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
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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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-------
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

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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

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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

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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

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Appendix E - Health Effects Summaries:  Overview

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      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

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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-
                                  E-2

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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
                                   E-3

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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

                                  E-4

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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).

                                  E-5

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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
                                  E-6

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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

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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.
                                  E-22

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E.23  REFERENCES

1.   U.S.  Environmental  Protection Agency.   Integrated Risk
     Information System  (IRIS)  Database,  National  Center for
     Environmental  Assessment,  Cincinnati,  OH.

2.   U.S.  Environmental  Protection Agency.   Health Effects  Notebook
     for Hazardous  Air Pollutants.   EPA-456-d-94-1003.   Air Risk
     Information Support Center,  RTF, NC  27711.

3.   U.S.  Environmental  Protection Agency.   Drinking Water  Regulations
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4.   ATSDR.   Toxicological  Profile for Arsenic  (Update).  Agency for
     Toxic Substances  and Disease Registry.  U.S.   Public Health
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5.   U.S.  Environmental  Protection Agency.   Integrated Risk
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6.   Ref.  5,  Carcin-3.

7.   Ref.  5,  Carcin-4.

8.   Electric Power Research Institute.   Electric  Utility Trace
     Substances  Synthesis Report.   EPRI TR-104614-V4.   pg.  G-l,  1994.

9.   Ref.  3.

10.  Ref.  5,  Carcin-3.

11.  Ref.  5,  Carcin-3.

12.  Ref.  4,  Section 2.2.

13.  Ref.  4,  Section 2.2.2.1.

14.  Ref.  4,  Section 2.2.

15.  Ref.  4,  Section 2.2.2.

16.  Ref.  5,  RfD-1.

17.  Tseng, W.P., H.M.   Chu,  S.W.   How, et  al.   Prevalence  of  skin
     cancer in an endemic area  of chronic arsenium in  Taiwan.  J.
     Natl.  Cancer  Inst.  40:453-463.  1968.   (as  cited in  Reference
     5)

18.  Ref.  5,  RFC-1.

19.  Ref.  5,  RfD-5  & 6.
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20.  Ref.  4,  Section  2.2.1.

21.  Ref.  4,  Section  2.2.

22.  Ref.  3.

23.  ATSDR.   Toxicological Profile for Chromium.  Agency for Toxic
     Substances and Disease Registry.  U.S.  Public Health Service.
     Section  2.2.  1991.

24.  Environmental Protection Agency.  Integrated Risk Information
     System on Chromium VI.  Carcin-1.  1995.

25.  Ref.  23, Section 2.2.

26.  Ref.  24, Carcin-1.

27.  Ref.  24, Carcin-1.

28.  Ref.  24, Carcin-1.

29.  Environmental Protection Agency.  Integrated Risk Information
     System on Chromium III.  Carcin-1.  1995.

30.  Ref.  24, Carcin-2.

31.  Ref.  24, Carcin-3.

32.  Ref.  24, Carcin-2.

33.  Ref.  29, Carcin-2.

34.  Ref.  24, Carcin-3.

35.  International Agency for Research on Cancer.  IARC Monographs on
     the  Evaluation of Carcinogenic Risks to Humans:  Chromium,
     Nickel,  and Welding.  Volume 49.  Lyon, France,  pp. 208-214.
     1990.

36.  Ref.  23, Section 2.2.

37.  Ref.  23, Section 2.2.

38.  Ref.  23, Section 2.2.

39.  Ref.  23, Section 2.2.

40.  Ref.  24, RfD-l.

41.  Ref.  24, RfD-2.
                                  E-24

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42.  Ref.  24,  RfD-3.

43.  Ref.  29,  RfD-1.

44.  Ref.  29,  RfD-3.

45.  Ref.  29,  RfC-1.

46.  Ref.  24,  RfC-1.

47.  Ref.  23,  Section  2.2.1.

48.  Ref.  23,  Section  2.2.

49.  The Merck Index:  An. Encyclopedia of Chemicals, Drugs, and
     Biologicals  (llth edition).  Ed. S. Budavari.  Merck & Co.,  Inc.
     Rahway, NJ.  p.756.  1989.

