WATER QUALITY CRITERIA


United States
Environmental Protection
Agency
Office of Science and Technology September 1993
Office of Water
Washington, D.C. 20460
SEPA 1993
Water Quality
Criteria

-------
Quality Criteria
for Water
1993
Prepared for
Health and Ecological Criteria Division
Office of Water
U.S. Environmental Protection Agency
Washington, D.C.
1993

-------
NOTE TO USERS
Quality Criteria for Water 1993 is produced as a summary document.
Individual criteria in this document are taken from previously published
criteria. To obtain further information on these criteria, review the
information in Appendix E. Each criteria entry lists the Federal Register
notification and publication date.
This publication contains criteria issued since 1976. Different meth-
odologies have been used to derive these criteria. Appendixes A through
D provide the various available methodologies; notes at the end of each
criterion indicate which methodology applies.
Individuals who use this document are encouraged to obtain a copy
of the Criteria Document for the Pollutant of Interest. These documents con-
tain the data sets and references upon which the criteria in this summary
document were derived.

-------
PREFACE
Section 304(a)(1) of the Clean Water Act (33 U.S.C. 1314[a][l] requires the
Environmental Protection Agency (EPA) to publish and periodically
update ambient water quality criteria. Intended neither as rules nor
regulations, these criteria present scientific data and guidance on the
environmental effects of pollutants that can be used to derive regulations
based on considerations of water quality impacts.
These criteria are to accurately reflect the latest scientific knowledge
(a)	on the kind and extent of all identifiable effects on health and welfare
including, but not limited to, plankton, fish, shellfish, wildlife, plant life,
shorelines, beaches, aesthetics, and recreation that may be expected from
the presence of pollutants in any body of water including groundwater;
(b)	on the concentration and dispersal of pollutants, or their byproducts,
through biological, physical, and chemical processes; and (c) on the ef-
fects of pollutants on biological community diversity, productivity, and
stability, including information on the factors affecting rates of eutrophi-
cation and organic and inorganic sedimentation for varying types of
receiving waters.
The first of these publications appeared in 1968 with The Report of the
National Technical Advisory Committee to the Secretary of the Interior "Green
Book." Water Quality Criteria 1972 (the "Blue Book") was published in
1973, followed three years later by the "Red Book," Quality Criteria for
Water.
On November 28, 1980 (45 F.R. 79318), EPA announced through the
Federal Register the publication of 64 individual ambient water quality
criteria documents for pollutants listed as toxic under section 307(a)(1)
of the Clean Water Act. On February 15,1984 (49 F.R. 5831); July 29,1985
(50 F.R. 30784); March 7, 1986 (51 F.R. 8012); June 24,1986 (51 F.R. 22978);
December 3, 1986 (51 F.R. 43665); and March 2, 1987 (52 F.R. 6213); EPA
published additional water quality criteria documents, followed by
Quality Criteria for Water 1986 (the "Gold Book"). The National Toxics
Rule (NTR) was promulgated on December 22, 1993 (57 F.R. 60848). The
NTR was a national rulemaking that provided the most current criteria
for the priority pollutants. This rulemaking recalculated the priority pol-
lutants criteria, based on data that was available at its release. Quality
Criteria for Water 1993 draws information from its predecessors in pre-
senting summaries of all the contaminants for which EPA has developed
criteria recommendations. The rationale for these recommendations can
be found in the reference identified at the end of each criteria summary.

-------
This document is intended as a summary only. Specific data on each cri-
terion can be found in individual criteria documents. Copies of the
individual ambient water quality criteria documents containing all the
data used to develop the criteria recommendations summarized here,
are available from National Technical Information Service, 5285 Port
Royal Road, Springfield, VA 22161, (703) 487-4650.
This book is intended for easy reference use. The Contents lists com-
pounds by their common names.
To obtain copies of this document and supplements, contact the Gov-
ernment Printing Office at (202) 783-3238.
EPA's goal is to continue to develop and make available ambient
water quality criteria reflecting the latest scientific information. This, we
believe, constitutes a major component in our ongoing commitment to
improve and protect the quality of our Nation's waters.
Margaret Stasikowski
Director, Health and Ecological Criteria Division
For further information contact:
HUMAN HEALTH
Dr. Frank Gostomski
Chief, Surface Water Health Assessment Section
U.S. Environmental Protection Agency
401 M Street, SW (WH-586)
Washington, DC 20460
AQUATIC LIFE
Margarete Heber
Chief, Criteria Section
U.S. Environmental Protection Agency
401 M Street, SW (WH-586)
Washington, DC 20460

-------
WATER QUALITY CRITERIA SUMMARY
CONCENTRATIONS (in ng/L)
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Science and Technology
Health and Ecological Criteria Division
Ecological Risk Assessment Branch (WH-586)
Human Risk Assessment Branch (WH-586)
CAS#
Sd
tr O
o. a
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)
PUBLISHED CRITERA
FRESH	FRESH	SALTWATER SALTWATER
ACUTE	CHRONIC ACUTE CHRONIC WATER 8 ORGANISMS
CRITERIA CRITERIA CRITERIA CRITERIA ORGANISMS ONLY
RECALCULATED VALUES
using IRIS, as of 9/90
WATER &
ORGANISMS
ORGANISMS
ONLY
DRINKING
WATER MCL
CRITERIA
FEDERAL
REGISTER
NOTICE
ACENAPHTHENE	83-32-9 Y N "1,700. *520.	*970.	*710	1,200.	2,700	57 FR 60890
ACENAPHTHYLENE
208-96-8
Y
Y





57 FR 60913
ACROLEIN
107-02-8
Y
N
*68. *21. *55. 320.
780



45 FR 79324
ACRYLONI'I'RILE
107-13-1
Y
Y
*7,550. *2,600. 0.058
0 65
0.059
0.66

45 FR 79324
ALACHLOR
15972-60-8
N
Y




20
—
ALDRIN
309-00-2
Y
Y
3.0 1.3 0 000074
0 000079
0 00013
0.00014

45 FR 79325
ALKALINITY
	
N
N
20,000




RB
ALUMINUM
7429-90-5
N
N
CRITERIA ARE pH DEPENDENT — SEE DOCUMENT




53 FR 33178
AMMONIA
7664-41-7
N
N
CRITERIA ARE pH AND TEMPERATURE DEPENDEN 1 — SEE DOCUMEN T



54 FR 19227
AN1HRACENE
120-12-7
Y
Y


0.00012
0.00054

57 FR 60913
ANTIMONY
7440-36-0
Y
N
/p/88 /p/30 /p/1500 /p/500 146
45,000
14.
4,300.

45 FR 79325
ARSENIC
7440-38-2
Y
Y
0.0022
0.0175
0.018
0.14

45 FR 79325
ARSENIC(V)
17428-41-0
Y
Y
*850. *2,319.




50 FR 30789
ARSENIC(III)
22569-72-8
Y
Y
360. 190. 69. 36.




50 FR 30786
ASBESTOS
1332-21-4
Y
Y
30k fibers/L



7 MFL
57 FR 60911
ATRAZINE
1912-24-9
N
N




3 0
—
BACTERIA
	
N
N
FOR PRIMARY RECREATION AND SHELLFISH USES — SEE DOCUMENT



51 FR 8012
BARIUM
7440-39-3
N
N
1,000



/ p/2.000
RB
BENZENE
71-43-2
Y
Y
*5,300. *5,100 *700. 0.66
40.
1 2
71.
5.0
57 FR 60911
BENZIDINE
92-87-5
Y
Y
*2,500. 0.00012
0 00053
0 00012
0 00054

57 FR 60913
BENZOFLUORAN THENE, 3,4-
205-99-2
Y
Y


0.0028
0.0311

57 FR 60913
BENZO(A)
ANTHRACENE
56-55-3
Y
Y


0.0028
0.0311

57 FR 60913
50-32-8 Y Y	0.0028	0.0311	57 FR 60913

-------
CAS#
0C 3
2d
cc O
a a.
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)
PUBLISHED CRITERA
FRESH	FRESH SALTWATER SALTWATER
ACUTE	CHRONIC ACUTE	CHRONIC WATER & ORGANISMS
CRITERIA CRITERIA CRITERIA CRITERIA ORGANISMS ONLY
RECALCULATED VALUES
using IRIS, as of 9/90
WATER &
ORGANISMS
ORGANISMS
ONLY
CRITERIA
FEDERAL
DRINKING REGISTER
WATER MCL NOTICE
BENZO(K)
FLUORANTHENE	207-08-9 Y Y	0.0028	0.0311	57 FR 60913
BERYLLIUM
7440-41-7
Y
Y
*130.
*5.3


0.0037
0.0641



57 FR 60848
BETA PARI ICLE and













PHOTON ACTIVITY
—
N
Y








4 mrem
—
BHC
680-73-1
N
Y
*100.

*0 34






45 FR 79335
BROMOFORM
75-25-2
Y
Y






4.3
360.

57 FR 60911
BUTYLBENZYL













PHTHALATE
85-68-7
Y
N






3,000.
5,200.

57 FR 60890
CADMIUM
7440-43-9
Y
N
3.9+
1.1 +
43.
9.3
10.



50
57 FR 60848
CARBOFURAN
1563-66-2
N
N








40.
—
CARBON TE'l'RACHLORIDE
56-23-5
Y
Y
*35,200.

*50,000.

0 4
6.94
0.25
4.4
5.0
57 FR 60911
CHLORDANE
57-74-9
Y
Y
2.4
0.0043
0.09
0.004
0 00046
0.00048
0.00057
0.00059
2.0
45 FR 79327
CHLORIDE
16887-00-6
N
N
860,000.
230,000.







53 FR 19028
CHLORINATED













BENZENES
—
Y
Y
*250.
*50
*160.
*129.


488.


45 FR 79327
CHLORINATED













NAPHTHALENES
—
Y
N
*1,600.

*7.5



1,700.
4,300.

57 FR 60890
CHLORINE
7782 50-5
N
N
19.
11.
13
7.5





50 FR 30788
CHLOROALKYL













ETHERS

Y
N
*238,000.








45 FR 79330
CHLOROBENZENE
108-90-7
Y
N




488

680.
21,000.
100.
57 FR 60911
CHI.ORODIBRO-













MOMETHANE
124-48-1
Y
Y






0.41
34.

57 FR 60911
CHLOROFORM
67-66-3
Y
Y
*28,900.
*1,240.


0.19
15.7
5.7
470.

57 FR 60911
CHLOROPHENOL, 2-
95-57-8
Y
N
*4,380.





120.
400.

57 FR 60890
CHLOROPIIENOL, 4
106-48-9
N
N


*29,700.






45 FR 79329
CHLOROPHENOL, 4 ,













METHYL, 3-
59-50-7
Y
N
*30.








45 FR 79329
CHLOROPHENOXY













HERBICIDE (2,4,5,-TP)
93-72-1
N
N




10



50
RB
CHLOROPHENOXY













HERBICIDE (2,4-D)
94-75-7
N
N




100.



70.
RB
CHLORPYRIFOS
2921-88-2
N
N
0.083
0.041
0.011
0.0056





51 FR 43666
CHROMIUM (VI)
7440-47-3
Y
N
16.
11.
1,100
50.
50.

170.
3,400.
100.
50 FR 30788

-------
CAS#
fcfS
s 3
Q d
DC O
q. a.
FRESH
ACUTE
CRITERIA
FRESH
CHRONIC
CRITERIA
SALTWATER
ACUTE
CRITERIA
SALTWATER
CHRONIC
CRITERIA
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)
PUBLISHED CRITERA
WATER &
ORGANISMS
RECALCULATED VALUES
using IRIS, as of 9/90
ORGANISMS
ONLY
WATER &
ORGANISMS
ORGANISMS
ONLY
DRINKING
WATER MCL
CRITERIA
FEDERAL
REGISTER
NOTICE
CHROMIUM (III)	1308-14-1 Y N 1,700.- 210.+	*10,300	170,000 3,433,000 33,000.	670,000. 100.	50 FR 30788
CHRYSENE
218-01-9
Y
Y




0 0028
0.0311

57 FR 60913
COLOR
—
N
N
NARRATIVE STATEMENT — SEE DOCUMENT





RB
COPPER
7440-50-8
Y
N
18 +
12.+ 2.9




/ p/1300.
50 FR 30789
CYANIDE
57-12-5
Y
N
22.
5.2 1.0
200

700.
220,000.

57 FR 60911
DDT
50-29-3
Y
Y
1.1
0 001 0.13 0 001
0.000024
0 000024
0.00059
0.00059

57 FR 60914
DDT METABOLITE











(DDD) ( IDE)
72-54-8
Y
Y
*0.6
*3.6


0 00083
0.00084

57 FR 60914
DDT ME TABOLITE











(DDE)
72-55-9
Y
Y
*1,050.
*14.


0 00059
0.00059

57 FR 60914
DEM ETON
8065-48-3
N
N

0.1 0.1





RB
DIBENZO(A,H)











ANTHRACENE
53-70-3
Y
Y




0.0028
0.0311

57 FR 60913
DIBROMOCHLORO-











PROPANE
96-12-8
N
Y






0.2
—
DI-N-BUTYL PHTHALATE
84-74-2
Y
N


34,000.
154,000.
2,700.
12,000.

57 FR 60913
DICHLOROBENZENE, 1,2-
95-50-1
Y
N




2,700.
17,000.
600.
57 FR 60913
DICHLOROBENZENE, 1,3-
541-73-1
Y
N




400.
2,600.
600.
57 FR 60913
DICHLOROBENZENE, 1,4-
106-46-7
Y
N




400.
2,600.
75.
57 FR 60913
DICHLOROBENZENES
25321-22-6
N
N
*1,120.
*763. *1,970.
400
2,600.



45 FR 79328
DICHLOROBENZIDINE, 3,3-
91-94-1
Y
Y


0 0103
0 0204
0.04
0.077

57 FR 60913
DICHLOROBROMO-











METHANE
75-27-4
Y
Y




0.27
2.2

57 FR 60911
DIC11LOROEITIANE, 1,2-
107-06-2
Y
Y
*118,000.
*20,000. *113,000
0 94
243
0.38
99.
5.0
57 FR 60912
DICHLOROE H IYLENE5
25323-30-3
N
Y
*11,600
*224,000.





45 FR 79332
DICHLOROEITIYLENE, 1,1-
75-35-4
Y
Y


0 033
1 85
0 057
3.2
7.0
57 FR 60912
DICHLOROE H IYLENE,











cis-1,2-
—
N
N






70
—
DICHLOROEITIYLENE,











trans, 1,2-
156-60-5
Y
N




700

100.
57 FR 60890
DICHLOROP! IENOL, 2,4-
120-83-2
Y
N
*2,020.
*365.
3090

93
790.

57 FR 60912
DICHLOROPROPANE
26638-19-7
N
N
*23,000.
*5,700. *10,300. *3,040.




5.0
45 FR 79333
DICHLOROPROPANE, 1,2-
78-87-5
Y
Y




0 52
39.
5 0
57 FR 60890
DICHLOROPROPENE
26952-23-8
N
N
*6,060
*244. *790
87.
14,1(X)



45 FR 79333
DICHLOROI'ROPYLENE, 1,3-
542-75-6
Y
N




10.
1,700.

57 FR 60912

-------
CAS#
Qd
tc. o
a. a.
FRESH
ACUTE
CRITERIA
FRESH
CHRONIC
CRITERIA
SALTWATER
ACUTE
CRITERIA
SALTWATER
CHRONIC
CRITERIA
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)
PUBLISHED CRITERA
RECALCULATED VALUES
using IRIS, as of 9/90
WATER &
ORGANISMS
ORGANISMS
ONLY
WATER &
ORGANISMS
ORGANISMS
ONLY
DRINKING
WATER MCL
CRITERIA
FEDERAL
REGISTER
NOTICE
DIELDRIN
60-57-1
Y
Y
2.5
0.0019 0.71 0.0019
0.000071
0 00076
0.00014
0.00014

57 FR 60914
DIETHYL PHTHALATE
84-66-2
Y
N


350,000.
1,800,000.
23,000.
120,000.

57 FR 60913
DIMETHYL PHENOL, 2,4-
105-67-9
Y
N
*2,120.



540.
2,300.

