry-bottom Furnace with
Nox Control (Oil-fired)
Unit Name
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
EPRI Site 119
1.00
0.57
See Note 1
0.44
1.00
0.69
0.62
0.56
0.35
1.00
1.00
0.72
0.64
EMF (Geometric mean)
0.44
1.00
0.69
0.79
0.56
0.35
1.00
1.00
0.64
0.64
Geometric standard deviation
N/A
N/A
N/A
1.40
N/A
N/A
N/A
N/A
1.17
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 NOX
control burning oil.
C-8
-------
Table C-15. Tested EMFs and Geometric Means Used in the Emission
Factor Program for Fabric Filters (baghouses)
Unit Name
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
Hydrogen Chloride
Hydrogen Fluoride
EPRI Site
10
0.001
0.004
0.004
0.002
0.018
0.002
NSP-
Riverside
#6-7
0.03
0.03
0.06
1.00
0.05
0.03
0.05
1.00
0.20
0.06
EPRI Site
115
0.03
0.003
0.05
0.013
0.007
0.007
0.005
0.27
0.05
0.02
DOE - Niles
#2 W/NOX
Control
0.005
0.004
0.001
0.01
0.003
0.001
0.001
0.005
0.92
0.001
0.79
DOE-
Boswell
0.06
0.009
0.01
0.06
0.004
0.006
0.01
0.003
0.39
0.007
0.31
Motel
Notel
EMF
(Geometric
mean)
0.02
0.01
0.01
0.08
0.01
0.004
0.01
0.01
0.56
0.01
0.12
0.56
1.00
Geometric
standard
deviation
1.66
2.20
2.51
3.47
1.84
1.36
2.70
1.66
1.17
11.26
3.09
Note 1 - These EMFs were developed from emission tests that examined HCI 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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
Hydrogen Chloride
Hydrogen Fluoride
EPRI Site 110
0.11
0.01
0.01
0.004
0.02
0.04
0.02
0.04
1.00
0.002
1.00
EPRI Site 1 10 w/NOx
Control
0.02
0.15
0.01
0.01
0.04
0.02
0.03
0.02
1.00
0.01
0.87
Note 2
Note 2
EMF (Geometric
mean)
0.04
0.04
0.01
0.01
0.03
0.03
0.02
0.02
1.00
0.004
0.93
1.00
1.00
Geometric standard
deviation
3.87
7.13
1.08
2.39
1.84
1.55
1.59
1.86
1.00
3.24
1.10
Note 2 - Because there were no data on HCI 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 HCI and HF emissions passed through the ESP.
C-9
-------
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
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
Hydrogen Chloride
Hydrogen Fluoride
EPRI Site 112
0.49
0.23
0.69
0.44
0.25
0.47
1.00
0.17
0.27
EPRI Site 118
0.55
0.10
0.44
0.08
0.43
0.83
0.58
0.07
0.648
Note 2
Note 2
EMF (Geometric
mean)
0.52
0.16
0.69
0.44
0.14
0.45
0.91
0.31
0.14
0.65
1.00
1.00
Geometric standard
deviation
1.09
1.76
N/A
1.00
2.27
1.07
1.14
2.39
2.50
N/A
Note 2 - Because there were no data on HCI 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 HCI 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
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Lead
Manganese
Mercury
Nickel
Selenium
Hydrogen Chloride
Hydrogen Fluoride
EPRI Site 125
0.10
0.10
0.02
0.09
0.03
0.02
0.03
0.01
0.96
0.01
1.00
0.06
0.09
EMF (Geometric mean)
0.10
0.10
0.02
0.09
0.03
0.02
0.03
0.01
0.96
0.01
1.00
0.06
0.09
Geometric standard deviation
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
C-10
-------
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Appendix D - Discussion of the Methodology Used to Develop
Nationwide Emission Totals
-------
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-l 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
D-l
-------
USGS coal data
(by State and coal type)
Apply coal cleaning
factor
UDI/EEI plant configuration
information
Trace elements (TE) to boiler
Apply boiler TE emission
modification factor
Apply the PM control TE
emission modification factor
Apply SO cgntrolTE emission
modification factor
kg/yr of specific trace element
exiting unit stack
USGS coal data
(by State and coal type)
No cleaning factor
What is
paniculate
matter (PM) control
type?'
Figure D-1. Trace elements in coal.
'Taken from UDI/EEI data.
D-2
-------
UDI/EEI plant configuration
information
Oil
Natural gas
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)
Used a denisty of 8.2 Ib/gal for
feed rate calculation
Trace elements (TE) to boiler
Apply boiler TE emission
modification factor
What is
particulate
patter (PM) control^
type?'
Apply the PM control TE
emission modification factor
What is
the SOzcontrol
type?'
Apply S02 control TE emission
modification factor
kg/yr of specific trace element
exiting unit stack
Figure D-2. Trace elements in oil and natural gas.
'Taken from UDI/EEI data.
D-3
-------
Obtain unifs fuel
consumption
Obtain unifs fuel
consumption
For individual HAPs,
find the median
Ib/trillion Btu emission
factor for a specific
HAP1
Obtain unit's fuel
consumption
For individual HAPs,
find the median
Ib/trillion Btu emission
factor for a specific
HAP2
For individual HAPs, find
the geometric mean
kg/10 * cu ft emission
factor for a specific
HAP3
Individual fuel
consumption x
emission factor x heat
content of 150,000
Btu/gal
Individual fuel
consumption x
emission factor
Individual fuel
consumption x
emission factor x
higher heating value
for bituminous coal
(1Z688 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)
Convert into kg/yr
stack emission for
HAP
'Only oi/*ed units were used to obtain these emission factors.
'Only cod-firsd units were used to obtain these emission factors,
"My gas-fired units were used to obtain tee emission factors.
Figure D-3. Organic emissions.
D-4
-------
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.1 Four units burning
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.2 These equations involved plotting nameplate megawatts
(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 (S02) . Early in the program's development,
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.4 The State of origin used in the EFP was the State
that contributed the highest percentage of the unit's coal. Coal
D-5
-------
consumption by State for each utility was found in volumes of Cost and
Quality of Fuels for Electric Utility Plants for 19905 and 1994.6
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 S02 control types could be sampled.
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 (NOX) control. Since only one type of gas-fired boiler
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 NOX control) unit contained information on two
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 NOX control) for which more data were
collected. The EMFs of the oil burning, front-fired unit (with NOX
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-6
-------
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 moire 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 S02
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 (0-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 NOX control, it was assumed that
the unit had no NOX control and the unit was assigned TANGDRYNONOX
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-7
-------
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-l 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 S02 and PM control. Particulate
matter scrubbers use water only, while flue gas clesulf irization 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 S02 emissions were classified
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 b\it 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-8
-------
Table D-l. Boiler Substitutions Used in the Emission Factor
Program (EFP)
Facility Boiler
CFBDRYNONOX
CFBDRYNOX
COMCYCLNONOX
COMCYCLNOX
COMCYCLNOX
CYCLWETNONOX
CYCLWETNONOX
CYCLWETNOX
FRONTDRYNONOX
FRONTDRYNONOX
FRONTDRYNOX
FRONTDRYNOX
FRONTWETNONOX
FRONTWETNONOX
OPPODRYNONOX
OPPODRYNONOX
OPPODRYNOX
QPPODRYNOX
OPPOWETNONOX
OPPOWETNONOX
OPPOWETNOX
OPPOWETNOX
REARDRYNONOX
REARDRYNONOX
STOKDRYNONOX
TANGDRYNONOX
TANGDRYNONOX
TANGDRYNOX
TANGDRYNOX
TANGWETNONOX
TANGWETNOX
TANGWETNOX
UNKNOWN
VERTDRYNONOX
VERTWETNONOX
G-CYCLWETNONOX
G-CYCLWETNOX
G-FRONTDRYNONOX
G-FRONTDRYNOX
G-FRONTWETNONOX
G-HORZDRYNONOX
G-OPPODRYNONOX
Boiler Used in EFP
CFBDRYNOX
CFBDRYNOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
O-FRONTDRYNONOX
fi
CYCLWETNONOX
TANGDRYNONOX
CYCLWETNONOX
FRONTDRYNONOX
TANGDRYNONOX
FRONTDRYNONOX
TANGDRYNONOX
CYCLWETNONOX
TANGDRYNONOX
TANGDRYNOX
OPPODRYNOX
TANGDRYNOX
OPPODRYNOX
TANGDRYNONOX
OPPOWETNONOX
TANGDRYNONOX
OPPOWETNONOX
FRONTDRYNONOX
TANGDRYNONOX
CYCLWETNONOX
TANGDRYNONOX
TANGDRYNONOX
TANGDRYNOX
TANGDRYNOX
CYCLWETNONOX
TANGDRYNONOX
OPPOWETNONOX
TANGDRYNONOX
VERTDRYNOX
CYCLWETNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
Associated Control Device
BAGHOUSE
BAGHOUSE
NO CONTROL
NO CONTROL
NO CONTROL
ESP, CS
ESP, HS
BAGHOUSE
BAGHOUSE
ESP, HS
BAGHOUSE
ESP, HS
BAGHOUSE
ESP, HS
ESP, HS
BAGHOUSE
ESP, HS
BAGHOUSE
ESP, HS
ESP, CS
ESP, HS
BAGHOUSE
ESP, CS
ESP, HS
ESP, CS
ESP, HS
BAGHOUSE .