50.  HSDB.  Hazardous  Substances Databank  (online database).  U.S.
     Department of Health and Human Services, National Library of
     Medicine,  National Toxicology Information Program.  Bethesda, MD.
     1992.

51.  Environmental Protection Agency.  Integrated Risk Information
     System on Hydrogen Chloride.  Carcin-1.  1995.

52.  Ref.  50.

53.  Ref.  50.

54.  RTECS.  Registry  of Toxic Effects of Chemical Substances.  U.S.
     Department of Health and Human Services, National Library of
     Medicine,  Bethesda, MD.  1992.

55.  Ref.  50.

56.  Ref.  51,  RfC-2.

57.  Ref.  51,  RfC-1.

58.  Ref.  51,  RfC-4.

59.  Ref.  51,  RfC-4.

60.  ATSDR.  Toxicological  Profile for Fluorides, Hydrogen Fluoride,
     and Fluorine.  Agency  for Toxic Substances and Disease Registry.
     US Public Health  Service.  Section 1.  1993.

61.  Ref.  3.

62.  Ref.  60,  Section  2.2.
                                 E-25

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63.  U.S. Environmental Protection Agency.  Integrated Risk
     Information System on Hydrogen Fluoride.  Carcin-1.  1995.

64.  Ref. 50.

65.  Ref. 60, Section 2.2.

66.  Ref. 60, Section 2.2.

67.  Ref. 60, Section 2.2.1.

68.  U.S. Environmental Protection Agency.  Health Issue Assessment:
     Summary Review of Health Effects Associated with Hydrogen
     Fluoride and Related Compounds.  Environmental Criteria and  '
     Assessment Office, Office of Health and Environmental Assessment,
     ORD, Cincinnati, OH.  Section 7.2.  1989.

69.  Ref. 60, Section 2.2.

70.  Ref. 68, Section 7.2.

71.  Ref. 63, RfC-1, RfD-1.

72.  Ref. 60, Section 2.2.

73.  Ref. 60, Section 2.2.

74.  U.S. Environmental Protection Agency.  Mercury Study Report to
     Congress.  Volume V.  Health Effects of Mercury and Mercury
     Compounds.  EPA-452/R-97-007.  Office of Air Quality Planning and
     Standards and Office of Research and Development.  1977.

75.  Ref. 35, pp. 407, 408.

76.  U.S. Environmental Protection Agency.  Integrated Risk
     Information System on Nickel Refinery Dust.  Carcin-1.  1995.

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.

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Appendix F — Documentation of The Inhalation Human Exposure Modeling
                        for  the Utility Study

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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

-------
      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

-------
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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F-6

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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

-------
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
Water Ingestion
Fish Ingestion
Direct Inhalation

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
3.56-12
2.1e-12
6.5e-10
1.86-15


8.56-15
4.56-13
2.06-12
1.26-12
3.66-10
1.86-15


9.16-16
4.86-14
4.7e-13
1.26-13
3.86-11
1.86-15

1.6e-13
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
Vegetation Ingestion
Animal Product Ingestion
Water Ingestion
Direct Inhalation

SOH-farmer
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Animal Product Ingestion
Water Ingestion
Direct Inhalation

LCC-resident
Soil Dermal
Soil Ingestion
Vegetation Ingestion
Water Ingestion
Direct Inhalation

LOC-resident
Soil Dermal
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
1.26-12
1.8e-15


5.0e-15
4.8e-14
1.0e-12
9.4e-12
1.2e-13
1.86-15


1.16-15
5.8e-14
2.4e-13
1.2e-13
2.96-15


5.86-16
3.16-14
1.86-13
7.86-14
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
2e-07
3e-10


8e-10
8e-09
2e-07
1e-06
2e-08
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%
100.00%


4.48%
4.48%
45.66%
18.54%
5.75%
99.97%


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

-------
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.


                                  H-3

-------
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

-------
      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
                                  H-6

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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.


                                  H-7

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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


                                  H-8

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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".

                                 H-10

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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


                                 H-ll

-------
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
                                  H-12

-------
      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


                                 H-13

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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

                                  H-14

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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.—


                                 H-15

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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.—
                                  H-16

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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


                                 H-17

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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-


                                 H-18

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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.