57 FR 60890
DIMETHYL PHTHALATE
131-11-3
Y
N


313,000
2,900,000.



45 FR 79339
DINITROPHENOL, 2,4-
51-28-5
Y
N




70.
14,000.

57 FR 60912
DINITROPH ENOL
25550-58-7
Y
N


70.
14,300.



45 FR 79337
DINITROTOLUENE, 2,4-
121-14-2
Y
Y
*330.
*230. *590. *370.
0.11
9.1



45 FR 79333
DINITRO-O-CRESOL, 2,4-"
534-52-1
Y
N


134
765.



45 FR 79333
DIOXIN (2,3,7,8 - I'CDD)
1746-01-6
Y
Y
* <0.01
* <0.00001
0.000000013
0.000000014



49 FR 5831
DIPHENYLHYDRAZINE, 1,2-
122-66-7
Y
Y
*270.

0.042
0.56
0.041
0.54

57 FR 60913
DI-2-ETHYLHEXYL
PHTHALATE
117-81-7
Y
Y
*2,100.
*160.
15,000.
50,000.
1.8
5.9

45 FR 79339
ENDOSULFAN
115-29-7
N
N
0.22
0.056 0.034 0 0087
74
159.



45 FR 79334
ENDOSULFAN SULFATE
1031-07-8
Y
N




0.93
2.0

57 FR 60915
ENDOSULFAN-ALPHA
959-98-8
Y
N
0.22
0.056 0 034 0 0087


0.93
2.0

57 FR 60914
ENDOSULFA N-BETA
33213-65-9
Y
N
0.22
0.056 0.034 0.0087


0.93
2.0

57 FR 60914
ENDRIN
72-20-8
Y
N
0.18
0.0023 0.037 0.0023
1 0

0.76
0.81

57 FR 60915
ENDRIN ALDEHYDE
7421-93-4
Y
N




0.76
0.81

57 FR 60915
ETHER, BIS
(2-CH LOROETHYL)
111-44-4
Y
Y


0.03
1.36
0.031
1.4

57 FR 60913
ETHER, BIS
(2-CHLOROISOPROPYL)
108-60-1
Y
N


34 7
4360.
1,400.
170,000.

57 FR 60913
ETHER, BIS
(CHLOROMETHYL)
542-88-1
N
Y


0.00(X)038
0 00184



45 FR 79330
ETHYLBENZENE
100-41-4
Y
N
*32,000.
*430
1,400.
3280
3,100.
29,000.
700.
57 FR 60912
ETHYLENE DIBROMIDE
106-93-4
N
Y






0 05
—
FLUORAN'I HENE
206-44-0
Y
N
*3,980
*40. *I6.
42.
54
300.
370.

57 FR 60848
FLUORENE
86-73-7
Y
Y




0.0028
0.031

—
GASES, TO TAL DISSOLVED
—
N
N
NARRATIVE STATEMENT— SEE DOCUMENT





RB
GROSS ALPHA
PARTICLE ACTIVITY

N
Y






15 pCi/L

GUTHION
86-50-0
N
N

0.01 0.01





RB
HALOETHERS
	
N
N
*360.
*122.





45 FR 79334
HALOME'H IANES
—
N
Y
*11,000.
*12,000. *6,400
0.19
15 7



45 FR 79334
HEPTACHLOR
76-44-8
Y
Y
0.52
0 0038 0 053 0.0036
0 00028
0 00029
0.00021
0.00021
0.4
45 FR 79335

-------
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)


l-
Ul




PUBLISHED CRITERA
RECALCULATED VALUES




f"
O






using IRIS, as of 9/90

CRfTERIA


2r 3
z
FRESH
FRESH
SALTWATER
SALTWATER





FEDERAL


o-i
o
tr
ACUTE
CHRONIC
ACUTE
CHRONIC
WATER &
ORGANISMS
WATER &
ORGANISMS
DRINKING
REGISTER

CAS#
12
<
o
CRITERIA
CRITERIA
CRITERIA
CRITERIA
ORGANISMS
ONLY
ORGANISMS
ONLY
WATER MCL
NOTICE
HEPTACHLOR EPOXIDE
1024-57-3
Y
Y
0.52
0.0038
0.053
0.0036


0.00010
0.00011
0.2
45 FR 79335
HEXACHLOROBENZENE
118-74-1
Y
Y
/p/6.0
/ p/3.68


0.00072
0.00074
0.00075
0.00077

57 FR 60912
HEXACHLOROBUTADIENE
87-68-3
Y
Y
*90.
*9.3
*32.

0 45
50.
0.44
50.

45 FR 79335
HEXACHLOROCYCLO-













HEXANE (LINDANE)
58-89-9
Y
Y
2.0
0.08
0.16

0.0186
0.0625
0.019
0.063
02
45 FR 79335
HEXACHLOROCYCLO-













HEXANE—ALPHA
319-84-6
Y
Y




0.0092
0.031
0.0039
0.013

57 FR 60914
HEXACHLOROCYCLO-













HEXANE—BETA
319-85-7
Y
Y




0.0163
0.0547
0.014
0.046

57 FR 60914
HEXACHLOROCYCLO-













HEXANE GAMMA
58-89-9
Y
Y
2.0
0.08
0.16

0.0186
0.0625
0.019
0.063
0.2
57 FR 60914
HEXACHLOROCYCLO-













HEXANE —TECHNICAL
319-86-8
N
Y




0.0123
0.0414



45 FR 79335
HEXACHLOROCYCLO-













PEN TADIENE
77-47-4
Y
N
*7.0
*5.2
*7.0

206.

240.
17,000.

57 FR 60914
HEXACHLOROETHANE
67-72-1
Y
Y
*980.
*540.
*940.

1.9
8.74
1.9
8.9

57 FR 60914
INDENO(l,2,3-CD)PYRENE
193-39-5
Y
Y






0.0028
0.0311

57 FR 60914
IRON
7439-89-6
N
N

1,000.


300.




RB
ISOPHORONE
78-59-1
Y
N
*117,000.

*12,900.

5,200
520,000
8.4
600.

57 FR 60914
LEAD
7439-92-1
Y
N
82.+
3.2+
220.
8.5
50.



/p/5.0
57 FR 60914
MA LATH ION
121-75-5
N
N

0.1

0 1





RB
MANGANESE
7439-96-5
N
N




50.
100



RB
MERCURY
7439-97-6
Y
N
2.4
0.012
2.1
0.025
0.144
0.146
0.14
0.15
2.0
50 FR 30791
METHOXYCHLOR
72-43-5
N
N

0.03

0.03
100



40.
RB
METHYL BROMIDE
74-83-9
Y
N






48.
4,000.

57 FR 60912
ME lT-IYL CHLORIDE
74-87-3
Y
Y









57 FR 60848
METHYLENE CHLORIDE
75-09-2
Y
Y






4.7
1,600.

57 FR 60912
MIREX
2385-85-5
N
N

0.001

0.001





RB
NAPHTHALENE
91-20-3
Y
N
*2,300.
*620.
*2,350.






45 FR 79337
NICKEL
7440-02-0
Y
N
1,400.+
160.+
75.
8.3
13.4
100
610.
4,600.

57 FR 60911
NITRATES
14797-55-8
N
N




10,000



10,OIK).
RB
NITRITE
—
N
N








1,000.
—
NITROBENZENE
98-95-3
Y
N
*27,000.

*6,680

19,800.

17.
1,900.

57 FR 60914
Nl'I-ROPHENOLS
	
Y
N
*230.
*150.
*4,850






45 FR 79337
NITROSAMINES
35576-91-1
Y
Y
*5,850

*3,300,000






45 FR 79337

-------
CAS#
(E
O
IE O
cl a.
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)
PUBLISHED CRITERA
FRESH	FRESH SALTWATER SALTWATER
ACUTE CHRONIC ACUTE CHRONIC WATER & ORGANISMS
CRITERIA CRITERIA CRITERIA CRITERIA ORGANISMS ONLY
RECALCULATED VALUES
using IRIS, as of 9/90
WATER &
ORGANISMS
ORGANISMS
ONLY
CRITERIA
FEDERAL
DRINKING REGISTER
WATER MCL NOTICE
NITROSODIBU'IYLAMINE, N-
924-16-3
N
Y




0.0064
0.587



45 FR 79338
NITROSODIETHYLAMINE, N-
55-18-5
N
Y




0.0008nK/L
0.0012



45 FR 79338
NITROSODIMETHYLAMINE,
N-
62-75-9
Y
Y




0.0014
16.
0.00069
8.1

57 FR 60914
NITROSODIPHENYLAMINE,
N-
86-30-6
Y
Y




4.9
16.1
50
16.

57 FR 60914
NITROSOPYROLIDINE, N-
930-55-2
N
Y




0.016
91.9



45 FR 79338
N-NITROSODI-N-
PROPYLAMINE
621-64-7
Y
Y






0.005
1.4

57 FR 60890
OIL AND GREASE
—
N
N
NARRATIVE STATEMENT — SEE DOCUMENT





RB
OXYGEN DISSOLVED
7782-44-7
N
N
WARMWATER AND COLDWATER CRITERIA MATRIX — SEE DOCUMENT



51 FR 22978
PARATHION
56-38-2
N
N
0.065
0.013







51 FR 43667
PCBs
1336-36-3
Y
Y
2.0
0.014
10.
0 03
0.000079
0 00(X)79


0 5
45 FR 79339
PCB-1016
12674-11-2
Y
Y






0.000044
0.000045

57 FR 60915
PCB-1221
11104-28-2
Y
Y






0.000044
0.000045

57 FR 60915
PCB-1232
11141-16-5
Y
Y






0.000044
0.000045

57 FR 60915
PCB-1242
5346-92-19
Y
Y






0.000044
0.000045

57 FR 60915
PCB-1248
12672-29-6
Y
Y






0.000044
0.000045

57 FR 60915
PCB-1254
11097-69-1
Y
Y






0.000044
0.000045

57 FR 60915
PCB-1260
11096-82-5
Y
Y






0 000044
0.000045

57 FR 60915
PENTACHLOROE1HANE
76-01-7
N
N
*7,240.
*1,100.
*390.
*281.





45 FR 79328
PENTACHLOROBENZENE
608-93-5
N
N




74.
85.



45 FR 79327
PENTACIILOROPHENOL
87-86-5
Y
N
*"20.
***13.
13.
7.9
1010

0.28
8.2
/p/1.0
57 FR 60912
pH
—
N
N

6.5-9

6.5-8.5
5-9




RB
PHENANITIRENE
85-01-8
Y
Y
/ p/30
/ p/6.3
/ p/7.7
/ p/4.6





57 FR 60848
PHENOL
108-95-2
Y-
N
*10,200.
*2,560.
*5,800.

3,500.

21,000.
4,600,000.

57 FR 60912
PHOSPHORUS ELEMENTAL
7723-14-0
N
N



0.1





RB
PHTHALATE ESTERS
—
Y
N










POLYNUCLEAR AROMATIC
HYDROCARBONS

Y
Y


*300.

0 0028
.0311



45 FR 79339
PYRENE
129-00-0
Y
Y






0.0028
0.0311

—
RADIUM 226/228
—
N
Y








5 pCi/L
—
SELENIUM
7782-49-2
Y
N
20.
5.0
300.
71.
10



50.
57 FR 60911
SILVER
7440-22-4
Y
N
4 1 + / p/0.92

0.12
2.3/p/7 2
/p/0.92
50



57 FR 60911

-------
CAS#
a. 3
Qd
DC o
0. £L
111
a
o
z
o
IE
4
o
FRESH
ACUTE
CRITERIA
FRESH
CHRONIC
CRfTERIA
SALTWATER
ACUTE
CRfTERIA
SALTWATER
CHRONIC
CRITERIA
HUMAN HEALTH (10-6 RISK LEVEL FOR CARCINOGENS)
PUBLISHED CRITERA
RECALCULATED VALUES
using IRIS, as of 9/90
WATER &
ORGANISMS
ORGANISMS
ONLY
WATER &
ORGANISMS
ORGANISMS
ONLY
DRINKING
WATER MCL
CRITERIA
FEDERAL
REGISTER
NOTICE
SOLIDS DISSOLVED
AND SALINITY	—	N N	250,000.	RB
SOLIDS SUSPENDED AND
TURBIDITY

N
N
NARRATIVE STATEMENT — SEE DOCUMENT





RB
STYRENE
100-42-5
N
Y








100.
	
SULFIDE-HYDROGEN
SULFIDE
7783-06-4
N
N

2.0

2.0





RB
TEMPERATURE
—
N
N
SPECIES DEPENDENT CRITERIA
— SEE DOCUMENT




RB
TETRACHLOROBENZENE,
1,2,4,5-
95-94-3
N
N




38.
48.



45 FR 79327
TETRACHLOROETHANE,
1,1,2,2-
79-34-5
Y
Y

*2,400.
*9,020.

0.17
10.7
0.17
11.

57 FR 60912
TETRACHLOROETHANES
25322-20-7
Y
N
*9,320








45 FR 79328
TETRACHLOROETHYLENE
127-18-4
Y
Y
*5,280
*840.
*10,200.
*450.
0.8
8.85


5.0
45 FR 79340
TETRACHLOROPHENOL,
2,3,5,6-
935-95-5
N
N


*440.






45 FR 79329
THALLIUM
7440-28-0
Y
N
*1,400.
*40.
*2,130.

13.
48.
1.7
6.3

57 FR 60913
TOLUENE
108-88-3
Y
N
*17,500.

*6,300.
*5,000.
14,300.
424,000.
6,800.
200,000.
1,000.
57 FR 60912
TOXAPHENE
8001-35-2
Y
Y
0.73
0.0002
0.21
0.0002
0.00071
0.00073
0.00073
0.00075
/p/5.0
57 FR 60915
TRICHLORINATED ETHANES
25323-89-1
Y
Y
*18,000.








45 FR 79328
TRICHLOROETHANE, 1,1,1-
71-55-6
Y
N


*31,200.

18,400.
1,030,000.


200.
57 FR 60848
TRICHLOROETHANE, 1,1,2-
79-00-5
Y
Y

*9,400.


0.6
41.8
0.60
42.

57 FR 60912
TRICHLOROETH YLEN E
79-01-6
Y
Y
*45,000.
*21,900.
*2,000.

2.7
80.7
2.7
81.
5.0
45 FR 79341
TRICHLOROPHENOL, 2,4,5-
95-95-4
N
N
/p/100
/p/63
/p/240
/p/11
2,600.




45 FR 79329
TRICHLOROPHENOL, 2,4,6-
88-06-2
Y
Y

*970.


1.2
3.6
2.1
6.5

57 FR 60912
VINYL CHLORIDE
75-01-4
Y
Y




2.0
525.


2.0
45 FR 79341
XYLENES
—
N
N








10,000.
—
ZINC	7440-66-6 Y N 120. + 110 +	95.	86.	52 FR 6214
+ = Hardness dependent criteria (100 mg/L CaC03 used)	RB = Red Book
* = Insufficient data to develop criteria. Value presented is the L.O.E.L. (Lowest Observed Effect Level) /p/= Proposed criterion
" = The preferred chemical name for 2,4 Dinitro-o-cresol listed in 45 FR 79333 is 4,6 - Dinitro-o-cresol	Y = Yes
*** = pH dependent criteria (7.8 pH used)	N = No
MCL= Maximum contaminant level (only for listed chemicals)	Marine criteria for lead reflect values updated after 1984.
Note: This chart is for general information. Please use criteria documents or detailed summaries in Quality Criteria for Water 1992 for regulatory purposes.