ESP, HS
BAGHOUSE
ESP, CS
ESP, HS
BAGHOUSE
ESP, CS
BAGHOUSE
ESP, CS
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
(continued)
D-9
-------
Table D-l. (Continued)
Facility Boiler
G-OPPODRYNOX
G-REARDRYNONOX
G-REARDRYNOX
G-REARDRYNOX
G-TANGDRYNONOX
G-TANGDRYNOX
G-TANGWETNONOX
G-TANGWETNONOX
G-UNKNOWN
G-UNKNOWNDRYNONOX
G-VERTDRYNONOX
G-VERTWETNONOX
O-CYCLDRYNONOX
O-CYCLEDRYNOX
O-CYCLWETNONOX
O-CYCLWETNOX
O-FRONTDRYNONOX
O-FRONTDRYNOX
O-FRONTWETNONOX
O-FRONTWETNOX
O-OPPODRYNONOX
O-OPPODRYNOX
O-OPPOWETNONOX
0-REARDRYNONOX
O-REARDRYNOX
O-TANGDRYNONOX
O-TANGDRYNOX
O-TANGWETNONOX
O-UNKNOWNDRYNONOX
O-VERTDRYNONOX
Boiler Used in EFP
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
G-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNOX
O-FRONTDRYNONOX
O-FRONTDRYNOX
O-OPPODRYNONOX
O-FRONTDRYNOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNOX
O-TANGDRYNONOX
O-TANGDRYNOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
O-FRONTDRYNONOX
Associated Control Device
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
O-ESP, CS
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
O-ESP, CS
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
O-ESP, CS
NO CONTROL
NO CONTROL
NO CONTROL
NO CONTROL
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
G- = gas
O- = oil
CFB boilers are coal fired
Boiler tiring type:
CFB Circulating fluidized bed
COMCYC Combined cycle
CYC Cyclone
FRONT Front
HORZ Horizontal
OPPO Opposed
REAR Rear
TANG Tangential
UNKNOWN Firing type not specified in Utility Data Institute
database (UDI)
VERT Vertical
Boiler bottom type:
DRY
WET
Dry ash
Molten ash
Nitrogen oxides control:
NONOX No nitrogen oxides control
NOX
Control device:
BAGHOUSE
ESP, CS
ESP, HS
O-ESP, CS
NO CONTROL
Nitrogen oxides control by any of several
means as specified in the UDI database
Fabric filter
Cold-side electrostatic precipitator (after trie air
preheater)
Hot-side electrostatic precipitator (before the air
preheater)
Cold-side electrstatic precipitator applied to an
oil-fired boiler
Particulate controls not applied to this boiler
(usually gas or oil fired)
D-10
-------
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 cleaned8 to meet customer
specifications on heat, ash, and sulfur content, a coal cleaning
factor was applied to most bituminous coals in the EFP.9 Two
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,10 and the
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.11 The two sets of
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,13 and from the Illinois State Geological Survey (ISGS) on 34
samples of Illinois cleaned coal.14 Arithmetic averages of the trace
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.15 In summary, USGS data
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-ll
-------
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 Ib/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/109 cubic feet was then multiplied by the unit's gas consumption to
obtain a kilogram HAP/year stack emission rate of each specific HAP
(see Equation 4 in Table D-l). This result was equivalent to a stack
emission because there were no PM control or SO2 control devices on
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-l) 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 (HC1) 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.16 For example, for each Ib/hr of chloride in the feed coal at
D-12
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D-13
-------
Table D-3. Fuel Consumption by Type of Boiler
Facility Boiler
CFBDRYNONOX
CFBDRYNOX
COMCYCLNONOX
CYCLWETNONOX
CYCLWETNOX
FRONTDRYNONOX
FRONTDRYNOX
FRONTWETNONOX
OPPODRYNONOX
OPPODRYNOX
OPPOWETNONOX
OPPOWETNOX
REARDRYNONOX
STOKDRYNONOX
TANGDRYNONOX
TANQDRYNOX
TANGWETNONOX
TANGWETNOX
UNKNOWN
VERTDRYNONOX
VERTWETNONOX
G-CYCLWETNONOX
G-CYCLWETNOX
G-FRONTDRYNONOX
G-FRONTDRYNOX
G-FRONTWETNONOX
G-HORZDRYNONOX
G-OPPODRYNONOX
G-OPPODRYNOX
G-REARDRYNONOX
G-REARDRYNOX
G-TANGDRYNONOX
G-TANGDRYNOX
G-TANGWETNONOX
G-UNKNOWNDRYNONOX
G- VERTDRYNONOX
O-CYCLDRYNONOX
0-CYCLWETNONOX
O-CYCLWETNOX
O-FRONTDRYNONOX
O-FRONTDRYNOX
O-FRONTWETNONOX
Fuel Consumed*
1,157
1885
166,767
56,723
3,098
5,713
59,686
10,045
121,708
138,237
6,217
29,590
733
731
217,266
138,085
1,546
16,606
93
5,008
792
10,245
789
278,662
269,047
4,262
190
409,885
662,119
758
29,747
416,979
221,779
6,287
1,337
139
898
228
792
2,772
35,929
182
Primary Fuel
Coal
Coal
Gas
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Coal
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Gas
Oil
Oil
Oil
Oil
Oil
Oil
D-14
(continued)
-------
Table D-3. (Continued)
Facility Boiler
O-FRONTWETNOX
O-OPPODRYNONOX
O-OPPODRYNOX
O-OPPOWETNONOX
O-REARDRYNONOX
O-REARDRYNOX
O-TANGDRYNONOX
O-TANGDRYNOX
O-TANGWETNONOX
Fuel Consumed*
134
7,967
6,589
766
1,407
1,287
32,654
7,304
9.226
Primary Fuel
Oil
Oil
Oil
Oil
Oil
Oil
Oil
Oil
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
include gas and oil used as starting or temporary fuels in boilers that normally bum other fuels as the primary fuel.
Note: Nationwide total fuel consumption from these boilers is, in the units given above:
Coal 815,135 (18.0x1015Btu)
Bituminous 405,013 (10.3 x 1015 Btu)
Subbituminous 330,978 (6.6 x 1015 Btu)
Lignite 79,t28 (1.1 x 1015 Btu)
Gas 2,708,342 (2.84 x 101S Btu)
Oil 146,148 (0.92 X10'5 Btu)
one of the test sites, 0.63 Ib/hr of HC1 was found in the gas stream
leaving the boiler. Similarly for HF, the boiler emissions were 0.64
Ib/hr for each Ib/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 S02 control. For the 1990 emission estimates, before
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-15
-------
Table D-4. Assigned Chloride ppmw and HCl ppmw Concentrations in
Coal, by State of Coal Origin —
State
Alaska
Illinois
Indiana
Iowa
Missouri
Utah
Washington
Conversion of assigned ppmw chloride to
assigned HCl ppmw
54 x 0.63 =
1,136x0.63 =
1,033x0.63 =
1,498x0.63 =
1,701 x0.63 =
220 x 0.63 =
104x0.63 =
Assigned ppmw HCl in
coal
34.0
715.7
650.8
943.7
1,071.6
138.6
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.18 These EMFs were averaged by
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 oil-
fired 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-16
-------
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
participate control device (when applicable), and the S02 control
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-l (Appendix C) shows two sets of EMF data for the DOE
Niles test site. One unit with NOX control is in a section designated
without NOX control. This apparent anomaly occurs because the NOX
control method used, SCR, is a postcombustion NOX control and does not
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 HC1 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 S02 control.
Hydrochloric acid and HF EMFs for PM and SO2 control devices were
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
SO2 control device. Using chlorine as an example, the measured amount
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
SO2 control device based on the measured mass of HCl or HF entering
that device (leaving the PM control device) and the mass measured at
the exit of the S02 control device. However, a modification was
required to account for flue-gas bypass around the S02 control device.
A portion of the flue gas is bypassed to maintain S02 removal at the
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-17
-------
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 HC1 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 assxuned 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 HC1 and HF EMFs.19 Future wet FGDs are not
expected to use flue-gas bypass in normal operation.
D.ll 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 (Ib/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.
D-18
-------
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Table D-6. Average Higher Heating Values of Coal21
Class and group"
Agglomerating
character
IS8^$Svv^©^3r,"
1 . Low-volatile bituminous coal
2. Medium-volatile bituminous
coal
3. High-volatile A bituminous coal
4. High-volatile B bituminous coal
5. High-volatile C bituminous coal
High-volatile C bituminous coal
commonly
agglomerating0
"
"
-
•
agglomerating
Fixed carbon limits, %
(dry, mineral-matter-
free basis)
Equal or
greater
than
78
69
...
...
...
...
Less than
-•*"'" ; -V
86
78
69
...
...
...
Volatile matter limits,
% (dry, mineral-matter-
free basis)
Equal or
greater
than
•>-' ' ' - >x<-s -s "' /
14
22
31
...
...
...
Less than
^ '*- \ i ••' ^
22
31
_
...
...
...
Calorific value limits,
Btu/lb (moist,6
mineral-matter-free
basis)
Equal or
greater
than
ffl:^
—
...
1 4,000"
1 3,000"
11,500
10,500
Less than
S;'5%>4
...
...
...
14,000
13,000
11,500
Average of Averages (Value used in EFP for bituminous coal)
\ -"'" '-,*-- 1-:"*'* ,,v'v<.: ; ,„'."*;"" /"v „-
li. SubbiUinvinouc •>• -, ^ ,-. ,. , • ; -, .
1. Subbituminous A Coal
2. Subbituminous B Coal
3. Subbituminous C Coal
nonagglomerating
"
•
-- = . '
...
...
...
...
...
...
...
...
...
....
...
10,500
9,500
8,300
11,500
10,500
9,500
Average of Averages (Value used in EFP for Subbituminous coal)
Average
|ff VI
14,000
13,500
12,250
11,000
12,688
11,000
10,000
8,900
9,967
migrttte '•'••'-' ,:.. ": .'.-';:' ; .; ;•-, . " :
1. Lignite A
2. Lignite B
nonagglomerating
-
...
...
...
...
...
...
—
...
6,300
...
8,300
6,300
Average of Averages (Value used in EFP for lignite coal)
7,300
6,300
6,800
' This classification does not include a few coals, principally nonbanded varieties, which have unusual physical and chemical properties and which
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.
° It is recognized that there may be nonagglomerating varieties in these groups of the bituminous class, and there are notable exceptions in high-
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
value.
D-20
-------
D.I2 TREATMENT OP 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 8.7ND
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.I3 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) .22 Then, instead of using an
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.
D-21
-------
Table D-7. Coal Consumption Scaling Factors for 2010
State of coal origin
Kentucky
Pennsylvania
West Virginia
Maryland
Ohio
Alabama
Louisiana
Texas
Virginia
Illinois
Indiana
Iowa
Kansas
Missouri
Oklahoma
Alaska
Arizona
Colorado
Montana
New Mexico
North Dakota
Utah
Washington
Wyoming
201 Of actor"
1.27
1.23
1.24
0.872
0.872
1.41
1.41
1.41
1.41
1
1
1
1
1
1
1.599
1.599
1.599
1.599
1.599
1.599
1.599
1.599
1.599
For each coal-fired unit, the 2010 coal consumption was determined as follows: The 1990 coal
consumption was multiplied by the 2010 factor that corresponded to the State of coal origin assigned to the
unit.
D-22
-------
Table D-8a. Trace Element Concentrations in Coal
State Coal type Compound
AK Subbituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM .
AL Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
AR Lignite ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
1.90
3.00
0.50
0.15
20.00
5.00
53.93
95.00
,5.40
88.00
0.07
10.00
1.60
1.82.