                                  H-20

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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


                                  H-21

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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


                                  H-22

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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.
                                  H-23

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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


                                 H-24

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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
                                  H-26

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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


                                  H-28

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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


                                 H-29

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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

                                 H-32

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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;

                                 H-35

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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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H.6  REFERENCES

1.   Ruprecht,  H.  and Sigg,  L.   Interactions  of  Aerosols  (Ammonium
     Sulfate, Ammonium Nitrate,  and Ammonium  Chloride)  and of  Gasses
      (HCl,  HN03) with Fogwater.  Atmospheric Environment.  Volume 24A,
     No.  3.   1990.  p.  573-584.

2.   Wegner,  et al.   Variations  of  tropospheric  HCl  amounts over
     Northern Sweden.   Geophysical  Research Letters.  Volume 24, No.
     8.   1997.   p.  849-852.

3.   Puxbaum, et al.   Seasonal variation of HN03, HCl, S02,  NH3 and PM
     at a rural site  in NE Austria,   Atmospheric Environment.  Volume
     27A, No. 15.   1993.   p.  2445-2447.

4.   Graedel, T. E. and W. C. Keene.   Tropospheric Budget  of Reactive
     Chlorine.   Global  Biogeochemical Cycles.  Volume 9, No. 1.  1995.
     p. 47-77.

5.   Lightowlers,  P.  J.  and  J. N. Cape.   Sources and Fate  of
     Atmospheric HCl  in the  U.K.  and Western  Europe.  Atmospheric
     Environment.  Volume  22., No.  1.   1988.  p.  7-15.

6.   Graedel, et al.  Global Emissions Inventories of Acid-Related
     Compounds.  Water,  Air,  and Soil Pollution.  Volume 85.   1995.
     p. 25-36.

7.   Kelly, T.  J.,  et al.  Concentrations and Transformations  of
     Hazardous  Air  Pollutants.   Environmental Science & Technology.
     Volume 28,  No. 8.   1994.  p. 378A-387A.

8.   Hutchinson, A. J.,  et al.   Stone Degradation Due to Dry
     Deposition of HCl  and S02 in a Laboratory-Based Exposure Chamber.
     Atmospheric Environment.  Volume 26A,  No. 15.   1992.   p.  2785-
     2793.

9.   U.S. Environmental Protection  Agency.  Hydrogen Chloride  and
     Hydrogen Fluoride  Emission  Factors for the  NAPAP Emission
     Inventory.  EPA/600/7-85-041.   Air and Energy Research
     Laboratory, Research  Triangle  Park,  NC.  October 1985.

10.  lapalucci,  T.L., R.J. Demski,  and D.  Bienstock.  Chlorine in Coal
     Combustion.  United States  Department of the Interior,  Bureau of
     Mines Report  of  Investigation  7260.   May 1969.

11.  Spicer,  et al.   A  Literature Review of Atmospheric
     Transformations  of Title III Hazardous Air  Pollutants,  Final
     Report.  U.S. Environmental Protection Agency.   EPA/600/R-94/088.
     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.
     Volume  88.  1983.  p.  8585-8612.

13.  Clarke, A. G. and M. Radojevic.  Chloride Ion Effects on the
     Aqueous Oxidation of S02.   Atmospheric Environment.   Volume 17.
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37.  Elrashidi, M. A., and W. L. Lindsay.  Solubility of Aluminum
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39.  Hodge, H. C. and F. A. Smith.  Fluorine Chemistry, Volume IV -
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42.  Klumpp, et al.  Fluoride Impact on Native Tree Species of the
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43.  Davies, F. B. M. and G. Notcutt.  Accumulation of Fluoride by
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44.  Amundson, R. G., A. J. Belsky and R. C. Dickie.  Fluoride
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46.  Boulton, I. C., J. A. Cooke,  and  M. S.  Johnson.  Experimental
     Fluoride Accumulation and Toxicity  in the Short-Tailed Field  Vole
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47.  Boulton, I. C., J. A. Cooke,  and  M. S.  Johnson.  Fluoride
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     Volume  15,  No.  6.  1995.   p.  423-431.

48.  Seel, D. C., A. G. Thomson,  and R.  E. Bryant.  Bone Fluoride  in
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49.  Cooke,  J.  A.,  S.  M.  Andrews  and M.  S. Johnson.   The Accumulation
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     55-63.