-------
CONTENTS
PREFACE	iii
Water Quality Criteria Summary Concentrations 	v
ACENAPHTHENE 	^	1
ACROLEIN		3
ACRYLONITRILE	5
AESTHETIC QUALITIES* 	7
ALDRIN 	9
ALKALINITY	11
ALUMINUM 	13
AMMONIA 	15
ANTIMONY 		25
ARSENIC		27
ASBESTOS 	31
BACTERIA 	33
BARIUM	35
BENZENE 	37
BENZIDINE 	39
BERYLLIUM 	41
BENZENE HEXACHLORIDE (BHC)	43
(See also: Hexachlorocyclohexane)
Bis(2-chloroethyl) ether-see: Chloroalkyl Ethers
Bis(2-chloroisopropyl) ether-see: Chloroalkyl Ethers
Bis(chloromethyl) ether-see: Chloroalkyl Ethers
BORON* 	45
CADMIUM	47
CARBON TETRACHLORIDE	49
CHLORDANE 	51
CHLORIDE	53
CHLORINATED BENZENES	55
CHLORINATED ETHANES 	57
CHLORINATED NAPHTHALENES	59
CHLORINATED PHENOLS 	61
CHLORINE 	65
CHLOROALKYL ETHERS 	67
Chlorobenzene-see: Chlorinated Benzenes
CHLOROFORM	69
•Entry not listed on 1990 Water Quality Chart.
xiii

-------
2-CHLOROPHENOL	71
4-Chlorophenol-see: Chlorinated Phenols
4-Chlorophenol, 3-methyl-see: Chlorinated Phenols
CHLOROPHENOXY HERBICIDES 	73
CHLORPYRIFOS 	75
CHROMIUM 		77
COLOR 						81
COPPER 				83
CYANIDE	85
DDT AND METABOLITES 			87
DEMETON 	89
Dibutyl phthalate-sce: Phthalate Esters
DICHLOROBENZENES 	91
DICHLOROBENZIDINE 	93
1,2-Dichloroethane-see: Chlorinated Ethanes
DICHLOROETHYLENES	95
2,4-DICHLOROPHENOL . 		.97
DICHLOROPROPANE/DICHLOROPROPENE	99
DIELDRIN 	101
Diethyl phthalate-see: Phthalate Esters
2,4-DIMETHYLPHEN OL	103
Dimethyl phthlate-see: Phthalate Esters
Dinitrophenol-see: Nitrophenols
DINITROTOLUENE 			105
Dinitro-o-cresol-see: Nitrophenols
Dioxin-see: Tetrachlorodibenzo-p-dioxin
DIPHENYLHYDRAZINE 	107
DI-2-ETHYLHEXYL PHTHALATE			109
(See also: Phthalate Esters)
DISSOLVED OXYGEN 	Ill
DISSOLVED SOLIDS AND SALINITY 		117
ENDOSULFAN 	119
ENDRIN 	121
Ether, bis(2-chloroethyl)-see: Chloroalkyl Ethers
Ether, bis(2-chloroisopropyl)-see: Chloroalkyl Ethers
Ether, bis(chloromethyl)-see: Chloroalkyl Ethers
ETHYLBENZENE 	123
FLUORANTHENE	125
GASES, TOTAL DISSOLVED 	127
GUTHION 	129
HALOETHERS 	131
xiv

-------
HALOMETHANES 	133
HARDNESS* 	135
HEPTACHLOR 	137
HEXACHLOROBENZENE	139
See also: Chlorinated Benzenes
HEXACHLOROBUTADIENE	141
HEXACHLOROCYCLOHEXANE	143
(Note: gamma-hexachlorocyclohexane = Lindane-see also: BHQ
HEX ACHLOROCYCLOPENTADIENE 	145
Hexachloroethane-see: Chlorinated Ethanes
Hydrogen sulfide-see: Sulfide-HiS
IRON 	147
ISOPHORONE 	151
LEAD 	153
MALATHION	155
Lindane-see; Hexachlorocyclohexane
MANGANESE 	157
MERCURY 	159
METHOXYCHLOR 	163
MIREX	165
NAPHTHALENE 	167
NICKEL 	179
NITRATES /NITRITES	171
NITROBENZENE 	173
NITROPHENOLS 	175
NITROSAMINES 	177
N-Nitrosodibutylamine-see: Nitrosamines
N-Nitrosodiethylamine-see: Nitrosamines
N-Nitrosodimethylamine-see: Nitrosamines
N-Nitrosodiphenylamine-see: Nitrosamines
N-Nitrosodipropylamine-s^e: Nitrosamines
N-Nitrosopyrolidine-see: Nitrosamines
OIL AND GREASE 	179
Oxygen, Dissolved-see: Dissolved Oxygen
PARATHION	183
PCBs-see: Polychlorinated Biphenyls
Pentachloroethane-see: Chlorinated Ethanes
Pentachlorobenzene-see: Chlorinated Benzenes
PENTACHLOROPHENOL (PCP) 	185
pH	187
PHENOL	189
XV

-------
PHOSPHORUS 	191
PHTHALATE ESTERS 	195
POLYCHLORINATED BIPHENYLS (PCBs)	197
POLYNUCLEAR AROMATIC HYDROCARBONS	199
SELENIUM 	201
SILVER	203
Solids, Dissolved and Salinity-see: Dissolved Solids
SOLIDS (SUSPENDED, SETTLEABLE) AND TURBIDITY 	205
SULFIDE - HYDROGEN SULFIDE	207
TAINTING SUBSTANCES*	209
TEMPERATURE	211
Tetrachlorobenzenes-see: Chlorinated Benzenes
2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN
(TCDD) (DIOXIN) 	.221
Tetrachloroethanes-see: Chlorinated Ethanes
TETRACHLOROETHYLENE 	.223
Tetrachlorophenol-see: Chlorinated Phenols
THALLIUM	225
TOLUENE 	227
TOXAPHENE	229
Trichlorinated Ethanes-see: Chlorinated Ethanes
TRICHLOROETHYLENE	231
Trichlorophenols-see: Chlorinated Phenols
VINYL CHLORIDE 	233
ZINC	235
Appendix A: Derivation of the 1985 Aquatic Life Criteria	237
Appendix B: Derivation of the 1980 Aquatic Life Criteria	255
Appendix C: Derivation of the 1980 Human Health Criteria	267
Appendix D: Derivation of 1976 Philosophy of Aquatic Life Criteria . 283
Appendix E: Bioconcentration Factors 	287
Appendix F: Water Quality Criteria Documents 	291

-------
ACENAPHTHENE
83-32-9
CRITERIA
Aquatic Life The available data for acenaphthene indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 1,700 ng/Land would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of
acenaphthene to sensitive freshwater aquatic animals, but toxicity to fresh-
water algae occurs at concentrations as low as 520 ng/L.
The available data for acenaphthene indicate that acute and chronic
toxicity to saltwater aquatic life occurs at concentrations as low as 970 and
710 ng/L, respectively, and would occur at lower concentrations among
species that are more sensitive than those tested. Toxicity to algae occurs at
concentrations as low as 500 |^g/L.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 RR. 60890). Recalcul-
ated IRIS values for acenaphthene are 1,200 M-g/L for ingestion of
contaminated water and organisms and 2,700 fig/L for ingestion of con-
taminated aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22,1993)
See Appendix C for Human Health Methodology.
1

-------
2

-------
ACROLEIN
107-02-8
CRITERIA
Aquatic Life The available data for acrolein indicate that acute and chronic toxicity to
freshwater aquatic life occurs at concentrations as low as 68 and 21 ng/L,
respectively, and would occur at lower concentrations among species that
are more sensitive than those tested.
The available data for acrolein indicate that acute toxicity to saltwater
aquatic life occurs at concentrations as low as 55 ng/L and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning the chronic toxicity of acrolein to
sensitive saltwater aquatic life.
Human Health For the protection of human health from the toxic properties of acrolein in-
gested through water and contaminated aquatic organisms, the ambient
water criterion is 320 jig/L.
For the protection of human health from the toxic properties of acro-
lein ingested through contaminated aquatic organisms alone, the ambient
water criterion is 780 ng/L.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
3

-------
ACRYLONITRILE
107-13-1
CRITERIA
Aquatic Life The available data for acrylonitrile indicate that acute toxicity to freshwa-
ter aquatic life occurs at concentrations as low as 7,550 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No definitive data are available concerning the chronic toxic-
ity of acrylonitrile to sensitive freshwater aquatic life, but mortality occurs
at concentrations as low as 2,600 ng/L, with a fish species exposed for 30
days.
Only one saltwater species has been tested with acrylonitrile, therefore
no statement can be made concerning acute or chronic toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects resulting from exposure to acrylonitrile through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at lO"3,10"6, and 10"7.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92. Recal-
culated IRIS values for acrylonitrile are 0.059 ng/L for ingestion of
contaminated water and organisms and 0.66 ng/L for ingestion of contam-
inated aquatic organisms only. IRIS values are based on a 10"6 risk level for
carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
5

-------
AESTHETIC QUALITIES
CRITERIA 	
All waters free from substances attributable to wastewater or other dis-
charges that
1.	settle to form objectionable deposits;
2.	float as debris, scum, oil, or other matter to form nuisances;
3.	produce objectionable color, odor, taste, or turbidity;
4.	injure or are toxic or produce adverse physiological responses in
humans, animals, or plants; and
5.	produce undesirable or nuisance aquatic life.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
7

-------
*ALDRIN
309-00-2
CRITERIA
Aquatic Life Not to exceed 3.0 ng/L in fresh water or 1.3 ng/L in salt water.
For freshwater aquatic life, the concentration of aldrin should not ex-
ceed 3.0 jxg/L at any time. No data are available concerning the chronic
toxicity of aldrin to sensitive freshwater aquatic life.
For saltwater aquatic life, the concentration of aldrin should not exceed
1.3 ng/L at any time. No data are available concerning the chronic toxicity
of aldrin to sensitive saltwater aquatic life.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to aldrin through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tion should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in incremental increase of cancer risk
over the lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848,
December 22, 1992). Recalculated IRIS values for aldrin are 0.00013 ng/L
for ingestion of contaminated water and organisms and 0.00014 fig/L for
ingestion of contaminated aquatic organisms only. IRIS values are based
on a 10"6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
•Indicates suspended, canceled, or restricted by U.S. EPA Office of Pesticides and Toxic
Substances
9

-------
10

-------
B
ALKALINITY
CRITERIA 	
20 mg/L, or more as CaCO^, for freshwater aquatic life except where natu-
ral concentrations are less!
Introduction Expressed commonly as milligrams per liter of calcium carbonate, alkalin-
ity is the sum total of components in the water that tend to elevate the pH
of the water above a value of about 4.5. It is measured by titration with
standardized acid to a pH value of about 4.5. Alkalinity, therefore, is a
measure of water's buffering capacity, and since pH has a direct effect on
organisms as well as an indirect effect on the toxicity of certain other pollu-
tants in the water, the buffering capacity is important to water quality.
Examples of commonly occurring materials in natural waters that increase
the alkalinity are carbonates, bicarbonates, phosphates, and hydroxides.
Rationale	The alkalinity of water used for municipal water supplies is important
because it affects the amount of chemicals that need to be added to ac-
complish coagulation, softening, and control of corrosion in distribution
systems. The alkalinity of water assists in the neutralization of excess
acid produced during the addition of such materials as aluminum sul-
fate during chemical coagulation. Waters having sufficient alkalinity do
not have to be supplemented with artificially added materials to in-
crease the alkalinity. Alkalinity resulting from naturally occurring
materials such as carbonate and bicarbonate is not considered a health
hazard in drinking water supplies, per se, and naturally occurring max-
imum levels up to approximately 400 mg/L as calcium carbonate are
not considered a problem to human health.
Alkalinity is important for fish and other aquatic life in freshwater sys-
tems because it buffers pH changes that occur naturally as a result of
photosynthetic activity of the chlorophyll-bearing vegetation. Components
of alkalinity such as carbonate and bicarbonate will complex some toxic
heavy metals and reduce their toxicity markedly. For these reasons, in 1968
the National Technical Advisory Committee recommended a minimum al-
kalinity of 20 mg/L. The subsequent 1974 National Academy of Sciences
(NAS) report recommended that natural alkalinity not be reduced by more
than 25 percent but did not place an absolute minimal value for it. The use
of the 25 percent reduction avoids the problem of establishing standards
on waters where natural alkalinity is at or below 20 mg/L. For such wa-
ters, alkalinity should not be further reduced.
The NAS Report recommends that adequate amounts of alkalinity be
maintained to buffer the pH within tolerable limits for marine waters. It
has been noted as a correlation that productive waterfowl habitats are
above 25 mg/L with higher alkalinities resulting in better waterfowl habi-
tats.
Excessive alkalinity can cause problems for swimmers by altering the
pH of the lacrimal fluid around the eye, causing irritation.
11

-------
For industrial water supplies high alkalinity can be damaging to indus-
tries involved in food production, especially those in which acidity accounts
for flavor and stability, such as the carbonated beverages. In other instances,
alkalinity is desirable because water with a high alkalinity is much less corro-
sive.
A brief summary of maximum alkalinities accepted as a source of raw
water by industry is included in Table 1. The concentrations listed in the
table are for water prior to treatment and thus are only desirable ranges
and not critical ranges for industrial use.
Table 1.*—Maximum alkalinity in waters used as a source of supply prior to
treatment.
INDUSTRY
ALKALINH Y
MG/L AS C'ACO,
Stcain generation boiler makeup 	
350
Steam generation cooling	
500
Textile mill products	 	
50-200
Paper and allied products	 	
75-150
Chemical and allied products	
500
Petroleum refining	 	
500
Pnmarv metals industries 	
200
Food canninc industries	
300
Bottled and canned soft drinks
85
Source: National Academy of Sciences (1974).
The effect of alkalinity in water used for irrigation may be important in
some instances because it may indirectly increase the relative proportion of
sodium in soil water. As an example, when bicarbonate concentrations are
high, calcium and magnesium ions that are in solution precipitate as car-
bonates in the soil water as the water becomes more concentrated through
evaporation and transpiration. As the calcium and magnesium ions de-
crease in concentration, the percentage of sodium increases and results in
soil and plant damage. Alkalinity may also lead to chlorosis in plants be-
cause it causes the iron to precipitate as a hydroxide. Hydroxyl ions react
with available iron in the soil water and make the iron unavailable to
plants. Such deficiencies induce chlorosis and further plant damage. Usu-
ally alkalinity must exceed 600 mg/L before such affects are noticed,
however.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
12

-------
ALUMINUM
7429-90-5
When the pH is between 6.5 and 9.0, four-day average concentration of
aluminum should not exceed 87 ng/L. One-hour average concentration of
aluminum should not exceed 750 ng/L for freshwater aquatic life.
Acute tests have been conducted on aluminum at pH betwen 6.5 and 9.0
with freshwater species in 14 genera. In many tests, less than 50 percent of
the organisms were affected at the highest concentration tests. Both
ceriodaphnids and brook trout were affected at concentrations below 4,000
Hg/L, whereas some other fish and invertebrate species were not affected
by 45,000 ng/L. Some researchers found that the acute toxicity of alumi-
num increased with pH, whereas others found the opposite to be true.
Three studies have been conducted on the chronic toxicity of alumi-
num to aquatic animals. The chronic values for Daphnia magna,
Ceriodaphnia dubia, and the fathead minnow were 742.2, 1,908, and 3,288
Hg/L, respectively. The diatom, Cyclotella meneghiniana, and the green alga,
Selenastrum capricornutum, were affected by concentrations of aluminum in
the range of 400 to 900 ng/L. Bioconcentration factors from 50 to 231 were
obtained in tests with young brook trout. At a pH of 6.5 to 6.6, 169 ng/L
caused a 24 percent reduction in the growth of young brook trout and 174
(xg/L killed 58 percent of the exposed striped bass.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably, when the pH is between 6.5 and 9.0, if the
four-day average concentration of aluminum does not exceed 87 ng/L
more than once every three years on the average and if the one-hour aver-
age concentration does not exeed 750 ng/L more than once every three
years on the average.
(53 F.R. 33178, August 30,1988)
See Appendix A for Aquatic Life Methodology.
13
CRITERIA
Aquatic Life
Summary