53.00
1.88
0.06
22.80
8.20
380.00
127.00
7.00
41.00
0.19
17.50
1.88
1.17
4.30
2.40
0.29
16.90
6.00
142.00
63.00
9.80
119.00
0.25
11.80
5.00
(continued)
D-23
-------
Table D-8a. (Continued)
State Coal type
AZ Subbituminous
CO Bituminous
CO Subbituminous
Compound
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
0.47
2.10
1.10
0.10
4.60
2.10
200.00
79.00
9.00
27.00
0.07
4.80
1.50
0.91
1.34
0.36
0.18
1.89
1.03
92.97
98.78
5.44
10.83
0.07
1.25
0.87
0.35
1.03
0.84
0.08
4.10
1.60
118.00
99.00
3.50
32.00
0.14
7.90
0.89
(continued)
D-24
-------
Table D-8a. (Continued)
State Coal type Compound
IA Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
IL Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
IN Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration. DDITIW
2.30
12.00
1.88
14.00
12.10
10.00
1498.36
77.00
68.00
259.00
0.19
31.00
3.60
0.82
6.78
1.31
0.98
12.66
3.19
1136.07
84.14
24.51
33.74
0.08
12.74
1.72
1.40
10.10
2.82
0.49
15.40
5.20
1032.79
65.00
10.90
38.00
0.11
17.90
2.17
(continued)
D-25
-------
Table D-8a. (Continued)
State Coal type
KS Bituminous
KY Bituminous
LA Lignite
Compound
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
0.85
25.00
1.47
10.00
10.10
15.00
2500.00
64.00
11 1 .00
160.00
0.19
41.00
2.70
1.13
19.10
3.17
0.16
16.30
6.60
1139.00
86.00
10.60
32.00
0.15
17.50
3.83
0.82
3.70
1.90
0.15
11.40
3.30
115.00
83.00
5.50
141.00
0.19
7.80
6.00
(continued)
D-26
-------
Table D-8a. (Continued)
State Coal type Compound
MD Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
MO Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
MT Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
0.81
26.00
2.01
0.14
26.70
11.00
914.00
107.00
10.00
13.00
0.42
22.00
3.80
1.60
10.00
2.01
0.80
12.20
6.70
1701.64
60.00
67.00
99.00
0.17
23.00
4.20
0.69
7.00
0.52
0.08
3.10
1.50
80.00
104.00
3.00
37.00
0.09
3.90
0.70
(continued)
D-27
-------
Table D-8a. (Continued)
State Coal type
MT Lignite
MT Subbituminous
ND Lignite
Compound
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
0.92
18.00
1.04
0.11
0.94
0.80
67.00
159.00
4.80
68.00
0.12
4.00
0.72
0.69
7.00
0.52
0.08
3.10
1.50
80.00
104.00
3.00
37.00
0.09
3.90
0.70
0.58
8.40
0.82
0.11
7.00
2.70
110.00
34.00
3.73
86.00
0.13
4.10
0.79
(continued)
D-28
-------
Table D-8a. (Continued)
State Coal type Compound
NM Subbituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
OH Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
OK Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
1.07
1.80
2.70
0.16
6.00
2.65
95.00
87.00
31.00
45.00
0.06
4.60
1.94
0.81
23.20
2.39
0.12
14.30
0.90
719.00
92.00
7.30
28.30
0.22
14.90
3.80
0.69
24.00
0.86
0.10
15.00
6.20
267.00
77.00
10.00
74.00
0.17
17.00
1.80
(continued)
D-29
-------
Table D-8a. (Continued)
State Coal type
PA Bituminous
TX Lignite
UT Bituminous
Compound
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
1.23
32.10
2.45
0.10
20.10
7.90
1096.00
78.00
10.80
23.50
0.29
20.40
3.55
0.82
3.70
1.90
0.15
11.40
3.30
115.00
83.00
5.50
141.00
0.19
7.80
6.00
0.23
0.89
0.61
0.08
7.70
2.70
219.67
57.00
3.90
8.00
0.04
4.10
2.00
(continued)
D-30
-------
Table D-8a. (Continued)
State Coal type Compound
VA Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
WA Subbituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
WV Bituminous ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
0.93
11.00
1.66
0.05
12.50
6.30
930.00
74.00
5.80
19.00
0.14
11.20
2.70
0.30
1.50
1.10
0.11
0.70
4.70
103.28
14.00
2.80
41.00
0.06
7.90
0.40
0.93
10.60
2.78
0.10
15.30
7.20
1216.00
58.00
7.20
19.10
0.16
14.20
3.97
(continued)
D-31
-------
State Coal type
WY Subbituminous
Compound
ANTIMONY
ARSENIC
BERYLLIUM
CADMIUM
CHROMIUM
COBALT
CHLORINE
FLUORINE
LEAD
MANGANESE
MERCURY
NICKEL
SELENIUM
Concentration, ppmw
0.73
0.69
0.18
0.13
2.82
0.87
118.30
43.70
2.07
5.65
0.08
2.17
0.51
Table D-8b.
estimates)
Trace Element Concentrations in Fuel Oil (for 1994
Trace Element
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Chlorine
Fluorine
Lead
Manganese
Mercury
Nickel
Selenium
Concentration in Oil, ppmw
0.306
0.027
0.020
0.31
1.63
131
17.5
1.41
0.35
0.0092
26
0.095
Table D-8c. Trace Element Concentrations in Gas
Trace Element
Arsenic
Cobalt
Lead
Mercury
Nickel
Concentration in gas mg/m3
0.000963
0.100
0.100
0.0000024
0.0500
D-32
-------
Table D-9. Coal Cleaning Factors for Bituminous Coals Used in
the Emission Factors Program*
Constituent
Antimony
Arsenic
Beryllium
Cadmium
Chromium
Cobalt
Chlorine
Fluorine
Lead
Manganese
Mercury
Nickel
Selenium
Cleaning factor
0.715
0.554
0.711
0.624
0.512
0.537
0.496
0.496
0.449
0.382
0.790
0.568
0.745
a Applying the cleaning factors to United States Geographical Survey (USGS) constituent concentrations for
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-33
-------
D.I4 REFERENCES
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-34
-------
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.
D-35
-------
Appendix E - Health Effects Summaries: Overview
-------
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 System1 or from an EPA document
titled Health Effects Notebook for Hazardous Air Pollutants.2 Each
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 weight-
of-evidence classification system is included. EPA uses a weight-of-
evidence, 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-l
-------
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 structure-
activity 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
A
B1
B2
C
D
E
Description
Known human carcinogen
Probable human carcinogen, limited human data
Probable Human carcinogen, sufficient evidence
inadequate or no evidence in humans
are available
in animals and
Possible human carcinogen
Not classifiable as to human carcinogenicity
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 (HIRE) 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 (yug/m3) of air for 70 years. The IURE is derived using
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-a-
million 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-
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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 /ug of pollutant per liter.
It is expressed in units of
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 (LD50) or median lethal concentration
(LC50). This is the dose (or concentration) estimated to kill 50
percent of the experimental animals. Results from these tests are
divided into the following toxicity categories:
Lethality
Oral LD50
Dermal LD,n
Inhalation LC50
Extreme
< 50 mg/kg
<200 mg/kg
<200 mg/m3
High
50 to 500 mg/kg
200 to 2,000 mg/kg
200 to 2,000 mg/m3
Moderate
500 to 5,000 mg/kg
2,000 to 20,000 mg/kg
2,000 to 20,000 mg/m3
Low
>5,000 mg/kg
>20,000 mg/kg
»20,000 mg/m3
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
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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/m3) . If available, the Oral Reference Dose (RfD) is
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 short-
term 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
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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"3 to 7.6 x 10"
3 (pig/m3)'1. Using the geometric mean of these data, EPA calculated an
inhalation unit risk estimate of 4.29 x 10'3 (^g/m3)"1 (EPA) .6 Based on
this unit risk estimate, EPA estimates that if an individual were to
breathe air containing arsenic at 0.0002 ,ug/m3a over his or her entire
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/in3 would result in an increased
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
0.0002 Mg/m3 (concentration corresponding to a 10"6 risk level) =10"6
(risk level)/4.29 x 10'3 (^g/m3)'1 (unit risk estimate).
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risk assessment methodology to recalculate the estimated risk. They
calculated a new unit risk of 1.43 x 10'3 (//g/m3)"1, which is one-third
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,9 they did
propose that a unit risk estimate of 5 x 1CT5 (ug/L)"1 from oral
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 othejr 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 jug/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
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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 LD50 for arsenic to range between 50 and 5,000 mg/kg.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
blackfobt 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 I.16 This RfD was based on two studies17 that
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
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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.27 Since the human studies could not differentiate between
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.28 EPA has classified
chromium III in Group D — Not Classifiable as to Human
Car cinogeni city.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
dug/m3) -1.30 Based upon this unit risk estimate, EPA estimates that if
an individual were to breathe air containing chromium VI at 0.00008
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//g/m3 b over his or her entire lifetime, that person would theoretically
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 yug/m3
would result in an increased chance of up to one in one hundred
thousand of developing cancer.31 EPA has not calculated a risk
estimate from oral exposure to chromium VI32 or from inhalation or oral
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 LC50 for chromium VI to be less than 200 mg/m3 and the LD50
to be less than 50 mg/kg. Chromium III has been shown to have
moderate toxicity from oral exposure, based on LD50 data in the range
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,
0.00008 jwg/m3 (concentration corresponding to a 10's risk level) = 10"6
(risk level)/!.2 x 10"2 (jug/m3)'1 (unit risk estimate).
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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 I.40 This was based on a study of rats,
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
I 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 I.43 This was based on no effects observed in rats
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.44 EPA has not established an RfC for chromium III45 or
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
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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.50 No studies
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/m3 hydrochloric acid for 30 minutes resulted in death from
laryngeal spasm, laryngeal edema, or rapidly developing pulmonary
edema.53 Acute inhalation exposure tests resulted in an LC50 of 1,108
ppm for exposed mice and 3,124 ppm for exposed rats (moderate to high
acute toxicity) . An LD50 of 900 mg/kg (moderate acute toxicity) was
reported for rabbits exposed orally to hydrochloric acid.54 No
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
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occupationally to hydrochloric acid.55 No other data are available
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 sguamous 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.56 No studies are
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?.
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/m3.57 The EPA has low
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.59 No animal studies are available on
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.60 EPA
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.62 No animal studies have been
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
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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
In 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.68 In
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.70 No information was
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
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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-l
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.75 Nickel refinery dust is defined as the "dust
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 exposure79'8°i81'82
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.83 The International Agency for Research on
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.