50.  Boulton, Cooke, and  Johnson.   Lead, Zinc, Cadmium, and Fluoride
     in  Small Mammals  from Contaminated  Grassland Established on
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     No. 1-2.   1990.   p.  43-54.

51.  Van Wensem, J., and  T. Adema.   Effects  of Fluoride on Soil  Fauna
     Mediated Litter Decomposition.  Environmental  Pollution.  Volume
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52.  Buse, A.   Fluoride Accumulation in  Invertebrates Near an Aluminum
     Reduction  Plant in Wales.  Environmental  Pollution.  Volume 41A.
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53.  Wang, J.,  and  Y.  Bian.  Fluoride  Effects  on the  Mulberry-Silkworm
     System.  Environmental Pollution.   Volume 52.  1988.  p.  11-18.

54.  Cerklewski, F.  L., and J.  W.  Ridlington.  Influence of Dietary
     Lead on Fluoride  Bioavailability  in the Rat. Biological  Trace
     Element Research.  Volume 14.   1987.  p.  105-113.

55.  Susheela,  A. K.,  and M. Bhatnagar.  Fluoride Toxicity:   A
     Biochemical and Scanning  Electron Microscopic  Study of Enamel
     Surface of Rabbit Teeth.   Archives  of Toxicology.  Volume 67.
     1993.   p.  573-579.
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56.  Pleasants,  E. W. ,  C. Waslien, and B. A. Naughton.  Dietary
     Modulation  of the  Symptoms of Cadmium Toxicity in Rats:  Effects
     of Vitamins A,  C,  D, D Hormone, and Fluoride.  Nutrition
     Research.   Volume  13. 1993.  p. 839-850.

57.  Yu,  et  al.  Effect of Dietary Fluoride on Selenite Toxicity in
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58.  Zeiger,  et  al.   Genetic Toxicity of Fluoride.  Environmental and
     Molecular Mutagenesis.  Volume 21.  1993.  p. 309-318.

59.  Camargo,  J. A.  and J. V. Tarazona.  Short-Term Toxicity of
     Fluoride Ion  (F~)  in Soft Water to Rainbow Trout and Brown Trout.
     Chemosphere.  Volume 22, Nos. 5-6.  1991.  p. 605-611.

60.  Camargo,  J. A.  and J. V. Tarazona.  Acute Toxicity to Freshwater
     Benthic Macroinvertebrates of Fluoride Ion in Soft Water.
     Bulletin of Environmental Contamination and  Toxicology.  Volume
     45.   1990.  p.  883-887.

61.  Khan, A., D. Kent,  J. Barbieri, and S. Khan.  Chronic Toxicity of
     a Fluoride  Mixture to Freshwater Organisms.  Water Science and
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62.  Joy,  C.  M., and K.  P. Balakrishnan.  Effect  of Fluoride on Axenic
     Cultures of Diatoms.  Water, Air, & Soil Pollution.  Volume 49.
     1990.   p. 241-249.
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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
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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.
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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

-------
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

-------
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

-------
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

-------
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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                                               1-16

-------
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

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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

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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

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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

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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

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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.


                                  1-31

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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


                                  1-32

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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


                                  1-33

-------
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


                                  1-34

-------
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


                                  1-35

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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
                                 1-36

-------
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


                                  1-37

-------
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


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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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I.3  REFERENCES

ABB, et al,  "Ultra High Efficiency ESP Development for Air Toxics,"
Draft Final  Report under Phase I DOE/FETC MegaPrda Program, period of
performance  - Sept. 1995 to July 1997.

Akers, David, R. Dospoy, C. Raleigh, 1993.  The Effect of Coal
Cleaning on  Trace Elements, Draft Report, Development of Algorithms.
December 16, 1993.  Prepared for EPRI by CQ, Inc.

Armstrong, M., 1994.  March 1994 notes from Mike Armstrong, Arkansas
Game and Fish Commission.

Attari, A. and S. Chao, 1993.  "Quality Survey of Natural Gas in the
United States."  Presented at the 1993 AICHE Spring National Meeting,
Houston, Texas.

Bailey, R. T., B. J. Jankura and K. H. Schulze, "Preliminary Results
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