-------
14

-------
AMMONIA
7664-41-7
SUMMARY
Fresh Water All concentrations used herein are expressed as un-ionized ammonia
(NH3) because NH3, not the ammonium ion (NH4+ ), has been demon-
stffteB to be the principal toxic form of ammonia. The data used in
deriving criteria are predominantly from flow through tests that measured
ammonia concentrations.
Ammonia was reported to be acutely toxic to freshwater organisms at
concentrations (uncorrected for pH) ranging from 0.53 to 22.8 mg/L NH3
for 19 invertebrate species representing 14 families and 16 genera and from
0.083 to 4.60 mg/L NH3 for 29 fish species from 9 families and 18 genera.
Among fish species, reported 96-hour LC50 ranged from 0.083 to 1.09
mg/L for salmonids and from 0.14 to 4.60 mg/L NH3 for nonsalmonids.
Reported data from chronic tests on ammonia with two freshwater inverte-
brate species, both daphnids, showed effects at concentrations
(uncorrected for pH) ranging from 0.304 to 1.2 mg/L NH3 and with 9
freshwater fish species, from 5 families and 7 genera, ranging from 0.0017
to 0.612 mg/L NH3.
Concentrations of ammonia are acutely toxic to fishes and can cause
loss of equilibrium, hyperexcitability, increased breathing, cardiac output
and oxygen uptake, and, in extreme cases, convulsions, coma, and death.
At lower concentrations, ammonia has many effects on fishes, including a
reduction in hatching success, reduction in growth rate and morphological
development, and pathologic changes in tissues of gills, livers, and kid-
neys.
Several factors have been shown to modify acute NH3 toxicity in fresh
water. Some factors alter the concentration of un-ionized ammonia in the
water by affecting the aqueous ammonia equilibrium, and some factors af-
fect the toxicity of un-ionized ammonia itself, either by ameliorating or
exacerbating the effects of ammonia. Factors that have been shown to af-
fect ammonia toxicity include dissolved oxygen concentration,
temperature, pH, previous acclimation to ammonia, fluctuating or inter-
mittent exposures, carbon dioxide concentration, salinity, and the presence
of other toxicants.
The most well-studied of these factors is pH: the acute toxicity of NH3
has been shown to increase as pH decreases. Sufficient data exist from tox-
icity tests conducted at different pH values to formulate a mathematical
expression to describe pH-dependent, acute NH3 toxicity. The very limited
amount of data regarding effects of pH on chronic NH3 toxicity also indi-
cates increasing NH3 toxicity with decreasing pH, but the data are
insufficient to derive a broadly applicable toxicity/pH relationship. Data
on temperature effects on acute NH3 toxicity are limited and somewhat
variable, but indications are that NH3 toxicity to fish is greater as tempera-
ture decreases. No information is available regarding temperature effects
on chronic NH3 toxicity.
15

-------
Examination of pH and temperature-corrected acute NH3 toxicity val-
ues among species and genera of freshwater organisms showed that
invertebrates are generally more tolerant than fishes, a notable exception
being the fingernail clam. No clear trend exists among groups of fish; the
several most sensitive tested species and genera include representatives
from diverse families (Salmonidae, Cyprinidae, Percidae, and Cen-
trarchidae). Available chronic toxicity data for freshwater organisms also
indicate invertebrates (cladocerans, one insect species) to be more tolerant
than fishes, again with the exception of the fingernail clam. When cor-
rected for the presumed effects of temperature and pH, chronic toxicity
values also show no clear trend among groups of fish, the most sensitive
species including representatives from five families (Salmonidae,
Cyprinidae, Ictaluridae, Centrarchidae, and Catostomidae) and having
chronic values ranging by not much more than a factor or two. The range
of acute-chronic ratios for 10 species from 6 families was 3 to 43; acute-
chronic ratios were higher for the species having chronic tolerance below
the median.
Available data indicate that differences in sensitivities between
warm water and coldwater families of aquatic organisms are inadequate to
warrant discrimination in the national ammonia criterion between bodies
of water with "warmwater" and "coldwater" fishes; rather, effects of or-
ganism sensitivities on the criterion are most appropriately handled by
site-specific criteria derivation procedures.
Data for concentrations of NH3 toxic to freshwater phytoplankton and
vascular plants, although limited, indicate that freshwater plant species are
appreciably more tolerant to NH3 than are invertebrates or fishes. The am-
monia criterion appropriate for the protection of aquatic animals,
therefore, will probably sufficiently protect plant life.
Salt Water	In aqueous solutions, the ammonium ion dissociates to un-ionized ammo-
nia and the hydrogen ion. The equilibrium equation can be written
Equation 1
H2O + NH4+ o nh3+h3o+
The total ammonia concentration is the sum of NH3 and NH4+.
The toxicity of aqueous ammonia solutions to aquatic organisms is pri-
marily attributable to the un-ionized form, the ammonium ion being less
toxic. It is necessary, therefore, to know the percentage of total ammonia in
the un-ionized form in order to establish the corresponding total ammonia
concentration toxic to aquatic life. The percentage of un-ionized ammonia
(UIA) can be calculated from the solution pH and pKa, the negative log of
stoichiometric dissociation,
Equation 2
% UIA= 100 [1 + 10(pKa-pH)]-1
The stoichiometric dissociation constant is defined
Equation 3
[NHaUhT]
Ka "	t	
[NHa+ ]
where the brackets represent molal concentrations. Ka is a function of the
temperature and ionic strength of the solution.
16

-------
Whitfield (1974) developed theoretical models to determine the pKa of
the ammonium ion in seawater. He combined his models with the infinite
dilution data of Bates and Pinching (1949) to define general equations for
the PKa of ammonium ion as a function of salinity and temperature.
Whitfield's models allow reasonable approximations of the percent of
un-ionized ammonia in sea water and have been substantiated experimen-
tally. Hampson's (1977) program for Whitfield's full seawater model has
been used to calculate the un-ionized ammonia fraction of measured total
ammonia concentrations in toxicity studies conducted by EPA and also in
deriving most other acute and chronic ammonia values that contribute to
the criteria. The equations for this model are
Equation 4
% UIA = 100 [1 + 10 (X + 0.0324 (298-T) + 0.0415 P/T- pH]'1
where
P = 1 AIM for all toxicity testing reported to date
T = temperature (K)
X = pksa or the stoichiometric acid hydrolysis constant of ammonium
ions in saline water based on I,
Equation 5
I = 19.9273 S (1000-1.005109 S)"1
where
I = molal ionic strength of the sea water
S = salinity (g/kg)
The Hampson program calculates the value for I for the test salinity
(Eq. 5), finds the corresponding pksa, then calculates % UIA (Eq. 4).
The major factors influencing the degree of ammonia dissociation are
pH and temperature. Both correlate positively with un-ionized ammonia.
Salinity, the least influential of the three water quality factors that control
the fraction of un-ionized ammonia, is inversely correlated.
NATIONAL CRITERIA
Fresh Water The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important species is
very sensitive, freshwater aquatic organisms and their uses should not be af-
fected unacceptably if
(1) the one-hour* average concentration of un-ionized ammonia (in
mg/LNH3) (see Tables 1 and 2) does not exceed more often than
once every three years on the average the numerical value given by
0.52/FT/FPH/2
where:
FT = 10°-O3(2O"TCAP) ; TCAPs Ts 30
10o.03(20-T)	; 0 s T s TCAP
*An averaging period of one hour may not be appropriate if excursions of
concentrations to greater than 1.5 times the average occur during the hour; in such
cases, a shorter averaging period may be needed.
17

-------
Table 1.—One-hour average concentrations for ammonia with salmonids or other

sensitive coldwater species present.*







Temperature fC)



PH
0
5
10
15
20
25
30

Un-ionized Ammonia (mg/L NH3)
6.50
0.0091
0.0129
0.0182
0.026
0.036
0.036
0.036
6.75
0.0149
0.021
0.030
0.042
0.059
0.059
0.059
7.00
0.023
0.033
0.046
0.066
0.093
0.093
0.093
7.25
0.034
0.048
0.068
0.095
0.135
0.135
0.135
7.50
0.045
0.064
0.091
0.128
0.181
0.181
0.181
7.75
0.056
0.080
0.113
0.159
0.22
0.22
0.22
8.00
0.065
0.092
0.130
0.184
0.26
0.26
0.26
8.25
0.065
0.092
0.130
0.184
0.26
0.26
0.26
8.50
0.065
0.092
0.130
0.184
0.26
0.26
0.26
8.75
0.065
0.092
0 130
0.184
0.26
0.26
0.26
9 00
0.065
0.092
0.130
0.184
0.26
0.26
0.26


Total Ammonia (
mg/L NH3)


6.50
35
33
31
30
29
20
14.3
6.75
32
30
28
27
27
18.6
13.2
7.00
28
26
25
24
23
16.4
11.6
7.25
23
22
20
19.7
19.2
13 4
9.5
7.50
17.4
16.3
15.5
14.9
14.6
10.2
7.3
7.75
12.2
11.4
10.9
10.5
10.3
7.2
5.2
8.00
8 0
7.5
7.1
6.9
6.8
4.8
3.5
8.25
4.5
4.2
4.1
4.0
3.9
2.8
2.1
8.50
2.6
24
2.3
2.3
2.3
1.71
1.28
8.75
1 47
1 40
1.37
1.38
1.42
1.07
0.83
9.00
0.86
0.83
0.83
0.86
0.91
0.72
0.58
*To convert these values to mg/L N, multiply by 0.822.
FPH = 1	8 s pH s 9
1 + 107'4 ~ pH
1.25	6-5 s pH s 8
TCAP = 20°C; salmonids or other sensitive coldwater species present
TCAP = 25°C; salmonids and other sensitive coldwater species absent
(2) the four-day average concentration of un-ionized ammonia (in
mg/L NH3) (see Tables 3 and 4) does not exceed, more often than
once every three years on the average, the average** numerical
value given by 0.80/FT/FPH/RATIO, where FT and FPH are as
above and
RATIO =13.5	7.7 £ pH £ 9
107.7 -pH
= 20 • — nU	6.5 £ pH £ 7.7
1+107.4 -pH
TCAP = 15°C; salmonids or other sensitive coldwater species present
TCAP = 20°C; salmonids and other sensitive coldwater species absent
"Because these formulas are nonlinear in pH and temperature, the criterion should be
the average of separate evaluations of the formulas reflective of the fluctuations of
flow, pH, and temperature within the averaging period; it is not appropriate in
general to simply apply the formula to average pH, temperature, and flow.

-------
Table 2.—One-hour average concentrations for ammonia with saimonids or other

sensitive coldwater species absent.
*







Temperature (*C)



PH
0
5
10
15
20
25
30

Un-ionized Ammonia (mg/L NH3)
6.50
0.0091
0.0129
0.0182
0.026
0.036
0.051
0.051
6.75
0.0149
0.021
0.030
0.042
0.059
0.084
0.084
7.00
0.023
0.033
0.046
0.066
0.093
0.131
0.131
7.25
0.034
0.048
0.068
0.095
0.135
0.190
0.190
7.50
0.045
0.064
0.091
0.128
0.181
0.26
0.26
7.75
0.056
0.080
0.113
0.159
0.22
0.32
0.32
8.00
0.065
0.092
0.130
0.184
0.26
0.37
0.37
8.25
0.065
0.092
0.130
0.184
0.26
0.37
0.37
8.50
0.065
0.092
0.130
0.184
0.26
0.37
0.37
8.75
0.065
0.092
0.130
0 184
0.26
0.37
0.37
9.00
0.065
0.092
0.130
0.184
0.26
0.37
0.37

Total Ammonia (mg/L NH3)
6.50
35
33
31
30
29
29
20
6.75
32
30
28
27
27
26
18.6
7.00
28
26
25
24
23
23
16.4
7.25
23
22
20
19.7
19.2
19.0
13.5
7.50
17.4
16.3
15.5
14.9
14.6
14.5
10.3
7.75
12.2
11.4
10.9
10.5
10.3"
10.2
7.3
8.00
8.0
7.5
7.1
6.9
6.8
6.8
4.9
8.25
4.5
4.2
4.1
4.0
3.9
4.0
2.9
8 50
2.6
2.4
2.3
2.3
2.3
2.4
1 81
8.75
1.47
1 40
1.37
1.38
1 42
1.52
1.18
9.00
0.86
0.83
0 83
0.86
0.91
1.01
0.82
*To convert these values to mg/L N, multiply by 0.822.
The extremes for temperature (0°C and 30°C) and pH (6.5, 9) given in
the above formulas are absolute. It is not permissible with current data to
conduct any extrapolations beyond these limits. In particular, there is rea-
son to believe that appropriate criteria at pH > 9 will be lower than the
plateau between pH 8 and 9 given above.
Limited data exists on the effect of temperature on chronic toxicity.
EPA will be conducting additional research on the effects of temperature
on ammonia toxicity to fill perceived data gaps. Because of this uncer-
tainty, additional site-specific information should be developed before
these criteria are used in wasteload allocation modeling. For example, the
chronic criteria tabulated for sites lacking salmonids are less certain at
temperatures much below 20°C than those tabulated at temperatures near
20°C. Where the treatment levels needed to meet these criteria below 20°C
may be substantial, use of site-specific criteria is strongly suggested. De-
velopment of such criteria should be based upon site-specific toxicity
tests.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average amount of time for an unstressed
system to recover from a pollution event in which exposure to ammonia
exceeds the criterion. A stressed system — for example, one in which sev-
eral outfalls occur in a limited area — would be expected to require more
19

-------
Table 3.—Four-day average concentrations for ammonia with salmonids or other
sensitive coidwater species present.*



Temperature fC)



pH
0
5
10
15
20
25
30

Un-ionized Ammonia (mg/L NH3)
6.50
0.0008
0.0011
0.0016
0.0022
0.0022
0.0022
0.0022
6.75
0.0014
0.0020
0.0028
0.0039
0.0039
0.0039
0.0039
7.00
0.0025
0.0035
0.0049
0.0070
0.0070
0.0070
0.0070
7.25
0.0044
0.0062
0.0088
0.0124
0.0124
0.0124
0.0124
7.50
0.0018
0.0111
0.0156
0.022
0.022
0.022
0.022
7.75
0.0129
0.0182
0.026
0.036
0.036
0.036
0.036
8.00
0.0149
0.021
0.030
0.042
0.042
0.042
0.042
8.25
0.0149
0.021
0.030
0.042
0.042
0.042
0.042
8.50
0.0149
0.021
0.030
0.042
0.042
0.042
0.042
8.75
0.0149
0.021
0.030
0.042
0.042
0.042
0.042
9 00
0.0149
0.021
0.030
0.042
0 042
0.042
0 042

Total Ammonia (mg/L NH3)
6.50
3.0
2.8
27
2.5
1 76
1.23
0.87
6 75
3.0
2.8
2.7
2.6
1 76
1.23
0.87
7.00
3 0
2.8
2 7
2.6
1.76
1.23
0.87
7.25
3.0
2.8
2.7
2.6
1.77
1.24
0.88
7.50
3.0
2.8
2.7
2.6
1.78
1.25
0.89
7.75
2.8
2.6
2.5
24
1.66
1.17
0.84
8.00
1.82
1.70
1.62
1.57
1.10
0.78
0.56
8.25
1.03
0.97
0.93
0.90
0.64
0.46
0.33
8.50
0.58
0.55
0.53
0.53
0.38
0.28
0.21
8.75
0.34
0.32
0.31
0.31
0.23
0.173
0.135
9.00
0.195
0.189
0.189
0.195
0.148
0.116
0.094
*To convert these values to mg/L N, multiply by 0.822.
time for recovery. The resilience of ecosystems and their ability to recover
differ greatly, however, and site-specific criteria may be established if ade-
quate justification is provided.
The use of criteria in designing waste treatment facilities requires select-
ing an appropriate wasteload allocation model. Dynamic models are
preferred for the application of these criteria. Limited data or other factors
may make their use impractical, in which case one should rely on a steady-
state model. The Agency recommends the interim use of 1Q5 or 1Q10 for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for the
Criterion Continuous Concentration design flow in steady-state models for
unstressed and stressed systems, respectively. The Agency acknowledges
that the Criterion Continuous Concentration stream flow averaging period
used for steady-state wasteload allocation modeling may be as long as 30
days in situations involving POTWs designed to remove ammonia where
limited variability of effluent pollutant concentration and resultant concen-
trations in receiving waters can be demonstrated. In cases where low
variability can be demonstrated, longer averaging periods for the ammonia
Criterion Continuous Concentration (e.g., 30-day averaging periods) would
be acceptable because the magnitude and duration of exceedences above
the Criterion Continuous Concentration would be sufficiently limited.