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Table E-l. Physical Properties of Some Forms of Nickel
Chemical Name
Metallic Nickel
Nickel Hydroxide
Nickel Subsulfide
Nickel Carbonyl
Nickel Sulfate
(anhydrous)
Nickel Monoxide
Nickel Acetate
Formula
Ni
Ni (OH)2
Ni3S2
Ni (0)4
NiS04
NiO
Ni(OCOCH,),
Description
Lustrous white, hard
ferromagnetic metal or grey
powder
Green crystals or
amorphous solid
Lustrous pale yellow or
bronze metallic crystals
Colorless to yellow liquid
Pale-green to yellow crystals
Grey, black, or green
powder
Dull-green crystals
Solubility
Soluble in dilute nitric acids; slightly
soluble in hydrochloric or suit uric acids;
insoluble in cold or hot water
Nearly insoluble in cold water; soluble in
acid, ammonium hydroxide
Insoluble in cold water; soluble in nitric
acid
Nearly insoluble in water; soluble in
ethanol, benzene, and nitric acid;
insoluble in dilute acids or dilute alkali
Soluble in water; insoluble in ethanol
Insoluble in water, soluble in acid
Soluble in water, insoluble in ethanol
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"4 (/^g/m31"1.84 Based upon this unit risk estimate, U.S.
Environmental Protection Agency estimates that if an individual were
to breathe air containing nickel refinery dust at 0.004 ug/m3 over his
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 ug/m3 would result in an increased
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
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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.89 This is based on epidemiological studies that
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 inj ections.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.91 For nickel subsulfide, U.S.
Environmental Protection Agency also used human delta to estimate the
unit cancer risk. U.S. Environmental Protection Agency calculated an
inhalation unit risk estimate of 4.8 x 10~4 (yUg/m3) ~1.92 U.S.
Environmental Protection Agency estimates that if an individual were
to breathe air containing this nickel compound at 0.002 ,ug/m3 over his
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 yug/m3 would result in an increased chance
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.95 The
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 jug Ni/m3. This
risk ranges from 2.8 x 10"3 for the maximum likelihood estimate to
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3.7 x 10"3 for the upper 95 percent confidence limit. This risk
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.98 IARC has classified nickel carbonyl as having limited
evidence in experimental animals for carcinogenicity.99 This is based
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.102 This is based on animal
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.104 This is based on animal
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.105 This is based on animal studies that showed an
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
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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/m3. This risk
ranges from 2.1 x 10"4 for the maximum likelihood estimate to 2.57 x 10"
4 for the upper 95 percent confidence limit. This risk estimate was
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/m3 and to exposure
to less soluble forms at concentrations greater than 10 mg/m3.111
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 LD50 test in rats, have shown nickel
compounds to exhibit acute toxicity values ranging from low to high,
based upon LD50 data in the range of 50 mg/kg to greater than 5,000
mg/kg. The soluble compounds, such as nickel acetate, were most
E-18
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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 ,ug of nickel.119 Other researchers have
estimated requirements ranging from 30 jj,g to 120 y.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 I.122 This was based on a study in rats that showed
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
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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.124 The database provided adequate
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 10'3 mg(Ni)/m3 for all
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.12S
Calabrese has calculated an ambient air level goal (AALG) for
soluble nickel compounds of 0.36 ng (Ni)/m3 and an AALG for insoluble
nickel compounds of 7.1 ng (Ni)/m3. An AALG is a health-based
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)/m3) measured in California near an industrial source, it is
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"5 mg(Ni)/m3. They have stated
that this MRL may not be protective for some hypersensitive
individuals.129 An MRL is a health-based guideline based on similar
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)/m3 for metallic nickel
and insoluble compounds, and 0.1 mg(Ni)/m3 for soluble nickel
compounds. 13°
The National Institute of Occupational Safety and Health has a
recommended exposure level for workplace air of 0.15 mg (Ni)/m3 for all
nickel compounds except nickel carbonyl and 7 jj,g (Ni)/m3 for nickel
carbonyl.131
The ACGIH has recommended a TLV of 0.05 mg(Ni)/m3 for an
8-hour exposure in the workplace to all nickel compounds (elemental,
insoluble, and soluble) .132
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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"4 to 5 x 10"4 pg/kg-day)'1; for all
cancers combined the unit risk estimate is 2 x 10"3 to 3 x 10'3
(pg/kg-day)'1 (U.S. Environmental Protection Agency136).
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.137 Routes of exposure in these
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. LD50 values
range from 0.6 /^g/kg in male guinea pigs to 5,500 /^g/kg in hamsters.
E-21
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Other effects on animals from acute oral exposure include loss of body
weight, hepatotoxicity, and decreased thymus weight.138 Information on
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.142 U.S. Environmental Protection Agency has not
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.144 In oral exposure studies, 2, 3,7, 8-TCDD has produced
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.
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E-24
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77. Ref. 76, Carcin 1 & 2.
78. ATSDR. Toxicological Profile for Nickel. Agency for Toxic
Substances and Disease Registry. U.S. Public Health Service.
Section 2.2. 1993.
79. Ref. 76, Carcin-2.
80. Ref. 78, Section 2.2.
81. U.S. Environmental Protection Agency. Integrated Risk
Information System on Nickel Carbonyl. Carcin-2. 1995.
82. U.S. Environmental Protection Agency. Integrated Risk
Information System on Nickel Subsulfide. Carcin-2. 1995.
E-26
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83. Ref. 76, Carcin-1.
84. Ref. 76, Carein-3.
85. Ref. 76, Carcin-3.
86. Ref. 76, Carcin-3.
87. National Toxicology Program. NTP Draft Technical Report on the
Toxicology and Carcinogenesis Studies of Nickel Sulfate
Hexahydrate in F344/N Rats and B6C3F1 Mice. U.S. Dept. of Health
and Human Services. Public Health Service. National Institute
of Health. 1994.
88. International Committee on Nickel Carcinogenesis in Man. Scand.
J. Work. Environ. Health. 16:1-82. 1990.
89. Ref. 35.
90. Ref. 35.
91. Ref. 82, Carcin-1.
92. Ref. 82, Carcin-2.
93. Ref. 82, Carcin-3.
94. National Toxicology Program. NTP Draft Technical Report on the
Toxicology and Carcinogenesis Studies of Nickel Subsulfide in
F344/N Rats and B6C3F1 Mice. U.S. Dept. of Health and Human
Services. Public Health Service. National Institute of Health.
1994.
95. Ref. 35.
96. Ref 88, p. 71-72.
97. California Air Resources Board. Initial Statement of Reasons for
Rulemaking. Proposed Identification of Nickel as a Toxic Air
Contaminant. p. 4-Part B. 1991.
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 B6C3F1 Mice. U.S. Dept. of Health and Human Services.
Public Health Service. National Institute of Health. 1994.
E-27
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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.
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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. 1-3.
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-Tetrachloro-
dibenzo-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.
E-29
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Appendix F — Documentation of The Inhalation Human Exposure Modeling
for the Utility Study
-------
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-l.)
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 Models1 as a preferred model for dealing
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
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-l
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Table F-l. Summary of HEM 1.5 Features
Characterization
Dispersion model
Meteorological database
Population database
Exposure calculations
Single pollutant, multiple source, nationwide
ISCLT2
Data set from locations/years available on OAQPS TIN and from the
National Weather Service
1990 Census Databases
Block level
6.9 million records
>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
TIN = 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.
F-2
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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:
u
•E
Where:
XA = total ambient concentration at receptor i, ug/m3
q± = emission rate for stack, g/s
Xi:j = normalized contribution from stack j to receptor i', ug/m3
J = total number of stacks.
Thus, the principal output of the dispersion modeling is a set of
normalized stack contributions (i.e., xi;j in the above equation) for
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 Models2 distinguishes between
urban and rural settings based on population density. "Urban" is
defined as a population density greater than 750 people per km2 in the
F-3
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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 km2.
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-4
-------
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-l
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-5
-------
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-------
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.s
F.5 RISK CALCULATIONS
In general, long-term exposure estimates are paired with
chemical-specific health benchmarks, such as inhalation unit risk
estimates (lUREs), 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.
F-7
-------
North
South
A Block Centroid
B&
*• Census Block
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-8
-------
The utility boiler HEM modeling application requires the input of
chemical-specific toxicity information. HEM 1.5 uses the lUREs 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 ug/m3 of air contaminant.
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 EP.A 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-9
-------
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-10
-------
F.8 REFERENCES
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-ll
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Appendix 6 — Data Tables for Dioxin Multipathway Assessment
-------
-------
Appendix 6-1
IBM Model Input Parameters
For Dioxin Screening Level
Multipathway Assessment
(Chapter 11)
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G-39
-------
Appendix G-6
Comparison of Dioxin Local-Scale ISCST3 and Long-Range RELMAP Modeling
Exposures and Risks
G-41
-------
Table G-6. Comparison of Local-Scale ISCST3 and Long-Range
RELMAP Modeling Exposures and Risks for Dioxin
LCH-fisher
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Fish Ingestion
Direct Inhalation
LOH-fisher
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Fish Ingestion
Direct Inhalation
SCH-fisher
Soil Dermal
Soil Ingestion
Vegetation Ingestion
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SOH-fisher
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Fish Ingestion
Direct Inhalation
LCH-farmer
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Animal Product Ingestion
Water Ingestion
Direct Inhalation
Total TEQ exposures
(LADD in mg/kg-d)
2.9e-14
1.6e-12
6.0e-12
3.2e-12
9.96-10
1.86-15
1.66-14
8.46-13
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3.86-11
1.86-15
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1.66-12
1.96-11
2.0e-10
3.2e-12
1.8e-15
Total TEQ
Risk
5e-09
2e-07
9e-07
5e-07
2e-04
3e-10
2e-09
1e-07
5e-07
3e-07
1e-04
3e-10
1e-09
7e-08
3e-07
2e-07
6e-05
3e-10
1e-10
8e-09
7e-08
2e-08
6e-06
3e-10
3e-08
2e-07
3e-06
3e-05
5e-07
3e-10
RELMAP % of Total
(mean data)
0.14%
0.14%
4.92%
0.22%
0.22%
100.00%
0.26%
0.26%
8.50%
0.33%
0.33%
100.00%
0.48%
0.48%
14.64%
0.61%
0.60%
100.00%
4.48%
4.48%
62.78%
5.75%
5.68%
99.97%
0.14%
0.14%
2.48%
0.85%
0.22%
100.00%
G-43
-------
Table G-6. (Continued)
LOH-farmer
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Animal Product Ingestion
Water Ingestion
Direct Inhalation
SCI-Marmer
Soil Dermal
Soil Ingestion
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Water Ingestion
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Soil Ingestion
Vegetation Ingestion
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Direct Inhalation
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Water Ingestion
Direct Inhalation
LOC-resident
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Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
SCC-resident
Soil Dermal
Total TEQ exposures
(LADD in mg/kg-d)
8.7e-14
8.4e-13
1.0e-11
1.4e-10
2.1e-12
1.8e-15
4.7e-14
4.56-13
5.9e-12
7.96-11
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1.8e-15
5.0e-15
4.8e-14
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9.4e-12
1.2e-13
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1.2e-13
2.96-15
5.86-16
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2.8e-15
4.1e-16
Total TEQ
Risk
1e-08
1e-07
2e-06
2e-05
3e-07
3e-10
7e-09
7e-08
9e-07
1e-05
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3e-10
8e-10
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2e-07
1e-06
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3e-10
2e-10
9e-09
4e-08
2e-08
5e-10
9e-11
5e-09
3e-08
1e-08
4e-10
6e-11
RELMAP% of Total
(mean data)
0.26%
0.26%
4.42%
1.24%
0.33%
100.00%
0.48%
0.48%
7.89%
2.20%
0.61%
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4.48%
4.48%
45.66%
18.54%
5.75%
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3.74%
3.74%
37.37%
5.76%
61.32%
7.01%
7.01%
49.78%
8.80%
63.60%
9.85%
G-44
-------
Table G-6. (Continued)
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
SOC-resident
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
LCC-child
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
LOC-child
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
SCO-child
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
SOC-child
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation
Total TEQ exposures
(LADD in mg/kg-d)
2.2e-14
2.2e-13
5.6e-14
4.1e-15
1.06-16
5.56-15
1.16-13
1.56-14
2.26-15
2.8e-15
9.56-14
8.16-14
4.96-14
1.96-15
1.56-15
5.16-14
6.36-14
3.26-14
1.8e-15
1.16-15
3.6e-14
8.3e-14
2.36-14
2.76-15
2.7e-16
9.0e-15
4.46-14
6.16-15
1.5e-15
Total TEQ
Risk
3e-09
3e-08
9e-09
6e-10
2e-11
9e-10
2e-08
2e-09
3e-10
4e-10
1e-08
1e-08
8e-09
3e-10
2e-10
8e-09
1e-08
5e-09
3e-10
2e-10
6e-09
1e-08
4e-09
4e-10
4e-11
1e-09
7e-09
1e-09
2e-10
RELMAP% of Total
(mean data)
9.85%
39.83%
12.19%
43.52%
39.65%
39.65%
77.27%
46.34%
79.02%
3.74%
3.74%
43.34%
5.76%
61.32%
7.01%
7.01%
55.25%
8.80%
63.60%
9.85%
9.85%
42.00%
12.19%
43.52%
39.65%
r39.65%
78.82%
46.34%
79.02%
G-45
-------
G.I 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 Dioxin-
Like Compounds, Vol. Ill: Site-Specific Assessment Procedures.
External Review Draft. EPA/6QO/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/R-
96-OOlc. Office of Air Quality Planning & Standards and Office of
Research and Development. 1996c.
G-46
-------
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.
G-47
-------
Appendix H - Literature Review of the Potential Impacts of Hydrogen
Chloride and Hydrogen Fluoride
-------
H.I OVERVIEW
The information presented in this appendix was collected to
expand the EPA's knowledge of the potential impacts of HC1 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.2 Puxbaum, et al. cite coal combustion as
the primary source of HCl in central Europe.3 Other sources include
biomass burning and the photolysis of Cl-atom precursors such as HCl
and C12, followed by hydrocarbon reactions.4 Natural sources of HCl
emissions include volcanic activity, marine plants/microorganisms,
H-l
-------
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 1977-1990.6 They predicted HCl
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 HN03 and/or H2S04. Total global
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 SOX (83 percent anthropogenic), 2.2 Teq H+
per year for NOX (57 percent anthropogenic), and 1.6 Teq H* per year
for HCl. However, because most of the HCl is thought to be generated
by acid-displacement reactions involving anthropogeaiically 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 SOX and NOX. Because SOX and NOX
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,
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
S02 and NOX, respectively. When evaluated at the scale of Western
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
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-2
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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,7 ranging from none detected to 4
Wegner, et al . found tropospheric HCl concentrations to vary -
substantially as a function of location and time, and to be strongly
correlated with CH20 concentrations ® = 0.93, 0.90, and 0.95 for polar,
midlatitude maritime, and midlatitude continental, respectively) .-
Average HCl concentrations measured in a tropospheric column were 1 . 15
x 1015 molecules per square centimeter. The concentrations of HCl
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/m3) were believed to originate
largely from an incinerator located approximately 3 km away from the
monitor. - Hutchinson, et al . , in a paper on HCl-induced stone
degradation, estimate typical atmospheric HCl concentrations ranges in
North America of 0.1-1.4 /ug/m3 in rural areas and 0.2-3 ^g/m3 in urban
areas . 8 Puxbaum, et al . measured HCl and several other chemical
species at a rural site in northeast Austria.- HCl concentrations
exhibited substantial variation between winter and summer, with
elevated values of 0.7 /ug/m3 found in winter. Annual average HCl
concentrations were 0.3
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 PM2.5 exceedances . The
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.9 A study conducted by
the U.S. Bureau of Mines indicated that the majority of chlorine
contained in coal volatilizes to form HCl.10 Additional tests found
that only small amounts of chlorine remain in the combustion ash.
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Other studies reveal the processes of chlorine and HC1 formation in
anthropogenic and natural systems.
Nonanthropogenic HC1 is emitted by volcanoes, or it can be formed
from deliquescent sea salt in the marine environment by the following
process:-
HN03(g) + NaCl,., - HCl(g, + NaN03(s) (1)
H2S04(S) + 2NaCl(s) - 2HCl(B) + Na2S04(s) (2)
Thus, sulfuric acid and nitric acid in the atmosphere can react with
sea salt spray to generate hydrogen chloride. HC1 in marine
environments can also be formed by indirect pathways, which generate
various chlorinated species such as ClN02 or C12. Once this happens,
photolysis of the chlorinated species will produce chlorine radicals,
which react as follows to produce HC1:
Cl« + RH - HC1 + R» (3)
Rupert and Sigg investigated the interaction between fogwater and
aerosols, which can create or destroy HCl.- The following reaction was
found to account for the transition of chlorine between the aerosol
and gaseous phase:
NH4Claerosol - NH3(g) + HCl(g) (4)
In their study of fogwater and aerosols in an urban area of
Switzerland, Ruprecht and Sigg found empirical values of NH4C1 to be
lower than theoretical values. One possible explanation is that
insufficient amounts of NH4C1 were available to sustain equilibrium
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 NH3 must have been too low to neutralize the HCl present,
hindering the formation of NH4C1. A significant portion of atmospheric
Cl" is present as fine aerosol (<2.4 ^on) 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
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-4
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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
C12, HOC1, and NOC1.— Analysis during laboratory simulation did not,
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: NaNO3, Na2SO4,
hydrocarbon radicals, and NH3.
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
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.-
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 PM2.5-sensitive
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.-'-
HCl's dry deposition rate is thought to be controlled by atmospheric
turbulence, rather than surface conditions. In general, HCl gas will
H-5
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be removed from the atmosphere much faster than S02 or NO2 and will be
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 S02
suspended in chloride-containing water droplets to be oxidized more
rapidly than if in pure water. If this is the case, SO2 deposition
close to the source will be decreased, but rain and cloudwater acidity
will be increased, subject to further transport. However, because the
solubility of S02 is pH dependent, if HCl lowers the pH below
approximately 3, chloride ion catalysis may be reduced. HCl's
solubility is the reason it is more efficiently removed by rain than.
either S02 or NOX. Thus the acidity in rainwater near a coal-fired
power plant may be predominately the result of HCl, rather than S02 or
NOX.
More evidence of chloride ion catalyzation of S02 was found in a
study by Clarke and Radojevic. This study was designed to provide
more applicable kinetic data concerning the oxidation of S02 in fresh
and salt water.13 Clarke and Radojevic found that various chloride
salts significantly increased the rate of S02 oxidation. The effects
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 HC1-
enriched plumes at high humidity. In these environments, rates of
oxidation of S02 in droplets will be elevated above the rates of pure
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 Cl2, HOC1, BrCl, arid ClNO2, which are
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 - H20 + Cl» (5)
HCl can impact the atmospheric chemistry of other species,
including other HAPs. For example, Seigneur, et al. identified the
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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)(a) + HCl(g) - products (6)
Reaction rate parameter = 1.0 x 10"19 cm3 molecule"1 s"1
In their theoretical study, Seigneur, 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 S02 concentration were all found to have a large
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 S02 or N02 and will be deposited in close proximity to the
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
PM2.5.- One way in which HCl may move long distances to sensitive
receptors is via the process of chloride ion catalyzation of SO2 to S04,
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
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.
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HC1 was found to affect the gas/liquid partitioning of Hg(0).^
For instance, when HCl is present, nearly all Hg(II) is present as
HgCl2. However, when HCl is absent, most Hg(II) is present in the form
of Hg(S03)22~ with a small portion present as HgS03. When liquid water
content, pH, and total S02 concentration are kept constant at O.I g/m3
and 4 and 10 ppb, respectively, moving from 0 yug/m3 HCl to 1 /zg/m3 HCl
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 long-
term 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 S02 to S04 and by
producing strong acidity, may have an impact on acid deposition.
Concentrations of gaseous HCl are limited by the equilibrium with
solid aerosol phase NH4C1 (see reaction #4) .- During fog events, high
concentrations of sulfate may be found in small aerosols as a result
of high-pH, aqueous-phase S02 oxidation. The presence of HCl can lower
the pH to the point that S02 oxidation is delayed, possibly altering
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
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.16 Acidic deposition may increase the
concentration of mobile strong-acid anions, thus hastening base cation
leaching in soils with medium to high base saturation, or leaching of
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acid cations (H+, Aln+) into surface waters from low base saturation
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.—
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.—
Vegetation takes up and retains sea-salt-deposited base cations,
producing strong mineral acid (HC1) 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
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-9
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In summary, by altering S02 oxidation, HCl can possibly impact
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.17 The pH, water temperature,
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 chlorine-
containing compounds such as OC1" ultimately releases chloride ion.
Skeffington lists a potential equilibrium constant (pKa) of -3
for the following reaction of hydrochloric acid and water:18
HCl -i- H2O - H3O* + Cl" (7)
A pKa value in this range means that the above reaction goes to
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.19 By combining previous
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".