-------
Table 4.—Four-day average concentrations for ammonia with salmonids or other
sensitive coldwater species absent* **



Temperature f C)



PH
0
5
10
15
20
25
30

Un-ionized Ammonia (mg/L NH3)
6.50
0.0008
0.0011
0.0016
0.0022
0.0031
0.0031
0.0031
6.75
0.0014
0.0020
0.0028
0.0039
0.0055
0.0055
0.0055
7.00
0.0025
0.0035
0.0049
0.0070
0.0099
0.0099
0.0099
7.25
0.0044
0.0062
0.0088
0.0124
0.0175
0.0175
0.0175
7.50
0.0078
0.0111
0.0156
0.022
0.031
0.031
0.031
7.75
0.0129
0.0182
0.026
0.036
0.051
0.051
0.051
8.00
0.0149
0.021
0.030
0.042
0.059
0.059
0.059
8.25
0.0149
0.021
0.030
0.042
0.059
0.059
0.059
8.50
0.0149
0.021
0.030
0.042
0.059
0.059
0.059
8.75
0.0149
0.021
0.030
0.042
0.059
0.059
0.059
9.00
0.0149
0.021
0.030
0.042
0.059
0.059
0.059

Total Ammonia (mg/L NH3)
6.50
3.0
2.8
2.7
2.5
2.5
1.73
1.23
6.75
3.0
2.8
2.7
2.6
2.5
1.74
1.23
7.00
3.0
2.8
2.7
2.6
2.5
1.74
1.23
7.25
3.0
2.8
2.7
2.6
2.5
1.75
1.24
7.50
3.0
2.8
2.7
2.6
2.5
1.76
1.25
7.75
2.8
2.6
2.5
2.4
2.3
1.65
1.18
8.00
1.82
1.70
1.62
1.57
1.55
1.10
0.79
8.25
1.03
0.97
0.93
0.90
0.90
0.64
0.47
8.50
0.58
0.55
0.53
0.53
0.53
0.39
0.29
8.75
0.34
0.32
0.31
0.31
0.32
0.24
0.190
9.00
0.195
0.189
0.189
0.195
0.21
0.163
0.133
*To convert t
••These valu
recommends
hese values to mg/L N, multiply by 0.82Z
bs may be conservative; however, if a more refined criterion is desired, EPA
a site-specific criteria modification.
These matters are discussed in more detail in EPA's "Technical Support
Document for Water Quality-Based Toxics Control."
Salt Water	The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important species
is very sensitive, saltwater aquatic organisms should not be affected unac-
ceptably if the four-day average concentration of un-ionized ammonia does
not exceed 0.035 mg/Lmore than once every three years on the average and
if the one-hour average concentration does not exceed 0.233 mg/L more
than once every three years on the average. Because sensitive saltwater ani-
mals appear to have narrow range of acute susceptibilities to ammonia, this
criterion will probably be as protective as intended only when the magni-
tudes and/or durations of excursions are appropriately small.
Criteria concentrations based on total ammonia for the pH range of 7.0
to 9.0, temperature range of 0 to 35°C, and salinities of 10, 20 and 30 g/kg
are provided in Tables 5 and 6. These values were calculated by
Hampson's (1977) program of Whitfield's (1974) model for hydrolysis of
ammonium ions in sea water (see original document).
21

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Table 5.—Water quality criteria for saltwater aquatic life based on total ammonia (mg/L).
Criteria Maximum Concentrations



Temperature (*C)




PH
0
5
10
15
20
25
30
35




Salinity =
lOg/kg



7.0
270
191
131
92
62
44
29
21
7.2
175
121
83
58
40
40
19
13
7.4
110
77
52
35
25
25
12
8.3
7.6
69
48
33
23
16
16
7.7
5.6
7.8
44
31
21
15
10
10
5.0
3.5
8.0
27
19
13
9.4
6.4
6.4
3.1
2.3
8.2
18
12
8.5
5.8
4.2
4.2
2.1
1.5
8.4
11
7.9
5.4
3.7
2.7
2.7
1.4
1.0
8.6
7.3
5.0
3.5
2.5
1.8
1.8
0.98
0.75
8.8
4.6
3.3
2.3
1.7
1.2
1.2
0.71
0.56
9.0
2.9
2.1
1.5
1.1
0.85
0.85
0.52
0.44




Salinity =
20 g/kg



7.0
291
200
137
96
64
44
31
21
7.2
183
125
87
60
42
29
20
14
7.4
166
79
54
37
27
18
12
8.7
7.6
73
50
35
23
17
11
7.9
5.6
7.8
46
31
23
15
11
7.5
5.2
3.5
8.0
29
20
14
9.8
6.7
4.8
3.3
2.3
8.2
19
13
8.9
6.2
4.4
3.1
2.1
1.6
8.4
12
8.1
5.6
4.0
2.9
2.0
1.5
1.1
8.6
7.5
5.2
3.7
2.7
1.9
1.4
1.0
0.77
8.8
4.8
3.3
2.5
1.7
1.3
0.94
0.73
0.56
9.0
3.1
2.3
1.6
1.2
0.87
0.69
0.54
0.44




Salinity =
30 g/kg



7.0
312
208
148
102
71
48
33
23
7.2
196
135
94
64
44
31
21
15
7.4
125
85
58
40
27
19
13
9.4
7.6
79
54
37
25
21
12
8.5
6.0
7.8
50
33
23
16
11
7.9
5.4
3.7
8.0
31
21
15
10
7.3
5.0
3.5
2.5
8.2
20
14
9.6
6.7
4.6
3.3
2.3
1.7
8.4
12.7
8.7
6.0
4.2
2.9
2.1
1.6
1.1
8.6
8.1
5.6
4.0
2.7
2.0
1.4
1.1
0.81
8.8
5.2
3.5
2.5
1.8
1.3
1.0
0.75
0.58
9.0
3.3
2.3
1.7
1.2
0.94
0.71
0.56
0.46
In the Agency's best scientific judgment, the average amount of time
aquatic ecosystems should be provided between excursions is three years.
The ability of ecosystems to recover differ greatly.
Site-specific criteria may be established if adequate justification is pro-
vided. This site-specific criterion may include not only site-specific criteria
concentrations, and mixing zone considerations, but also site-specific du-
rations of averaging periods and site-specific frequencies of allowed
exceedences.
22

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Table 6.-
-Water quality criteria for saltwater aquatic life based on total ammonia (mg/L).


Criteria Continuous Concentrations





Temperature (*C)




PH
0
5
10
15
20
25
30
35




Salinity =
10 g/kg



7.0
41
29
20
14
9.4
6.6
4.4
3.1
7.2
26
18
12
8.7
5.9
4.1
2.8
2.0
7.4
17
12
7.8
5.3
3.7
2.6
1.8
1.2
7.6
10
7.2
5.0
3.4
2.4
1.7
1.2
0.84
7.8
6.6
4.7
3.1
2.2
1.5
1.1
0.75
0.53
8.0
4.1
2.9
2.0
1.4
0.97
0.69
0.47
0.34
8.2
2.7
1.8
1.3
0.87
0.62
0.44
0.31
0.23
8.4
1.7
1.2
0.81
0.56
0.41
0.29
0.21
0.16
8.6
1.1
0.75
0.53
0.37
0.27
0.20
0.15
0.11
8.8
0.69
0.50
0.34
0.25
0.18
0.14
0 11
0.08
9.0
0.44
0.31
0.23
0.17
0.13
0.10
0.08
0.07




Salinity =
20 g/kg



7.0
44
30
21
14
9.7
6.6
4.7
3.1
7.2
27
19
13
9.0
6.2
4.4
3.0
2.1
7.4
18
12
8.1
5.6
4.1
2.7
1.9
1.3
7.6
11
7.5
5.3
3.4
2.5
1.7
1.2
0.84
7.8
6.9
4.7
3.4
2.3
1.6
1.1
0.78
0.53
8.0
4.4
3.0
2.1
1.5
1.0
0.72
0.50
0.34
8.2
2.8
1.9
1.3
0.94
0.66
0.47
0.31
0.24
8.4
1.8
1.2
0.84
0.59
0.44
0.30
0.22
0.16
8.6
1.1
0.78
0.56
0.41
0.28
0.20
0,15
0.12
8.8
0.72
0.50
0.37
0.26
0.19
0.14
0.11
0.08
9.0
0.47
0.34
0.24
0.18
0.13
0.10
0.08
0.07




Salinity =
30 g/kg



7.0
47
31
22
15
11
7.2
5.0
3.4
7.2
29
20
14
9.7
6.6
4.7
3.1
2.2
7.4
19
13
8.7
5.9
4.1
2.9
2.0
1.4
7.6
12
8.1
5.6
3.7
3.1
1.8
1.3
0.90
7.8
7.5
5.0
3.4
2.4
1.7
1.2
0.81
0.56
8.0
4.7
3.1
2.2
1.6
1.1
0.75
0.53
0.37
8.2
3.0
2.1
1.4
1.0
0.69
0.50
0.34
0.25
8.4
1.9
1.3
0.90
0.62
0.44
0.31
0.23
0.17
8.6
" 1.2
0.83
0.59
0.41
0.30
0.22
0.16
0.12
8.8
0.78
0.53
0.37
0.27
0.20
0.15
0.11
0.09
9.0
0.50
0.34
0.26
0.19
0.14
0.11
0.08
0.07
Use of criteria for developing water quality-based permit limits and for
designing waste treatment facilities requires selecting an appropriate
wasteload allocation model. Dynamic models are preferred for the applica-
tion of these criteria. Limited data or other considerations might make
their use impractical, causing reliance on a steady-state model.
Implementation Water quality standards for ammonia developed from these criteria should
specify use of environmental monitoring methods that are comparable to
the analytical methods employed to generate the toxicity data base. Total
ammonia may be measured using an automated idophenol blue method,
23

-------
such as described by Technicon Industrial Systems (1973) or U.S. EPA
(1979) method 350.1. Un-ionized ammonia concentrations should be calcu-
lated using the dissociation model of Whitfield (1974) as programmed by
Hampson (1977). This program was used to calculate most of the un-ion-
ized values for saltwater organisms. Accurate measurement of sample pH
is crucial in the calculation of the un-ionized ammonia fraction. The fol-
lowing equipment and procedures were used by EPA in the ammonia
toxicity studies to enhance the precision of pH measurements in salt water.
The pH meter reported two decimal places. A Ross electrode with ceramic
junction was used because of its rapid response time; an automatic temper-
ature compensation probe provided temperature correction. Note that the
responsiveness of a new electrode may be enhanced by holding it in sea-
water for several days prior to use. Two National Institute of Standards
and Technology buffer solutions for calibration preferred for their stability
were (1) potassium hydrogen phthalate (pH 4.00) and (2) disodium hydro-
gen phosphate (pH 7.4). For overnight or weekend storage, the electrode
was held in salt water, leaving the fill hole open. For daily use, the outer
half-cell was filled with electrolyte to the fill hole and the electrode
checked for stability. The electrode pair was calibrated once daily prior to
measuring pH of samples; it was never recalibrated during a series of mea-
surements. Following calibration, the electrode was soaked in sea water, of
salinity similar to the sample, for at least 15 minutes to achieve chemical
equilibrium and a steady-state junction potential. When measuring pH,
the sample was initially gently agitated or stirred to assure good mixing at
the electrode tip but without entraining air bubbles in the sample. Stirring
was stopped to read the meter. The electrode was allowed to equilibrate, so
the change in meter reading was less than 0.02 pH unit/minute before re-
cording. Following each measurement, the electrode was rinsed with sea
water and placed in fresh sea water for the temporary storage between
measurements.
Fresh Water — (50 F.R. 30784, July 29,1985)
Salt Water — (54 F.R. 19227, May 4,1989)
See Appendix A for Aquatic Life Methodology.
24

-------
ANTIMONY
7440-36-0
CRITERIA
Aquatic Life The available data for antimony indicate that acute and chronic toxicity to
freshwater aquatic life occur at concentrations as low as 9,000 and 1,600
Hg/L, respectively, and would occur at lower concentrations among spe-
cies that are more sensitive than those tested. Toxicity to algae occurs at
concentrations as low as 610 ng/L.
No saltwater organisms have been adequately tested with antimony,
and no statement can be made concerning acute or chronic toxicity.
Human Health For the protection of human health from the toxic properties of antimony
ingested through water and contaminated aquatic organisms, the ambient
water criterion is 146 ng/L.
For the protection of human health from the toxic properties of anti-
mony ingested through contaminated aquatic organisms alone, the
ambient water criterion is 45 mg/L.
Published human health critiera was recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 10/92. Recal-
culated IRIS values for antimony are 146.0 ng/L for ingestion of
contaminated water and organisms and 4,500 ng/L for ingestion of con-
taminated aquatic organisms only.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
25

-------
26

-------
ARSENIC
7440-38-2
CRITERIA
Aquatic Life Arsenic (III) — Freshwater — 1-hour average of 360 ng/L
4-day average of 190 ng/L
Saltwater — 1-hour average of 69 ng/L
4-day average of 36 ng/L
Summary	The chemistry of arsenic in water is complex, and the form present in a so-
lution is dependent on environmental conditions such as Eh, pH, organic
content, suspended solids, and sediment. The relative toxicities of the vari-
ous forms of arsenic apparently vary from species to species. For inorganic
arsenic (III), acute values for 16 freshwater animal species ranged from 812
Hg/L for a cladoceran to 97,000 n-g/L for a midge, but the three acute-
chronic ratios only ranged from 4.660 to 4.862. The five acute values for
inorganic arsenic (V) covered about the same range, but the single acute-
chronic ratio was 28.71. The six acute values for MSMA ranged from 3,243
to 1,403,000 ng/L. The freshwater residue data indicated that arsenic is not
bioconcentrated to a high degree, but that lower forms of aquatic life may
accumulate higher arsenic residues than fish. The low bioconcentration
factor and short half-life of arsenic in fish tissue suggest that residues
should not be a problem to predators of aquatic life.
The available data indicate that freshwater plants differ a great deal as
to their sensitivity to arsenic (III) and arsenic (V). In comparable tests, the
alga, Selenastrum capricornutum, was 45 times more sensitive to arsenic (V)
than to arsenic (III), although other data present conflicting information on
the sensitivity of this alga to arsenic (V). Many plant values for inorganic
arsenic (III) were in the same range as the available chronic values for
freshwater animals; several plant values for arsenic (V) were lower than
the one available chronic value.
The other toxicological data revealed a wide range of toxicity based on
tests with a variety of freshwater species and endpoints. Tests with early
life stages appeared to be the most sensitive indicator of arsenic toxicity.
Values obtained from this type of test with inorganic arsenic (III) were
lower than chronic values. For example, an effect concentration of 40 ng/L
was obtained in a test on inorganic arsenic (III) with embryos and toad
larvae.
Twelve species of saltwater animals have acute values for inorganic ar-
senic (III) from 232 to 16,030 ng/L, and the single acute-chronic ratio is
1.945. The only values available for inorganic arsenic (V) are for two inver-
tebrate and are between 2,000 and 3,000 n-g/L. Arsenic (HI) and arsenic (V)
are equally toxic to various species of saltwater algae, but the sensitivities
of the species range from 19 M-g/L to more than 1,000 ng/L- In a test with
an oyster, a BCF of 350 was obtained for inorganic arsenic (III).
27