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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.20 The positive DOC interference was found to bias colorimetric
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.21 Within the study area in southern Germany, it was found
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/km2/yr, and chloride deposition varies
from 670 to 1,170 kg/km2/yr. Although it is qualitative, this work
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
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pollution and land use or atmospheric deposition a potential screening
instrument for groundwater contamination by anthropogenic activities,
such as utility HC1 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 HC1 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 RfCj. for HCl of 20 ,ug/m3 to
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 yug/m3 for HCl. Additional
information on HCl acute toxicity was obtained from a Lab Chemical
Safety Summary compiled by the National Research Council and EPA
Region III.23 The National Research Council reported a Threshold Limit
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-l contains the National Research Council's reported toxicity
values. EPA Region Ill's Superfund Risk Based Concentration (RBC) for
chlorine is 370
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.
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 long-
term exposure may cause tooth erosion. The WHO concluded in a review
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Table H-l. National Research Council Acute Toxicity
Values for Chlorine
Chemical
Hydrogen chloride
Chlorine
TLV-TWA
5ppm
0.5 ppm
PEL
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).u Most nonanthropogenic chloride is
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
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.—
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
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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
treatment as a disinfectant for drinking water and wastewater prior to
discharge; Manning, Wilson, and Chapman found it to be toxic to
aquatic life.26 Many treatment plants use ammonia and chlorine
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
LC50 obtained in these tests, the environmental concern level for
chlorine in marine situations is 0.0018 rag/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
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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
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.28 Relatively small amounts are
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 /ug/m3, which is at the limit of
detection.29 Urban fluoride concentrations ranging from 0.1 jUg/m3 to
1.0 /t/g/m3 have been observed in British industrial cities.— However,
HF accidentally released from chemical plants may produce
concentrations on the order of 50 /ug/m3 for several hours. As of
August 1994, only one HF monitoring station had been operated in the
United States, with 20 samples measured.- Measured concentrations of
HF ranged between 1 /ug/m3 and 8
As in the case of HC1, 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 PM2 5 exceedances .
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
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
alumina with HF adsorbed to its surface, and particle diameters range
from <0.1 yum to approximately 10 /zm.— Davison reports that fugitive
fluoridated alumina particle sizes near a dry-scrubbed aluminum
smelter can range from <1 yum to about 5 /uin in diameter.—
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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)8 at ambient temperatures through
hydrogen bonding. HF can also be present as single molecules at
higher temperatures. HP'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 HN03, H2S04 and HC1,
present in the atmosphere.— Therefore, it is possible that HF may
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 HC1 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.—
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
0.54 days, while fluoride aerosol has a half-life of 2.1 days.—
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.—
Hance, et al. found precipitation to be the major global route of
atmospheric HF removal.— Slooff, et al. also cite wet and dry
deposition as the primary routes of HF removal from the atmosphere.—
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.—
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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 (Ftotal:SOx ratio) was used as an atmospheric
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.32 Slooff, et al. report that atmospheric
fluoride, especially aerosol fluoride, can be dispersed over great
distances by wind or atmospheric turbulence.— Fine particles will
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.33 Background foliar
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
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 18F isotope (1.8 hours) has precluded its use for
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
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a rapid rate.— Some possible mechanisms have been investigated with
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.34 Analyses of
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
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
ecosystem. For instance, Murray found that fire increases the water-
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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.36 The experiments were designed to test the hypothesis that
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 water-
soluble 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
with pH greater than 5, fluoride is mainly present as fluoride ion,
based on stability diagrams derived from thermodynamic calculations.
In fresh water, f luorapatite, Ca(P04)3F or Ca10 (P04) 6F2, is only
somewhat soluble.— Slooff, et al. provide the following equilibrium
reactions:
Ca(P04)3F + 6H* * 5Ca2+ + 3H2P
-------
Ca5(P04)3F + H20 ** Ca5(P04)3OH + 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.—
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/m3. Additional literature reviewed on human
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
contains toxicity values for several fluoride species.
Whitford reviewed fluoride ingestion findings in a number of
studies.38 The difficulty in estimating the lethal dose or potentially
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,
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.
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Table H-2. National Research Council Acute Toxicity
Values for Fluorides
Chemical
Hydrogen fluoride
Fluorine
Fluorides
TLV-TWA
3ppm
1 ppm
2.5 mg/m3
PEL
3 ppm
0.1 ppm
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.
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,
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.40 Insoluble forms, such as calcium fluoride
and cryolite (Na3AlF6), are less toxic than sodium fluoride, as they
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: Ml) 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."—
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 HFVegetation 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.42 All were pioneer tree species known to be ubiquitous in
the study area. The vegetation and soils within a 60 km2 radius of
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
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
salts may form, thereby rendering both cations and anions
physiologically inactive. Reactions are as follows:
HF(B) - H+ + F- (11)
Ca2+ + 2F~ - CaF2 (12)
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.43 Lichen fluoride concentrations were
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.
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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.44 Pine stands within 0.8 km of the smelter had foliar
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.45 For
instance, a review conducted by Davison noted that ripening peach
fruit tissues can be visibly injured by fluoride concentrations of 0.3
£ig/m3 for 12 hours, 5 days a week.— However, most plant species are
considered resistant to fluoride and can tolerate concentrations of up
to 30 /ug/m3.
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.—
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
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, NOX, ammonia, and SOX, are likely to be
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
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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.46 Fluoride
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 Ca2+ and A13+. The
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 (Mis 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.47 The order
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-25
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nisus), kestrel (Falco tinnunculus), barn owl (Tyto alba), and tawny
owl (Strix aluco),48 Fluoride loads increased with age in all four
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.2°
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.—
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
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
nothing was known about fluoride transport to invertebrates or
detritus chain organisms. However, one study on this topic was found
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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.—
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 (Mzcrotus 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
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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 scaJber) and found neither to be adversely affected by soil
fluoride concentrations of up to 170 micro mol per gram.51 A previous
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 eitmospheric
deposition of fluoride.53 High accumulation of substances in isopods
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
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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.54 Fluoride failed to
influence the increased lead concentration in the plasma, femur,
liver, and kidney, as well as the increased excretion of delta-
aminolevulinic 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.55 The authors examined the effects of long-term fluoride
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 (Cd2*) toxicity in rats.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 D3 appeared to have experience reduced Cd2+
hematoxicity. In another study, Yu, et al. examined whether high
dietary fluoride inhibits selenite toxicity in rats.57 The significant
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
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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.58 The review covers microbial, in vitro
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 ,ug/m3 to 0.5 ^g/m3 for a
continuous period of several months. If higher concentrations are
present, more severe fluorosis symptoms will occur more rapidly.—
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.—
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 chemiccil 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 fauna-
mediated 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 yug/m3 to 0.5 /wg/m3 concentrations
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 life-
long 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
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 LC50 values of 92.4, 85.1, 73.4 and 64.1 ppm F" (rainbow trout)
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.60 These five
freshwater benthic macroinvertebrates, all native to the rivers of the
Iberian Peninsula, were found to have 96-hour LC50 values of 44.90,
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 prcraelas) and water flea (Ceriodaphnia duhia).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 LC50 of 2.26 percent, a survival NOEC of 1.25
percent, and a growth NOEC of 0.625 percent. After an 8-day exposure
period, water fleas had an LC50 of 3.11 percent, a survival NOEC of 2.5
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 LC50, a
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.62 The number of
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
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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
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 fluoride-
containing 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 SCI'8 impacts on the spatial and temporal deposition of
SOxr a species implicated in both acid rain and PM2 5?
S02 and HCl interact in the atmosphere. For instance, chloride
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 S02 will be dry-deposited
close to the source, but more H2S04 will be generated in cloud and
rain water, subject to further transport. HCl might, therefore,
alter the spatial and temporal deposition of S02, which may be an
important way in which utility HCl emissions indirectly affect
acid rain and PM2.5.
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, Seigneur, Wrobel, and Constantinou found
that HCl concentration, among other factors (e.g., atmospheric
liquid water content, pH, and S02 concentration) , had a strong
effect on both the Hg(0)/Hg(II) ratio and the gas/liquid
partitioning of Hg(0).i±
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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 SCI and OF particulate is in the fine
fraction?
EPA has promulgated a primary 24-hour PM2.5 standard of 65 fj.g/m3
and a primary annual PM2.S standard of 15 /ug/m3. As part of the
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
jUg/m3. Information on rural HF concentrations was more scarce,
with a reported high end concentration of I yug/m3. However,
accidental industrial releases of HF can produce short-term
concentrations of up to 50 ,ug/m3. Davison reports fluoride
particulates range in diameter from <0.1 /urn 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 anyf
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 PM2.5-sensitive receptors and if so, how long this process
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.—
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
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Ridlington found that dietary Pb may reduce fluoride
bioavailability, but that fluoride does not influence Pb
utilization.— Further examination of HF and HC1 emissions may be
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
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
resistance of some tropical tree species may be related to their
capacity to accumulate Al.—
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
are limited, it is generally assumed that fluoride concentrations
do not increase with trophic level.— Resolving the question of
biomagnification is essential to evaluating the ecological
impacts of HF and HC1.
9. Information on the following SCI topics was not available or not
found.
Information pertaining to several topics concerning HC1 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 by-
products ;
2. terrestrial HCl transport, mobility, and partitioning;
3. aquatic stability and persistence of HCl and by-products;
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4. terrestrial animal HCl ingestion, inhalation, dermal
absorption, distribution, accumulation; and
5. aquatic animal HCl ingestion, distribution, and
accumulation.
H.5 GLOSSARY
Biomagnification
Chlorosis -
Fluorosis -
Hornblende -
Interspecific -
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.
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.
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.
A mineral, CaNa(Mg,Fe)4 (Al,Fe,Ti)3Si6022 (OH,F)2,
commonly green to black in color, formed late
in the cooling of igneous rock.
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.
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Hydrogen Fluoride Emission Factors for the NAPAP Emission
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Research Triangle Park, NC. July 1993.
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12. Patrinos, A. A., Dana, M. T., and Saylor, R. E. Wetfall.
Chemistry Studies Around a Large Coal-fired Power Plant in the
Southeastern United States. Journal of Geophysical Research.
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13. Clarke, A. G. and M. Radojevic. Chloride Ion Effects on the
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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.
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1.1 INTRODUCTION
1.1.1 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-eyelone 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
1-1
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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.
1.1.2 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 150°C [300 to 400°F]), 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 (HC1) in the flue gas stream can result in the
formation of mercuric chloride (HgCl2), which is readily adsorbed onto
carbon-containing particulate matter. Conversely, sulfur dioxide
{SC>2) in flue gas can act as a reducing agent to convert oxidized
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 150°C [300 to 400°F]),
the presence of an effective mercury sorbent and a method to
collect the sorbent. In general, high lesvels 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 HgCl2, which
is readily adsorbed onto carbon-containing particulate
matter, so it can be efficiently scrubbed by a wet FGD
system. Conversely, sulfur dioxide (SO2) in flue gas can
1-2
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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 HgCl2 by wet scrubbers).
• 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.
1.2 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.
1-3
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1.2.1 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.
1.2.1.1 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 heavy-
media-cyclone reduced mercury concentrations in the raw coal by 42 to
45 percent (ICF Kaiser Engineers, 1995). The water-only-cyclone and
1-4
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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 post-
combustion 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.
1.2.1.2 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. Microcel™ 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-Flote™ is another type of column froth
flotation cell.
1.2.2 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-
1-5
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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.
1.2.3 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 1.2.4 below.
1.2.3.1 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 1-1, 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 cultivated carbon
1-6
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Table 1-1. Comparison of Typical Uncontrolled Flue Gas Parameters at
Utilities and MWCsa-b
Uncontrolled Flue Gas
Parameters
Temperature (°C)
Mercury Content (/^g/dscrn)
Chloride Content (Atg/dscm)
Flow Rate (dscm/min)1
Coal-Fired Utility
Boilers*"
121 - 177
1-25
1,000-140,000
11,000-4,000,000
Oil-Fired
Utility Boilers"-"-'
121 - 177
0.2 - 2'
1,000-3,000
10,000-2,000,000
MWC9*
177-299
400-1,400
200,000 - 400,000
80,000 - 200,000
Standard conditions are 0°C and 1 atmosphere.
Moisture content in the MWC flue gas was assumed to be 13.2 percent.
Radian Corporation, 1993a, UNDEERC, 1996, CONSOL INC, 1997.
Heath, 1994.
Radian Corporation, 1994.
Radian Corporation, 1993b.
Brown and Felsvang.
Nebel and White, 1991.
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
essentially unstudied.
Min = minute
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 1-1 and presented in greater
detail in section 1.2.3.2. 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 (107-121°C, low carbon injection) to a high of 95 to 99 percent
with a median of 98 percent (88-107°C, 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.
1.2.3.1.1 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 HC1, nitrogen oxides, sulfur oxides, metal oxides on the
1-7
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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
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; Ca'rey 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 96°C (225 to 210 to 205°F), 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 121°C (190 to 250°F) 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 96°C
[190 to 205°F]), mercury concentrations were reduced by 97.7 percent
with an activated carbon injection rate of approximately 155 fj.g
carbon/pig of inlet mercury, while at higher temperatures (110 to 121°C
[230 to 250°F]) mercury concentrations were reduced by only 75 to 87
1-9
-------
percent with an activated carbon injection rate of approximately 3,500
//g carbon/yug 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
65 percent at 121°C (250°F) ; 67 percent at 93°C (200°F) , across a pilot-
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
average of 30 percent at 93 - 109°C (200 - 228°F) for the native fly
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 /ug/m3 levels, respectively,
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 /ug/m3 of the ionic form as measured by the respective methods. The
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
(346°F) , and baghouse outlet of 166°C (330°F) . Approximately 40 percent
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
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.
1-10
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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,
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
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 177°C (250 to 351°F) upstream of any particulate control device
to avoid acid condensation and, consequently, equipment corrosion.
The stack gas temperature may be lowered below 96°C (205°F) 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 medium-
to 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
1-11
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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 121°C (250°F), 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 ug/dscm, at
7 percent 02) . High mass carbon-to-mercury ratios will be required due
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 eire substantial
variations in the concentrations of elemental versus ionic mercury.
The percentage of Hg° is from near zero percent to >70 percent. The
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. Understcmding the rate
controlling mechanisms (i.e. transport, equilibrium, and kinetics)
will aid in predicting the species formed and eventually will aid in
1-12
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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
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 1-2 for coal- and oil-fired utility flue gas (Bloom et al,
1993) .
As shown in Table 1-2, the distribution of ionic mercury, most
likely HgCl2 in coal-fired utility flue gas, ranged from 12 to 99
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 1-2). 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 bench-
and 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;
1-13
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-------
and Laudal et al, December, 1997). Further studies at UNDEERC
indicated this finding is related to high S02 concentrations with the
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 SO2 and
NOX in the flue gas. When S02 is present in concentrations >500 ppm
and NOX is present at >250 ppm, the MESA method can overestimate the
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 HgCl2. Results showed that both the Ontario Hydro and U.S.
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)
1-17
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(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 subbiturrtinous 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 vapor-
phase 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 matericils 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
HC1 and possibly NOX affecting oxidation.
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
1-18
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flue gas conditions (with and without HCl and various levels of NOX ) .
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 HgCl2 in
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
1-19
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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
commercially available activated carbons. Dynamic adsorption capacity
of these carbons as high as 4,000 ug Hg/g was measured using a fixed-
bed absorber with an empty bed contact time of 0.Oil second and
influent mercury concentration of 55 ug/m3. This capacity is almost
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
- 1112°F) . This promoted a more uniform distribution of short linear
chains of sulfur allotropes (S2 and S6) on the carbon surface as
opposed to having predominately S8 rings condensed in the macropore
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
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 FF-
equipped 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 FF-
equipped facilities.
Further investigation, development, and enhancement of activated
carbons and chemically impregnated carbons for mercxiry 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).
1-20
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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 SO2 and HCl
on the adsorption of elemental mercury and mercuric chloride (HgCl2) by a
lignite-based activated carbon (Carey et al, 1997). Equilibrium
adsorption capacities were determined for fixed beds of the carbon at
275°F and three flue gas compositions: one containing 1,600 ppm S02 and
50 ppm HCl (the baseline composition) ; a second containing no S02 and 50
ppm HCl; and a third containing 1,600 ppm S02 and no HCl. (All three
compositions of flue gas had the same concentration of elemental mercury,
mercuric chloride, C02, water, and O2) .
Figure 1-2 illustrates the effect of SO2 and HCl on the
equilibrium adsorption capacity of the lignite-based activated carbon
for elemental mercury and mercuric chloride. Removing SO2 from the
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 S02 from the flue gas, the equilibrium
adsorption capacity for elemental mercury increased by a factor of
about 5.5 compared to 3.5 for mercuric chloride.
1-21
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Figure 1-2
Equilibrium Adsorption Capacity of Elemental Mercury
(Hg(0)) and Mercuric Chloride (HgC12) by
a Lignite-Based Activated Carbon
4J
•H
U
(0
u
fl
o
•H
JJ
a
M
o
[Q
•H
-H
•H
-H
&
16000-
14000-
12000-
10000-
8000-
6000-
4000-
2000-
o-
1200
1 1
HgC12
Baseline
4300
1200
, 1 1
HgC12, HgC12
No SO2 No HQ
2800
Hg(0)
Baseline
1J**VA.
Hg(0) Hg(0)
No SO2 No HQ
1-22
-------
Removing HC1 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 1-2 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 1-2 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 SO2 and
HCl on adsorption capacity could also be sorbent dependent. Other
carbons may not be affected by the presence of HCl and S02 if the
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.
1.2.3.1.2 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 full-
scale 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 1.2.3.2. 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 coal-
fired flue gas or, in other words, the system removed 72.5 percent of
the total mercury. During testing with activated carbon injection,
1-23
-------
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 iodide-
impregnated 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 200°C (257 to 392°F). Iodine-
impregnated 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 S02, HCl, NOX, HF, and possibly
CO. Lignite-based activated carbon removed approximately 50 percent
of elemental mercury at 130°C; 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 125°C, the iodine-impregnated carbon was
somewhat effective at removing oxidized mercury/ while it removed no
oxidized mercury at 175°C. 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,
1-24
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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 m2/g and a nominal
equilibrium adsorption capacity of 500 ug 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 ug 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
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°C
(222°F) upstream of an pilot-scale ESP with a residence time ranging
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 poal 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,
1-25
-------
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 te'sts at the PSCO facility also
indicated a substantial increase in carbon injection or mass carbon-
to-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)
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
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
>8700:1 were needed to remove mercury at 87 percent for a temperature
of 113°C (236°). The mercury concentrations were steady during these
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
approximately 12 ft/min (in this case EPRI's COHPAC or TOXICON). ESP
pilot-scale tests indicated mercury removals of 83 percent at 105°C
(221°F) and a mercury removal of 35 percent at 133°C (272°F) at the same
mass carbon-to-mercury ratio of 45,000:1. Low-sulfur Eastern
bituminous coal was fired at the utility and the fly ash mercury
1-26
-------
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-to-
mercury 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)
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 air-
to-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
Ib/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
1-27
-------
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 under-
utilized 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.
1.2.3.2 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 1-3 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 107°C (225°F) and 121°C (250°F), 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 88°C and 107°C, 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
107°C (225°F) and 121°C (250°F) 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 88°C (190°F) and 107°C (225°F) produced a
median mercury removal efficiency of 98 percent, with a range from
95 percent to 99 percent removal (Chang et al., 1993).
Table 1-4 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
1-28
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Table 1-3. Activated Carbon Injection Before Fabric Filter Data3
Unit
. t x „ J>, i ir^Sfe '» -*», *S
^=i'to:;hy«^lfeiSiKte^l4^ffi
Test #4, Run #1
Test #4, Run #2
Test #4, Run #3
Test #6, Run #3
sMt^!tomMSffljl««ri^
Test #5, Run #1
Test #5, Run #2
Test #5, Run #3
Test #6, Run #1
Test #6, Run #2
"Hp " "^"VlK, ''/ '•"• V
ttOWt^l^PBIwul^ 4*" i"ȣJB CStCr1
Test #2, Run #1
Test #2, Run #2
Test #2, Run #3
Test #3, Run #2
Test #3, Run #3
Control Device
** >
AC + FF (88°C (190°F) and 216 wt C/wt inlet Hg; inlet Hg
concentration of 5.35 ^g/dscm)
AC + FF (88°C (190°F) and 126 wt C/wt inlet Hg; inlet Hg
concentration of 8.19 /uo/dscm)
AC + FF (91 °C (196°F) and 123 wt C/wt inlet Hg; inlet Hg
concentration of 8.62 ,ug/dscm)
AC + FF (102°C (216°F) and 727 wt C/wt inlet Hg; inlet Hg
concentration of 1 .94 MQ/dscm)
^j^^^iiif^^^^affiMo^^ irttaW'1- & " • '
AC + FF (107°C (225°F) and 362 wt C/wt inlet Hg; inlet Hg
concentration of 5.53 MS/dscm)
AC + FF (1 10°C (230°F) and 373 wt C/wt inlet Hg; inlet Hg
concentration of 4.45 /ug/dscm)
AC + FF (1 16°C (241 °F) and 457 wt C/wt inlet Hg; inlet Hg
concentration of 3.47 /ug/dscm)
AC + FF (121 °C (250°F) and 286 wt C/wt inlet Hg; inlet Hg
concentration of 5.04 ^g/dscm)
AC + FF (1 18°C (244°F) and 367 wt C/wt inlet Hg; inlet Hg
concentration of 4.22 //o^dscm)
7l?S<*...v - - •„ ^3^'^- 'Si ^2% Ife^ ^ %Li L^V /."'•'*
EJII^ H^6*?ysM»-*fi!t|R5\^:^0tSi4JMl ^^^ef^ffM^^^^^-f^^ti^Lw^^^'-.