-------
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of arse-
nic (III) does not exceed 190 ng/L more than once every three years on the
average and if the one-hour average concentration does not exceed 360
Hg/L more than once every three years on the average.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of arsenic (III) does not exceed 36 ng/L more
than once every three years on the average and if the one-hour average
concentration does not exceed 69 ng/L more than once every three years
on the average. This criterion might be too high wherever Skeletonema cos-
rarum or Thalassiosira aestivalis are ecologically important.
Not enough data are available to allow derivation of numerical na-
tional water quality criteria for freshwater aquatic life for inorganic arsenic
(V) or any organic arsenic compound. Inorganic arsenic (V) is acutely toxic
to freshwater aquatic animals at concentrations as low as 850 Hg/L, and an
acute-chronic ratio of 28 was obtained with the fathead minnow. Arsenic
(V) affected freshwater aquatic plants at concentrations as low as 48 ng/L.
Monosodium methanearsenace (MSMA) is acutely toxic to aquatic animals
at concentrations as low as 1,900 ng/L, but no data are available concern-
ing chronic toxicity to animals or toxicity to plants.
Very few data are available concerning the toxicity of any form of arse-
nic other than inorganic arsenic (III) to saltwater aquatic life. The available
data do show that inorganic arsenic (V) is acutely toxic to saltwater ani-
mals at concentrations as low as 2,319 ng/L and affected some saltwater
plants at 13 to 56 n-g/L- No data are available concerning the chronic toxic-
ity of any form of arsenic other than inorganic arsenic (III) to saltwater
aquatic life.
EPA believes that a measurement such as "acid-soluble" would pro-
vide a more scientifically correct basis upon which to establish criteria for
metals. The criteria were developed on this basis. However, at this time no
EPA-approved methods for such a measurement are available to imple-
ment the criteria through the regulatory programs of the Agency and the
States. The Agency is considering development and approval of methods
for a measurement such as acid-soluble. Until available, however, EPA rec-
ommends applying the criteria using the total recoverable method. This
has two impacts: (1) certain species of some metals cannot be analyzed di-
rectly because the total recoverable method does not distinguish between
individual oxidation states, and (2) these criteria may be overly protective
when based on the total recoverable method.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average amount of time it will take an un-
stressed system to recover from a pollution event in which exposure to
arsenic (III) exceeds the criterion. A stressed system, for example, one in
which several outfalls occur in a limited area, would be expected to require
more time for recovery. The resilience of ecosystems and their ability to re-
cover differ greatly, however, and site-specific criteria maybe established if
adequate justification is provided.
The use of criteria in designing waste treatment facilities requires the
selection of an appropriate wasteload allocation model. Dynamic models
are preferred for the application of these criteria. Limited data or other

-------
factors may make their use impractical, in which case one should rely on a
steady-state model. The Agency recommends the interim use of 1Q5 or
1Q10 for Criterion Maximum Concentration design flow and 7Q5 or 7Q10
for the Criterion Continuous Concentration design flow in steady-state
models for unstressed and stressed systems, respectively. These matters
are discussed in more detail in EPA's "Technical Support Document for
Water Quality-Based Toxics Control."
(50F.R. 30784, July 29,1985)
Human Health
Criteria	For the maximum protection of human health from the potential carcino-
genic effects due to exposure of inorganic arsenic through ingestion of
water and aquatic organisms that are contaminated, the ambient water
concentration should be zero based on the non-threshold assumption for
this chemical. However, zero level may not be attainable at the present
time. Therefore, the levels that may result in incremental increase of cancer
risk over the lifetime are estimated at 10"5, 10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for arsenic are 0.018 |xg/L for ingestion of con-
taminated water and organisms and 0.14 ng/L for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
(45 F.R. 79318, November 28,1980) (50 F.R. 30784, July 29,1985)
(57 F.R. 60848, December 22,1993)
See Appendix A for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
29

-------
30

-------
ASBESTOS
1332-21-4
CRITERIA
Aquatic Life No freshwater or saltwater organisms have been tested with any asbesti-
form mineral, therefore no statement can be made concerning its acute or
chronic toxicity.
Human Health Published human health criteria were recalculated to reflect data as of
12/92 (57 F.R. 60911). For the maximum protection of human health from
the potential carcinogenic effects of exposure to asbestos through ingestion
of water and contaminated aquatic organisms, the ambient water concen-
tration should be zero. The estimated level that would result in increased
lifetime cancer risks of 10"6 is 7,000,000 fibers/L. Estimates are for con-
sumption of aquatic organisms only, excluding the consumption of water.
(45 F.R. 79318, November 28,1980) (57 F.R. 60911, December 22,1993)
See Appendix C for Human Health Methodology.
31

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32

-------
BACTERIA
CRITERIA
Bathing Waters
(Full Body Contact) Based on a statistically sufficient number of samples (generally not less
than five samples equally spaced over a 30-day period), the geometric
mean of the indicated bacterial densities should not exceed one of the fol-
lowing (as displayed in Table 1).
Freshwater Based on a statistically sufficient number of samples (generally not less than
five samples equally spaced over a 30-day period), the geometric mean of the
indicated bacterial densities should not exceed one or the other of the follow-
ing:*
E. coli 126 per 100 ml; or
enterococci 33 per 100 ml;
no sample should exceed a one-sided confidence limit (C.L.) calculated
using the following as guidance:
designated bathing beach 75% C.L.
moderate use for bathing 82% C.L.
light use for bathing 90% C.L.
infrequent use for bathing 95% C.L.
based on a site-specific log standard deviation; or if site data are
insufficient to establish a log standard deviation, then using 0.4 as the log
standard deviation for both indicators.
Marine Water Based on a statistically sufficient number of samples (generally not less
than five samples equally spaced over a 30-day period), the geometric
mean of the enterococci densities should not exceed 35 per 100 ml; no sam-
ple should exceed a one-sided confidence limit using the following as
guidance:
designated bathing beach 75% C.L.
moderate use for bathing 82% C.L.
light use for bathing 90% C.L.
infrequent use for bathing 95% C.L.
based on a site-specific log standard deviation; or if site data are insufficient
to establish a log standard deviation, then using 0.7 as the log standard
deviation.
'Only one indicator should be used. The regulatory agency should select the
appropriate indicator for its conditions.
33

-------
Shellfish
Harvesting
Waters	The median fecal coliform bacterial concentration should not exceed 14
Most Probable Number (MPN) per 100 mL, with not more than 10 percent
of samples exceeding 43 MPN per 100 mL for the taking of shellfish.
The microbiological criterion for shellfish water quality has been estab-
lished by international agreement to be 70 total coliforms per 100 mL,
using a median MPN, with no more than 10 percent of the values exceed-
ing 230 total coliforms per 100 mL. For evaluation of waters for
recreational taking of shellfish, EPA recommends fecal coliform bacteria
rather than total coliform bacteria.
REFERENCES:
Bathing Water Criteria - EPA 440/5-84-002
Bathing Water Criteria Laboratory Methods - EPA 600/4-85/076
Shellfish Water Criteria - Quality Criteria for Water (1976)
GPO Access #055-001-01049-4
(51 F.R. 8012, March 7,1986)
Table 1.—Criteria for indicator for biological densities.
SINGLE SAMPLE MAXIMUM ALLOWABLE DENSITY (4), (5)
ACCEPTABLE
SWIMMING
ASSOCIATED
GASTROENTERITIS
RATE PER 1000
SWIMMERS
STEADY-STATE
GEOMETRIC
MEAN
INDICATOR
DENSITY
DESIGNATED
BEACHAREA
(UPPER 75%
MODERTE FULL
BODY CONTACT
RECREATION
(UPPER 82%C.L)
LIGHTLY USED
FULL BODY
CONTACT
RECREATION
(UPPER 90% C.L)
INFREQUENTLY
USED FULL BODY
CONTACT
RECREATION
(UPPER 95%C.L)
Freshwater*
enterococci
E. coli
33'
0)
126

61
235
78
298
107
409
151
575
* Only one indicator should be used. The regulatory agency should select the appropriate indicator for its conditions.
Marine Water
enterococci
NOTES:
19
35
(3)
104
158
276
501
(1)	Calculated to nearest whole number using equation:
illness rate/1000 people + 6.28
(mean enterococci density) = antilog10 	9 40 	
(2)	Calculated to nearest whole number using equation:
illness rate/1000 people + 11.74
9.40
(mean E. coli density) = antilog 10
(3)	Calculated to nearest whole number using equation:
(mean enterococci density) = antilog 10
(4)	Single sample limit = antilog10
(log10 indicator geometric mean
density/100ml)
illness rate/1000 people - 0.20
12.17
x (log10 stand.
10 ¦
deviation)
-34
factor determined from areas
under the normal probability
curve for the assumed level of
probability
The appropriate factors for the indicated one-sided confidence levels:
75% C.L. — .675
82% C.L— .935
90% C.L. — 1.28
95% C.L. — 1.65
(5) Based on the observed log standard deviations during the EPA studies: 0.4 for freshwater E. coli and
enterococci; and 0.7 for marine water enterococci. Each jurisdiction should establish its own standard
deviation for its conditions, which would than vary the single sample limit.

-------
BARIUM
7440-39-3
CRITERIA
Aquatic Life
Introduction
Human Health
1 mg/L for domestic water supply (health).
Barium is a yellowish white metal of the alkaline earth group. It occurs in
nature chiefly as barite (BaS04) and witherite (BaC03), both highly insolu-
ble salts. The metal is stable in dry air but readily oxidized by humid air or
water.
Many of the salts of barium are soluble in both water and acid; soluble
barium salts are reported to be poisonous. However, barium ions are gen-
erally thought to precipitate rapidly or be removed from a solution by
absorption and sedimentation.
While barium is a malleable, ductile metal, its major commercial value
is in its compounds. Barium compounds are used in a variety of industrial
applications, including the metallurgic, paint, glass, and electronics indus-
tries, as well as for medicinal purposes.
For the protection of human health from the toxic properties of barium in-
gested through water and contaminated aquatic organisms, the ambient
water criterion is 1 mg/L.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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BENZENE
71-43-2
CRITERIA
Aquatic Life The available data for benzene indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 5,300 ng/L and would occur
at lower concentrations among species that are more sensitive than those
tested. No data are available concerning the chronic toxicity of benzene to
sensitive freshwater aquatic life.
The available data for benzene indicate that acute toxicity to saltwater
aquatic life occurs at concentrations as low as 5,100 ng/L and would occur
at lower concentrations among species that are more sensitive than those
tested. No definitive data are available concerning the chronic toxicity of
benzene to sensitive saltwater aquatic life, but adverse effects occur at con-
centrations as low as 700 pig/L with a fish species exposed for 168 days.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to benzene through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tions should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60911). Recalculated IRIS values for benzene are 1.2 ng/L for inges-
tion of contaminated water and organisms and 71 ng/L for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60911, December 22,1992)
See Appendix C for Human Health Methodology.
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38

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BENZIDINE
92-87-5
CRITERIA
Aquatic Life The available data for benzidine indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 2,500 ng/L and would occur
at lower concentrations among species that are more sensitive than those
tested. No data are available concerning the chronic toxicity of benzidine
to sensitive freshwater aquatic life.
Since saltwater organisms have not been tested with benzidine, no
statement can be made concerning its acute and chronic toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to benzidine through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tions should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in an incremental increase of cancer
risk over a lifetime are estimated at 10'5,10"6, and 10"7.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60913). Recalculated IRIS values for benzidine are 0.00012 fxg/L for
ingestion of contaminated water and organisms and 0.00054 ng/L for in-
gestion of contaminated aquatic organisms only. IRIS values are based on a
10"6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60913, December 22,1993)
See Appendix C for Human Health Methodology.
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BERYLLIUM
7440-41-7
CRITERIA
Aquatic Life The available data for beryllium indicate that acute and chronic toxicity to
freshwater aquatic life occur at concentrations as low as 130 and 5.3 ng/L,
respectively, and would occur at lower concentrations among species that
are more sensitive than those tested. Hardness has a substantial effect on
acute toxicity.
The limited saltwater database available for beryllium does not permit
any statement concerning acute or chronic toxicity.
Human Health Human health criteria have been withdrawn for this compound (see
57 F.R. 60885, December 22, 1992). Although the human health criteria are
withdrawn, EPA published a document for this compound that may con-
tain useful human health information. This document was originally
noticed in 45 F.R. 79326, November 28,1980.
(45 F.R. 79318, November 28, 1980) (57 F.R. 60911, December 22,1992)
See Appendix C for Human Health Methodology.
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42

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BENZENE HEXACHLORIDE (BHC)
680-73-1
(See also: Hexachlorocyclohexane)
CRITERIA
Aquatic Life The available data for a mixture of isomers of BHC indicate that acute tox-
icity to freshwater aquatic life occurs at concentrations as low as 100 ng/L
and would occur at lower concentrations among species that are more sen-
sitive than those tested.
No data are available concerning the chronic toxicity of a mixture of
isomers of BHC to sensitive freshwater aquatic life.
The available data for a mixture of isomers of BHC indicate that acute
toxicity to saltwater aquatic life occurs at concentrations as low as 0.34
Hg/L and would occur at lower concentrations among species that are
more sensitive than those tested. No data are available concerning the
chronic toxicity of a mixture of isomers of BHC to sensitive saltwater
aquatic life.
(45 F.R. 79318, November 28,1980)
See Appendix B for Aquatic Life Methodology.
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BORON
CRITERION 	
750 ng/L for long-term irrigation on sensitive crops. Data are insufficient
to determine acute or chronic toxicity of boron to freshwater or saltwater
aquatic life.
Introduction Boron is usually found as a sodium or calcium borate salt in nature, rather
than in an elemental form. Boron salts are used in fire retardants, produc-
tion of glass, leather tanning and finishing industries, cosmetics,
photographic materials, metallurgy, and for high energy rocket fuels. Ele-
mental boron also can be used in nuclear reactors for neutron absorption.
Borates are used as "burnable" poisons.
Rationale	Boron is an essential element for the growth of plants, but no evidence in-
dicates that it is required by animals. The maximum concentration found
in 1,546 samples of river and lake waters from various parts of the United
States was 5.0 mg/L; the mean value was 0.1 mg/L. Groundwaters could
contain substantially higher concentrations in certain places. The concen-
tration in sea water is reported as 4.5 mg/L in the form of borate. Naturally
occurring concentrations of boron should have no effect on aquatic life.
The minimum lethal dose for minnows exposed to boric acid at 20°C
for 6 hours was reported to be 18,000 to 19,000 mg/L in distilled water and
19,000 to 19,500 mg/L in hard water. In the dairy cow, 16 to 20 g/day of
boric acid for 40 days produced no ill effects.
Sensitive crops have shown toxic effects at 1,000 ug/L or less of boron.
When the boron concentration in irrigation waters was greater than 0.75
Ug/L, some sensitive plants such as citrus began to show injury. Water con-
taining 2 ug/L boron (with neutral and alkaline soils of high absorption
capacities) might be used for some time without injury to sensitive plants.
The criterion of 750 ng/L is thought to protect sensitive crops during long-
term irrigation.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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46

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D
CADMIUM
7440-43-9
CRITERIA
Aquatic Life Saltwater — 1-hour average of 43.0 ng/L
4-day average of 9.3 ^g/L
Freshwater criteria are hardness dependant. See text.
Criteria	Freshwater acute values for cadmium are available for species in 44 genera
and range from 1.0 ng/L for rainbow trout to 28,000 ng/Lfor a mayfly. The
antagonistic effect of hardness on acute toxicity has been demonstrated
with five species. Chronic tests have been conducted on cadmium with 12
freshwater fish species and four invertebrate species with chronic values
ranging from 0.15 ng/L for Daphnia magna to 156 |xg/L for the Atlantic
salmon. Acute-chronic ratios are available for eight species and range from
0.9021 for the chinook salmon to 433.8 for the flagfish. Freshwater aquatic
plants are affected by cadmium at concentrations ranging from 2 to 7,400
Hg/L. These values are in the same range as the acute toxicity values for
fish and invertebrate species and are considerably above the chronic val-
ues. Bioconcentration factors (BCFs) for cadmium in freshwater range
from 164 to 4,190 for invertebrates and from 3 to 2,213 for fishes.
Saltwater acute values for cadmium and five species of fishes range
from 577 ng/L for larval Atlantic silverside to 114,000 ng/L for juvenile
mummichog. Acute values for 30 species of invertebrates range from
15.5 fig/L for a mysid to 135,000 ng/L for an oligochaete worm. The acute
toxicity of cadmium generally increases as salinity decreases.
The effect of temperature seems to be species-specific. Two life-cycle
tests with Mysidopsis bahia under different test conditions resulted in sim-
ilar chronic values of 82 and 7.1 |xg/L, but the acute-chronic ratios were 1.9
and 15, respectively. The acute values appear to reflect effects of salinity
and temperature, whereas the few available chronic values apparently do
not. A life-cycle test with Mysidopsis bigelowi also resulted in a chronic
value of 7.1 fig/L and an acute-chronic ratio of 15. Studies with microalgae
and macroalgae revealed effects at 22.8 to 860 ng/L.
BCFs determined with a variety of saltwater invertebrates ranged from
5 to 3,160. BCFs for bivalve molluscs were above 1,000 in long exposures,
with no indication that steady state had been reached. Cadmium mortality
is cumulative for exposure periods beyond four days. Chronic cadmium
exposure resulted in significant effects on the growth of bay scallops at
78 ng/L and on reproduction of a copepod at 44 ng/L.
The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms
and Their Uses" indicate that, except possibly where a locally important
species is very sensitive, freshwater aquatic organisms and their uses
should not be affected unacceptably if the four-day average concentration
(in |xg/L) of cadmium does not exceed the numerical value given by
g(0.7852[ln(hardness)]-3.490)
47