AC + FF (91 °C (196°F) and 2843 wt C/wt inlet Hg; inlet Hg
concentration not measured but assumed to be 7.00
MO/dscm)
AC + FF (96°C (205°F) and 3132 wt C/wt inlet Hg; inlet Hg
concentration not measured but assumed to be 7.00
//g/dscm)
AC + FF (93°C (200°F) and 3121 wt C/wt inlet Hg; inlet Hg
concentration not measured but assumed to be 7.00
MQ/dscm)
AC + FF (93°C (200°F) and 4361 wt C/wt inlet Hg; inlet Hg
concentration of 6.23 ^g/dscm)
AC + FF (96°C (205°F) and 3850 wt C/wt inlet Hg; inlet Hg
concentration of 6.91 Atg/dscm)
Hg removal %
^ ,,^, ^ -*,^
97
99
97
76
1.,,,, , *•• \
' ^^^ ,/-' ^
14
28
47
29
35
*&r. ' i,T, -:
:fc. ^' iM^j - >s*; ••' <
95
98
98
99
99
(continued)
1-29
-------
Table 1-3. (Continued)
Unit 1 Control Device
Test #3, Run #1
Test #7, Run #1
Test #7, Run #2
Test #7, Run #3
Hg removal %
AC + FF (1 10°C (230°F) and 3332 wt C/wt inlet Hg; inlet Hg
concentration of 7.95 ^g/dscm)
AC + FF (121 °C (250°F) and 1296 wt C/wt inlet Hg; inlet Hg
concentration of 4.66 ^g/dscm)
AC + FF (121 °C (250°F) and 1954 wt C/wt inlet Hg; inlet Hg
concentration of 4.30 ^g/dscm)
AC + FF (1 16°C (241 °F) and 3649 wt C/wt inlet Hg; inlet Hg
concentration of 2.09 ^g/dscm)
91
69
76
87
a Source: Chang et al., 1993
Table 1-4. Activated Carbon Injection Before Spray Dryer
Absorption Data3
Unit
/ NAfuMME fe '/
vrA. ^-j i^,-,sp*iflnS &,,~. A * -/
AC + SDA/FF (inlet Hg concentration unknown)
AC + SDA/FF (inlet Hg concentration of 3.9
Mg/dscm)
Hg Removal %
^ ** " ' -:
80.3, 85.8, 75.8, 74.5, 90.9,
89.5, 89.3, 86.7, 85.9
r . %" ., 5 ^". i
50-60
>99
8 Source: Felsvang, 1993
1-30
-------
efficiency of 60 percent, with a range from 50 percent to 99 percent
removal (Felsvang, 1993).
1.2.3.3 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 S02 emissions from
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 particulate-
bound 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
(HgCl2) (Redinger et al, 1997). The proportion of elemental mercury to
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, S02, and NOX. Because oxidized mercury is much more
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 pilot-
scale 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.
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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 tp 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 ug/m3 of mercury in
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 vapo-r 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
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
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dictated by the desired removal efficiency of S02. In general, high
efficiency (95 percent SO2 removal) systems are designed with L/G
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
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
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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 HgCl2 injected separately into the flue gas
stream at the inlet of the wet FGD. Results indicated 96 percent of
the HgCl2 (naturally occurring and spiked) was collected across the wet
FGD and the increase in Hg° across the FGD was from 0.66 to 0.96 ug/m3.
The results for the Hg° spiking indicated 37 percent of spike was
measured in the acidified peroxide solutions and the total Hg removal
was only 29 percent. These results indicated the injected HgCl2 was
being effectively collected in the scrubber solutions and not being
reduced and subsequently re-emitted as Hg°. In addition, M29 was not
effective in speciating the mercury at the inlet of this wet FGD
system when Hg° was spiked.
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 SO2 concentrations and moderate to
high concentrations of NOX) have an impact on the overestimation of the
oxidized form of mercury - S02 for the U.S. EPA M29 and the combination
of SO2 and NOX for the MESA. These findings are also detailed in the
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 HgCl2.
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 S02 in the flue gas as compared to the previous validation
studies for M29. The Ontario Hydro method was not impacted by the SO
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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.
1.2.3.4 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 S02 emissions from boiler flue gas;
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 S02 and total mercury from a spray dryer FGD
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 S02 and total mercury from a
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) .
1.2.4 Research and Emerging Technologies for Controlling Mercury
Emissions from Utilities
Considerable research continues to develop efficient and cost-
effective 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 1-5 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 coal-
fired utility boilers. The approaches included those listed in Table
1-5. DOE has selected six Phase II proposals (two to three year
efforts) to further investigate and develop fine particulate and
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Table 1-5. Current Mercury Control Research for Utility Boilers
Sponsor
U.S. Environmental Protection Agency
U.S. Department of Energy
Electric Power Research Institute
Research Area
Mercury speciation/High temperature control
Fundamental reactions/Low temperature control
Combined SOj/Mercury control
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)
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
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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 gas-
phase 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
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
Hg is predicted to exist as HgCl2. Between 725 and 975 K, the split
between HgCl2 and Hg is determined by the chlorine content of the coal
(via the HC1 content of the gas). HC1 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
HgCl2(g), if equilibrium is attained in the flue gas.
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 S02 do not have time to equilibrate as the
gas cools. For example, the oxidation of S02 to S03 in coal combustion
flue gas does not proceed at a fast rate below about 1500 K (Flagan
and Seinfeld, 1988) and thus the SO3 concentration is effectively
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 C12 is frozen as the flue gas cools
(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*2 to 95 percent Hg*2
upstream of the APCD (Prestbo and Bloom, 1990). There is some
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evidence from laboratory and pilot data that the kinetics of Hg
oxidation are slow at low temperatures. Based on pilot data, the
addition of HC1 at temperatures below 450°K (180° C) did not increase
the amount of HgCl2 in coal combustion flue gas, indicating no reaction
at those temperatures (Galbreath and Zygarlicke, 19.96) . 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).
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 750-900°K.
Kinetic calculations on the formation of C12, which is highly reactive
with elemental mercury, indicate that the conversion of HC1 to C12 does
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 (400-600°K) 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.
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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 420°K to 640°K (300°F to 700°F) .
Fly ash from five different coals was tested. At 420°K, 20-50 percent
of the elemental mercury was converted to a gaseous oxidized form,
probably HgCl2 based on equilibrium considerations, while 0-80 percent
was converted to an adsorbed form on the solids. The adsorbed species
might be HgCl2, HgO, or HgS04. There was a wide variation in the
amount of adsorbed mercury depending on coal type. At 640°K, 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 N2, 02, C02/ S02, H20, HCl, and elemental mercury. In order
to better understand the forms of adsorbed mercury, X-ray absorption
fine structure (XAFS) spectra were collected at the mercury Lm edge at
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 Cl-
edge 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.
1.2.4.1 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
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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
a more uniform distribution of short linear chains of sulfur
allotropes (S2 and S6) on the carbon surface as opposed to having
predominately Ss rings condensed in the macropore region of
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
impregnated carbons exhibited high elemental mercury uptake efficiency
at 140°C (284°F) when compared to commercially available activated
carbons. Dynamic adsorption capacity of these carbons were measure as
high as 4,000 ug 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 carbon-
bound iodide reacts with mercury to form mercuric iodide (HgI2) on the
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
HgCl2 on the carbon, and the carbon is removed by a particulate control
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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 1.2.3.1.2, "Current Research on Activated Carbon
Injection for Utilities" (Waugh et al, 1997).
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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°C
(300°F) . Compounding the mercury mass transfer limitation(s) is the
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)2) with
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)2-impregnated
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 low-
cost, 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 Na2S solution into the flue gas prior to the acid gas control
device. The specific reactions of Na2S and Hg are not totally clear
but appear to be (Nebel and White, 1991) :
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Hg (gas) + Na2S 4- 2H20 =» HgS (Solid) + 2NaOH + H2 and
HgCl2 (gas) + Na2S =» HgS (Solid) + 2NaCl.
The resulting solid, HgS, can be collected by a FF.
There are several potential limitations to Na2S injection. These
include reaction of Na2S with calcium (Ca) in the sorbent (as found in
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 (Na2S is a corrosive
material), clogging and plugging of the screw conveyor due to
solidification of Na2S, and sludge formation due to the presence of
inorganic salts in the mixing water (Nebel and White, 1991).
At present, full-scale operational injection of Na2S has been
done only in MWCs. Mo 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 S02, HCl, and H20. Mercsorbent can be used for sorbent
injection or it can be used as a coating on a FF. A bench-scale duct-
injection 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
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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 ug/m3 to zero) at 110° C (230° F) . The iodine-impregnated
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
ug/m3 after 100 percent breakthrough (carbon fully utilized) was
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).
1.2.4.2 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 S02 and
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, bench-
scale 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
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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/m3 to less than 5 mg/m3 (<0.005 Ibs/MMBtu) at a
gas temperature of 150°C (300°F) . The particles in the 2.5 micron
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 W2.3.2, 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
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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 difficult-
to-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) .
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/m3 to less then 1 Aig/m3 at the outlet after more than
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 S02, S03, particulates, and trace elements from
coal combustion flue gas. The condensing heat exchanger is a two-
pass, 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 upwaird through the
second cooling stage. An alkali reagent is sprayed from the top of
1-46
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the second stage to aid in the removal of S02. Testing of the
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-LIDS™) process
combines furnace limestone injection with dry scrubbing to achieve
high efficiency SO2 particulate, and trace element emissions control.
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 SO2 in the flue gas. The flue gas passes through a
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 S02 scrubbing reagent for the
spray dryer.
Application of the E-LIDS™ 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 Ibs S02/106 Btu.
Total mercury emissions were reduced from an uncontrolled level of
17.6 /ug/dscm to less than 0.2 ,ug/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 106 Btu/hr small boiler simulator (SBS) facility. Mercury
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 ^ug/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 SO3 to serve
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).
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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 HC1 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).
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