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more than once every three years on the average and if the one-hour
average concentration (in (ig/L) does not exceed the numerical value given
by
(1.128 [In (hardness)]-3.828)
c
more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaCC>3, the four-day average
concentrations of cadmium are 0.66, 1.1, and 2.0 |xg/L, respectively, and
the one-hour average concentrations are 1.8, 3.9, and 8.6 ng/L. If brook
trout, brown trout, and striped bass are as sensitive as some data indicate,
they might not be protected by this criterion.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of cadmium does not exceed 9.3 ng/L more
than once every three years on the average, and if the one-hour average
concentration does not exceed 43 ng/L more than once every three years
on the average.
The little information that is available concerning the sensitivity of the
American lobster to cadmium indicates that this important species might
not be protected by this criterion. In addition, data suggest that the acute
toxicity of cadmium is salinity dependent; therefore, the one-hour average
concentration might be under protective at low salinities and overprotec-
tive at high salinities.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average amount of time it will take an un-
stressed system to recover from a pollution event in which exposure to
cadmium exceeds the criterion. A stressed system, for example, one in
which several outfalls occur in a limited area, would be expected to require
more time for recovery. The resilience of ecosystems and their ability to re-
cover differ greatly, however, and site-specific criteria may be established if
adequate justification is provided.
The use of criteria in designing waste treatment facilities requires the
selection of an appropriate wasteload allocation model. Dynamic models
are preferred for the application of these criteria. Limited data or other fac-
tors may make their use impractical, in which case one should rely on a
steady-state model. The Agency recommends the interim use of 1Q5 or
1Q10 for Criterion Maximum Concentration design flow and 7Q5 or 7Q10
for the Criterion Continuous Concentration design flow in steady- state
models for unstressed and stressed systems, respectively. These matters
are discussed in more detail in the "Technical Support Document for Water
Quality-Based Toxics Control."
Human Health
Criteria	Human health criteria have been withdrawn for this compound (see
57 F.R. 60885, December 22, 1992). Although the human health criteria are
withdrawn, EPA published a document for this compound that may con-
tain useful human health information. This document was originally
noticed in 45 F.R. 79326, November 28,1980.
(45 F.R. 79318, November 28,1980) (50 F.R. 30784, July 29,1985)
(57 F.R. 60885, December 22,1992)
See Appendix A for Aquatic Life Methodology.
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CARBON TETRACHLORIDE
56-23-5
CRITERIA
Aquatic Life The available data for carbon tetrachloride indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 35,200 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
carbon tetrachloride to sensitive freshwater aquatic life.
The available data for carbon tetrachloride indicate that acute toxicity
to saltwater aquatic life occurs at concentrations as low as 50,000 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
carbon tetrachloride to sensitive saltwater aquatic life.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to carbon tetrachloride through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10"5,10"6, and 10"7.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60911, December 22,1992). Recalculated IRIS values for carbon tet-
rachloride are 0.25 [ig/L for ingestion of contaminated water and
organisms and 4.4 ng/L for ingestion of contaminated aquatic life organ-
isms. IRIS values are based on a 10"6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60911, December 22,1992)
See Appendix C for Human Health Methodology.
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CHLORDANE
57-74-9
CRITERIA
Aquatic Life For chlordane, the criterion to protect freshwater aquatic life as derived
using the guidelines is 0.0043 n-g/L as a 24-hour average. The concentra-
tion should not exceed 2.4 pig/L at any time.
For chlordane, the criterion to protect saltwater aquatic life as derived
using the guidelines is 0.0040 ng/L as a 24-hour average. The concentra-
tion should not exceed 0.09 ng/L at any time.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to chlordane through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tion should be zero based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in incremental increase of cancer risk
over a lifetime are estimated at 10'5,10"6, and 10"7.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60914, December 22, 1992). Recalculated IRIS values for chlordane
are 0.00057 ng/L for ingestion of contaminated water and organisms and
0.00059 ng/L for ingestion of contaminated aquatic organisms only. IRIS
values are based on a 10"6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60914, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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CHLORIDE
16887-00-6
CRITERIA
Aquatic Life Not to exceed a one-hour average of 860 mg/L or a four-day average of
230 mg/L for freshwater aquatic life.
The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms
and Their Uses" indicate that, except possibly where a locally important
species is very sensitive, freshwater aquatic organisms and their uses
should not be affected unacceptably if the four-day average concentration
of dissolved chloride, when associated with sodium, does not exceed 230
mg/L more than once every three years on the average and if the one-hour
average concentration does not exceed 860 mg/L more than once every
three years on the average.
These criteria probably will not be adequately protective when the
chloride is associated with potassium, calcium, or magnesium, rather than
sodium. In addition, because freshwater animals have a narrow range of
acute susceptibilities to chloride, excursions above these criteria might af-
fect a substantial number of species.
(53 F.R. 19028, May 26,1988)
See Appendix A for Aquatic Life Methodology.
53

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54

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CHLORINATED BENZENES
CRITERIA
Aquatic Life The available data for chlorinated benzenes indicate that acute toxicity to
freshwater aquatic life would occur at concentrations as low as 250 ng/L
and at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of the
more toxic of the chlorinated benzenes to sensitive freshwater aquatic life,
but toxicity occurs at concentrations as low as 50 fxg/L for a fish species ex-
posed for 7.5 days.
The available data for chlorinated benzenes indicate that acute and
chronic toxicity to saltwater aquatic life occur at concentrations as low as
160 and 129 |j.g/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested.
Human Health
Monochlorobenzene (Chlorobenzene) 108-90-7
Published human health criteria were recalculated using Integrated Risk
Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60911, December 22,1992). Recalculated IRIS values for chloroben-
zene are 680.0 ng/L for contaminated water and organisms and 21,000
M-g/L for ingestion of contaminated aquatic organisms only.
Trichlorobenzenes
Because of insufficiency in the available information for the trichloro-
benzenes, a criterion cannot be derived at this time using the present
guidelines.
1,2,4,5-Tetrachlorobenzene 95-94-3
For the protection of human health from the toxic properties of 1,2,4,5-
tetrachlorobenzene ingested through water and contaminated aquatic
"organisms, the ambient water criterion is determined to be 38 ng/L.
For the protection of human health from the toxic properties of 1,2,4,5-
tetrachlorobenzene ingested through contaminated aquatic organisms
alone, the ambient water criterion is determined to be 48 ng/L.
Pentachlorobenzene 608-93-5
For the protection of human health from the toxic properties of pen-
tachlorobenzene ingested through water and contaminated aquatic
organisms, the ambient water criterion is determined to be 74 ng/L.
For the protection of human health from the toxic properties of pen-
tachlorobenzene ingested through contaminated aquatic organisms alone,
the ambient water criterion is determined to be 85 ng/L.
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Hexachlorobenzene 118-74-1
For the maximum protection of human health from the potential carcino-
genic effects due to exposure of hexachlorobenzene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero based on the nonthreshold assumption
for this chemical. However, zero level may not be attainable at the present
time. Therefore, the levels that may result in incremental increase of cancer
risk over a lifetime are estimated at 10'5,10"6, and 10'7.
Published human health criteria were recalculated using IRIS to reflect
available data as of 12/92 (57 F.R. 609121, December 22,1992). The recal-
culated IRIS value for chlorinated benzenes is .00075 jig/L for ingestion of
contaminated water and organisms and 0.00077 ng/L for ingestion of con-
taminated organisms only. IRIS values are based on a 10" risk level for
carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology
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CHLORINATED ETHANES
CRITERIA
Aquatic Life The available freshwater data for chlorinated ethanes indicate that toxicity
increases greatly with increasing chlorination and that acute toxicity
occurs at concentrations as low as 118,000 ng/L for 1,2-dichloroethane;
18,000 (ig/L for two trichloroethanes; 9,320 ng/L for two tetrachloro-
ethanes; 7,240 ng/L for pentachloroethane; and 980 ng/L for hexachloro-
ethane.
Chronic toxicity occurs at concentrations as low as 20,000 ng/L for 1,2-
dichloroethane; 9,400 ng/L for 1,1,2-trichchloroethane; 2,400 ng/L for
1,1,2,2-tetrachloroethane; 1,100 n-g/L for pentachloroethane; and 540 ng/L
for hexachloroethane. Acute and chronic toxicity would occur at lower
concentrations among species that are more sensitive than those tested.
The available saltwater data for chlorinated ethanes indicate that toxic-
ity increases greatly with increasing chlorination and that acute toxicity to
fish and invertebrate species occurs at concentrations as low as 113,000
Hg/L for 1,2-dichloroethane; 31,200 (xg/L for 1,1,1-trichloroethane; 9,020
Hg/L for 1,1,2,2-tetrachloroethane; 390 ng/L for pentachloroethane; and
940 ng/L for hexachloroethane.
Chronic toxicity occurs at concentrations as low as 281 (ig/L for pen-
tachloroethane. Acute and chronic toxicity would occur at lower
concentrations among species that are more sensitive than those tested.
Human Health
1,2-Dichloroethane 107-06-2
For the maximum protection of human health from the potential carcino-
genic effects of exposure to 1,2-dichloroethane through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10'5,10"6, and 10'7.
Health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for 1,2-dichloroethane are 0.38 ng/L for ingestion of
contaminated water and organisms and 99 ng/L for ingestion of contami-
nated aquatic organisms only. IRIS values are based on 10"6 risk level for
carcinogens.
1,1,2-Trichloroethane 79-00-5
For the maximum protection of human health from the potential carcino-
genic effects of exposure to 1,1,2-trichloroethane through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
57

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present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10~5,10"6, and 10"7.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60948). Recalculated IRIS values for 1,1,2-
trichloroethane are 0.60 ng/L for ingestion of contaminated water and
organisms and 42 M-g/L for ingestion of contaminated aquatic organisms
only. IRIS values are based on a 10"6 risk level for carcinogens.
1,1,2,2-Tetrachloroethane 79-34-5
For the maximum protection of human health from the potential carcino-
genic effects of exposure to 1,1,2,2-tetrachloroethane through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60848). Recalculated IRIS values for 1,1,2,2- tetra-
chloroethane are 0.17 ng/L for ingestion of contaminated water and
organisms and 11.0 ng/L for ingestion of contaminated aquatic organisms
only. IRIS values are based on a 10"6 risk level for carcinogens.
Hexachloroethane 67-72-1
For the maximum protection of human health from the potential carcino-
genic effects of exposure to hexachloroethane through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60848). Recalculated IRIS values for hexachloro-
ethane are 1.9 ng/L for ingestion of contaminated water and organisms
and 8.9 ng/L for ingestion of contaminated aquatic organisms only. IRIS
values are based on 10"6 risk level for carcinogens.
1,1,1-Trichloroethane 71-55-6
Human health criteria have been withdrawn for one chlorinated ethane,
1,1,1-trichloroethane (see 57 F.R. 60885, December 22,1992). Although the
human health criteria are withdrawn, EPA published a document for this
compound that may contain useful human health information. This docu-
ment was originally noticed in 45 F.R. 79328, November 28,1980.
Because of insufficient available data for monochloroethane, 1,1-
dichloroethane, 1,1,1,2-tetrachloroethane, and pentachloroethane, satisfac-
tory criteria cannot be derived at this time using the present guidelines.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
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CHLORINATED NAPHTHALENES
CRITERIA
Aquatic Life The available data for chlorinated naphthalenes indicate that acute toxicity
to freshwater aquatic life occurs at concentrations as low as 1,600 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
chlorinated naphthalenes to sensitive freshwater aquatic life.
The available data for chlorinated naphthalenes indicate that acute tox-
icity to saltwater aquatic life occurs at concentrations as low as 7.5 ng/L
and would occur at lower concentrations among species that are more sen-
sitive than those tested. No data are available concerning the chronic
toxicity of chlorinated naphthalenes to sensitive saltwater aquatic life.
Human Health Human health criteria was calculated for 2-chloronaphthelene using Inte-
grated Risk Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60890). The calculated IRIS values for 2-chloronaphthelene is
1,700 ng/L for ingestion of contaminated water and organisms and
4,300 mg/L for organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22, 1992)
See Appendix B for Aquatic Life Methodology.
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CHLORINATED PHENOLS
CRITERIA
Aquatic Life The available freshwater data for chlorinated phenols indicate that toxicity
generally increases with increased chlorination and that acute toxicity oc-
curs at concentrations as low as 30 |ig/L for 4-chloro-3-methylphenol to
greater than 500,000 ng/L for other compounds. Chronic toxicity occurs at
concentrations as low as 970 pig/L for 2,4,6-trichlorophenol. Acute and
chronic toxicity would occur at lower concentrations among species that
are more sensitive than those tested.
The available saltwater data for chlorinated phenols indicate that toxic-
ity generally increases with increasing chlorination and that acute toxicity
occurs at concentrations as low as 440 ng/L for 2,3,5,6-tetrachlorophenol
and 29,700 jig/L for 4-chlorophenol. Acute toxicity would occur at lower
concentrations among species that are more sensitive than those tested. No
data are available concerning the chronic toxicity of chlorinated phenols to
sensitive saltwater aquatic life.
Human Health
3-Chlorophenol
Sufficient data are not available for 3-chlorophenol to derive a level that
would protect against the potential toxicity of this compound. According
to available organoleptic data, the estimated level is 0.1 ng/L to control un-
desirable taste and odor qualities of ambient water. Organoleptic data do
have limitations as a basis for establishing a water quality criterion but do
not have a demonstrated relationship to potentially adverse effects on
human health.
4-Chlorophenol	106-48-9
Sufficient data are not available for 4-chlorophenol to derive a level that
would protect against the potential toxicity of this compound. According
to available organoleptic data, to control undesirable taste and odor quali-
ties of ambient water the estimated level is 0.1 M-g/L. Organoleptic data do
have limitations as a basis for establishing a water quality criterion but do
not have a demonstrated relationship to potentially adverse effects on
human health.
2,3-Dichlorophenol
Sufficient data are not available for 2,3-dichlorophenol to derive a level
that would protect against the potential toxicity of this compound. Accord-
ing to available organoleptic data, the estimated level is 0.04 ng/L to
control undesirable taste and odor qualities of ambient water. Organolep-
tic data do have limitations as a basis for establishing a water quality
criterion but do not have a demonstrated relationship to potentially ad-
verse effects on human health.
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2.5-Dichlorophenol
Sufficient data are not available for 2,5-dichlorophenol to derive a level
that would protect against the potential toxicity of this compound. Accord-
ing to available organoleptic data, the estimated level is 0.5 M-g/L to control
undesirable taste and odor qualities of ambient water. Organoleptic data
do have limitations as a basis for establishing a water quality criterion but
do not have a demonstrated relationship to potentially adverse effects on
human health.
2.6-Dichlorophenol
Sufficient data are not available for 2,6-dichlorophenol to derive a level
that would protect against the potential toxicity of this compound. Accord-
ing to available organoleptic data, to control undesirable taste and odor
qualities of ambient water, the estimated level is 0.2 ng/L. Organoleptic
data do have limitations as a basis for establishing a water quality criterion
but do not have a demonstrated relationship to potentially adverse effects
on human health.
3,4-Dichlorophenol
Sufficient data are not available for 3,4-dichlorophenol to derive a level
that would protect against the potential toxicity of this compound. Accord-
ing to available organoleptic data, to control undesirable taste and odor
qualities of ambient water, the estimated level is 0.3 p.g/L. Organoleptic
data do have limitations as a basis for establishing a water quality criterion
but do not have a demonstrated relationship to potentially adverse effects
on human health.
2.4.5-Trichlorophenol	95-95-4
For comparison purposes, two approaches were used to derive criterion
levels for 2,4,5-trichlorophenol. Based on available toxicity data, to protect
public health, the derived level is 2.6 mg/L. Using available organoleptic
data, to control undesirable taste and odor quality of ambient water, the es-
timated level is 1.0 ng/L. Organoleptic data do have limitations as a basis
for establishing a water quality criterion but do not have a demonstrated
relationship to potentially adverse effects on human health.
2.4.6-Trichlorophenol	88-06-2
For the maximum protection of human health from the potential carcino-
genic effects of exposure to 2,4,6-trichlorophenol through the ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10"5,10"6, and 10'7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for 2,4,6-Trichlorophenol are 2.1 fig/L for inges-
tion of contaminated water and organisms and 6.5 ng/L for ingestion of
contaminated organisms only. IRIS values are based on a 10"6 risk level for
carcinogens.
62

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2,3,4,6-Tetrachlorophenol
Sufficient data are not available for 2,3,4,6-tetrachlorophenol to derive a
level that would protect against the potential toxicity of this compound.
According to available organoleptic data, to control undesirable taste and
odor qualities of ambient water the estimated level is 1.0 ng/L. Organolep-
tic data have limitations as a basis for establishing a water quality criterion
and a demonstrated relationship to potentially adverse effects on human
health.
2-Methyl-4-Chlorophenol
Sufficient data are not available for 2-methyl-4-chlorophenol to derive a
criterion level that would protect against any potential toxicity of this com-
pound. According to available organoleptic data, to control undesirable
taste and odor qualities of ambient water, the estimated level is 1,800 ng/L.
Organoleptic data do have limitations as a basis for establishing a water
quality criterion but do not have a demonstrated relationship to poten-
tially adverse effects on human health.
3-Methyl-4-Chlorophenol	59-50-7
Sufficient data are not available for 3-methyl-4-chlorophenol to derive a
criterion level that would protect against any potential toxicity of this com-
pound. According to available organoleptic data, to control undesirable
taste and odor qualities of ambient water, the estimated level is 3,000 ng/L.
Organoleptic data do have limitations as a basis for establishing a water
quality criterion but do not have a demonstrated relationship to poten-
tially adverse effects on human health.
3-Methyl-6-Chlorophenol
Sufficient data are not available for 3-methyl-6-chlorophenol to derive a
criterion level that would protect against any potential toxicity of this com-
pound. According to available organoleptic data, the estimated level is 20
H-g/L to control undesirable taste and odor qualities of ambient water. Or-
ganoleptic data do have limitations as a basis for establishing a water
quality criterion but do not have a demonstrated relationship to poten-
tially adverse effects on human health.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
63

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CHLORINE
7782-50-5
Freshwater — 4-day average of 11 ng/L
1-hour average of 19 ng/L
Saltwater — 4-day average of 7.5 ng/L
1-hour average of 13 ng/L
Thirty-three freshwater species in 28 genera have been exposed to total re-
sidual chlorine (TRC); the acute values range from 28 ng/L for Daphnia
magna to 710 |xg/L for the threespine stickleback. Fish and invertebrate spe-
cies had similar ranges of sensitivity. Freshwater chronic tests have been
conducted with two invertebrate and one fish species. The chronic values
for these three species ranged from less than 3.4 to 26 ng/L, with acute-
chronic ratios from 3.7 to greater than 78.
The acute sensitivities of 24 species of saltwater animals in 21 genera
have been determined for CPO, and the LC50 range is from 26 M-g/L for the
eastern oyster to 1,418 ng/L for a mixture of two shore crab species. This
range is very similar to that observed with freshwater species: fishes and
invertebrates had similar sensitivities. Only one chronic test has been con-
ducted with a saltwater species, Menidia peninsulae; the acute chronic ratio
was 1.162.
The available data indicate that aquatic plants are more resistant to
chlorine than fish and invertebrate species.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of total
residual chlorine does not exceed 11 ng/L more than once every three
years on the average, and if the one-hour average concentration does not
exceed 19 ^g/L more than once every three years on the average.
The procedures described in the guidelines indicate that saltwater
aquatic organisms and their uses should not be affected unacceptably (ex-
cept possibly where a locally important species is very sensitive) if the
four-day average concentration of chlorine-produced oxidants does not ex-
ceed 7.5 ng/L more than once every three years on the average, and if the
one-hour average concentration does not exceed 13 (xg/L more than once
every three years on the average.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average amount of time for an unstressed
system to recover from a pollution event in which exposure to chlorine ex-
ceeds the criterion. A stressed system — for example, one in which several
outfalls occur in a limited area — would be expected to require more time
CRITERIA
Aquatic Life
Summary
65

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for recovery. An ecosystem's resilience and ability to recover differ greatly,
however, and site-specific criteria may be established if adequate justifica-
tion is provided.
The use of criteria in designing waste treatment facilities requires the
selection of an appropriate wasteload allocation model. Dynamic models
are preferred for the application of these criteria; however, if limited data
or other factors make their use impractical, one should rely on a steady-
state model. The Agency recommends the interim use of 1Q5 or 1Q10 for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for the
Criterion Continuous Concentration design flow in steady-state models
for unstressed and stressed systems, respectively. These matters are dis-
cussed in more detail in EPA's "Technical Support Document for Water
Quality-Based Toxics Control."
(50 F.R. 30784, July 29,1985)
See Appendix A for Aquatic Life Methodology.

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CHLOROALKYLETHERS
CRITERIA
Aquatic Life The available data for chloroalkyl ethers indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 238,000 jig/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No definitive data are available concerning the chronic
toxicity of chloroalkyl ethers to sensitive freshwater aquatic life.
No saltwater organism has been tested with any chloroalkyl ether, and
therefore, no statement can be made concerning acute or chronic toxicity.
Human Health
Bis(Z-Chloroisopropyl) 108-60-1
Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for bis(2-chloroisopropyl) ether are 1,400 M-g/L for
ingestion of contaminated water and organisms and 170,000 ng/L for in-
gestion of contaminated aquatic organisms only.
Bis(Chloromethyl)
For the maximum protection of human health from the potential carcino-
genic effects of exposure to bis(chloromethyl) ether through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10"5,10"6, and 10"7. The cor-
responding recommended criteria are 37.6 x 10"6 ng/L, and 0.376 x 10"6
M-g/L, respectively. If these estimates are made for consumption of aquatic
organisms only, excluding consumption of water, the levels are 18.4 x 10"3
Hg/L, 1.84 x 10"3 ng/L, respectively.
Bis(2-Chloroethyl) 111-44-4
For the maximum protection of human health from the potential carcino-
genic effects of exposure to bis(2-chloroethyl) ether through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10'5, 10"6, and 10"7.The corre-
sponding recommended criteria are 0.30 jxg/L, 0.030 ng/L, and 0.003
Hg/L, respectively. If these estimates are made for consumption of aquatic
organisms only, excluding consumption of water, the levels are 13.6 ng/L,
1.36 |xg/L, and 0.136 ng/L, respectively.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
67

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68

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CHLOROFORM
67-66-3
CRITERIA
Aquatic Life The available data for chloroform indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 28,900 ng/L and would
occur at lower concentrations among species that are more sensitive than
the three tested species. As indicated by 27-day LC50 values, chronic toxic-
ity occurs at concentrations as low as 1,240 ng/L and could occur at lower
concentrations among species or other life stages that are more sensitive
than the earliest life-cycle stages of the rainbow trout. The data base for
saltwater species is limited to one test, and therefore, no statement can be
made concerning acute or chronic toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to chloroform through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tions should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in incremental increase of cancer risk
over the lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for chloroform are 5.7 M-g/L for ingestion of con-
taminated water and organisms and 470 |ig/L for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
69

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70

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2-CHLOROPHENOL
95-57-8
CRITERIA
Aquatic Life The available data for 2-chlorophenol indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 4,380 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No definitive data are available concerning the chronic toxic-
ity of 2-chlorophenol to sensitive freshwater aquatic life, but flavor
impairment occurs in one species of fish at concentrations as low as
2,000 ng/L.
No saltwater organisms have been tested with 2-chlorophenol, and
therefore, no statement can be made concerning acute or chronic toxicity.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60890). Recalcul-
ated IRIS values for 2-chlorophenol are 120 ng/L for ingestion of
contaminated water and organisms and 400 (ig/L for ingestion organisms
only. According to available organoleptic data, the estimated level is 0.1
Hg/L to control undesirable taste and odor qualities of ambient water. Or-
ganoleptic data do have limitations as a basis for establishing a water
quality criterion but do not have a demonstrated relationship to poten-
tially adverse effects on human health.
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22,1992)
See Appendix C for Human Health Methodology.
71

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72

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CHLOROPHENOXY HERBICIDES
2,4-D 94-75-7; 2,4,5-TP 93-72-1
CRITERIA
2,4-D 100 jig/L for domestic water supply (health).
2,4,5-TP 10 ng/L for domestic water supply (health).
Rationale Two widely used herbicides are 2,4-D (2,4-dichlorophenoxy) acetic acid and
2,4,5-TP (silvex) 2-(2,4,5-trichlorophenoxy) propionic acid. These com-
pounds may exhibit different herbicidal properties, but all are hydolyzed
rapidly to the corresponding acid in the body.
The subacute oral toxicity of chlorophenoxy herbicides has been inves-
tigated in a number of animals. The dog was found to be sensitive and
often displayed mild injury in response to doses of 10 mg/kg/day for 90
days and serious effects from doses of 20 mg/kg/day for 90 days. The no-
effect level of 2,4-D is 0.5 mg/kg/day in the rat and 8.0 mg/kg/day in the
dog.
Table 1.—Derivation of approval limits (AL) for chlorophenoxy herbicides.

LOWEST LONG
TERM LEVELS
WITH MINIMAL OR
NO EFFECTS
CALCULATED MAXIMUM SAFE
LEVELS FROM ALL
SOURCES OF EXPOSURE
WATER
Compound
Species
me/kc/daya
Safely
facior(X)
mg/kg/day
mg/man/davb
9c ot
Safe level
AL
mg/lc
2,4-D ....
2.4,5-TP...
Rat
Dog
Rat
Dog
0.5
8.0
2.6
0.9
1/500
1/500
1/500
1/500
0.1
0.016
0.005
0.002
7.0
1.12"
0.35
0.14d
20
20
0.1
0.01
'Assume weight of rai = 0 3 kg and of dog = 10 tg. assume average daily food consumption of rat = 0 05 kg
and of dog = 0 2 kg
p Assume average weight of human jduii = 70 kg
v Assume average daily intake of water for man = 2 liters
dChosen as basis on which to derive AL
Table 1 illustrates the derivation of the criteria for the two
chlorophenoxy herbicides. The long-term, no-effect levels (mg/kg/day)
are listed for the rat and the dog. These values are adjusted by a factor of
1/500 for 2,4-D and 2,4,5-TP. The safe levels are then readjusted to reflect
total allowable intake per person. Since little 2,4-D or 2,4,5-TP is expected
to occur in foods, 20 percent of the safe exposure level can reasonably be
allocated to water without jeopardizing the health of the consumer.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
73

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CHLORPYRIFOS
2921-88-2
CRITERIA
Aquatic Life Freshwater— 4-day average of 0.041 ng/L
1-hour average of 0.083 ng/L
Saltwater — 4-day average of 0.0056 |xg/L
1-hour average of 0.011 ng/L
Summary	The acute values for 18 freshwater species in 15 genera range from
0.11 n-g/L for an amphipod to greater than 806 fig/L for 2 fishes and a
snail. The bluegill is the most sensitive fish species with an acute value of
10 M-g/L, but 7 intervetebrate genera are more sensitive. Smaller organisms
seem to be more acutely sensitive than larger ones.
Chronic toxicity data are available for 1 freshwater species, the fathead
minnow. Unacceptable effects occurred in second generation larvae at 0.12
Hg/L, which was the lowest concentration tested. The resulting acute-
chronic ratio was greater than 1,417.
Little information is available on the toxicity of chlorpyrifos to fresh-
water plants, although algal blooms frequently follow field applications.
The only available bioconcentration test on chlorpyrifos with a freshwater
species (the fathead minnow) resulted in a bioconcentration factor of 1,673.
The acute toxicity of chlorpyrifos has been determined for 15 species of
saltwater animals in 12 genera with the acute values ranging from 0.01
M-g/L for the Korean shrimp, Palaemon macrodactylus, to 1,911 M-g/L for lar-
vae of the eastern oyster, Crassostrea virginica. Arthropods are particularly
sensitive to chlorpyrifos. Among the 10 species of fish tested, the 96-hour
LC50s range from 0.58 fig/L for striped bass to 520 M-g/L for gulf toadfish.
Fish larvae are more sensitive than other life stages. Growth of the mysid,
Mysidopsis bahia, was reduced at 0.004 fig/Lin a life-cycle test. In early life-
stage tests, the California grunion, Leuresthes tenuis, was the most sensitive
of the six fishes, with growth being reduced at 0.30 M-g/L. Of the seven
acute-chronic ratios that have been determined with saltwater species, the
five lowest range from 2.388 to 12.50, whereas the highest is 228.5.
Concentrations of chlorpyrifos affecting six species of saltwater phyto-
plankton range from 138 to 10,000 fig/L. Bioconcentration factors (BCFs)
ranged from 100 to 5,100 when the gulf toadfish was exposed to concentra-
tions increasing from 1.4 to 150 M-g/L. Steady-state BCFs averaged from
100 to 757 for five fishes exposed in early life-stage tests.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of
chlorpyrifos does not exceed 0.041 ng/L more than once every three years
on the average, and if the one-hour average concentration does not exceed
0.083 M-g/L more than once every three years on the average.
75

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The procedures described in the guidelines also indicate that, except
possibly where a locally important species is very sensitive, saltwater
aquatic organisms and their uses should not be affected unacceptably if the
four-day average concentration of chlorpyrifos does not exceed 0.0056
Hg/L more than once every three years on the average, and if the one-hour
average concentration does not exceed 0.011 |xg/L more than once every
three years on the average.
In the Agency's best scientific judgment, three years is the average
amount of time aquatic ecosystems should be provided between excur-
sions. The resiliences of ecosystems and their abilities to recover differ
greatly, however, and site-specific allowed excursion frequencies can be es-
tablished if adequate justification is provided.
When designing waste treatment facilities, criteria for developing
water quality-based permit limits must be applied to an appropriate
wasteload allocation model. Dynamic models are preferred for the applica-
tion of these criteria. Limited data or other considerations might make
their use impractical, in which case one must rely on a steady-state model.
(51 F.R. 43665, December 3,1986)
See Appendix A for Aquatic Life Methodology.

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CHROMIUM
CRITERIA
Aquatic Life Chromium (VI) Freshwater — 4-day average of 11 ng/L
1-hour average ofl6 ng/L
Saltwater — 4-day average of 50 ng/L
1-hour average of 1,100 ng/L
Summary
Chromium (VI) 7440-47-3
Acute toxicity values for chromium (VI) are available for freshwater ani-
mal species in 27 genera and range from 23.07 jxg/L for a cladoceran to
1,870,000 ng/L for a stonefly. These species include a wide variety of ani-
mals that perform a spectrum of ecological functions. All five tested
species of daphnids are especially sensitive. The few data that are available
indicate that the acute toxicity of chromium (VI) decreases as hardness and
pH increase.
The chronic value for both rainbow trout and brook trout is 264.6 mg/ L,
which is much lower than the chronic value of 1,987 (ig/L for the fathead
minnow. The acute-chronic ratios for these three fishes range from 18.55 to
260.8. In all three chronic tests, a temporary reduction in growth occurred
at low concentrations. Six chronic tests with five species of daphnids gave
chronic values that range from 2.5 to 40 fxg/L; the acute-chronic ratios
range from 1.130 to 9.680. Except for the fathead minnow, all the chronic
tests were conducted in soft water. Green algae are quite sensitive to chro-
mium (VI). The bioconcentration factor obtained with rainbow trout is less
than 3. Growth of chinook salmon was reduced at a measured concentra-
tion of 16 ^/L.
The acute toxicity of chromium (VI) to 23 saltwater vertebrate and in-
vertebrate species ranges from 2,000 ng/L for a polychaete worm and a
mysid to 105,000 jxg/L for the mud snail. The chronic values for a poly-
chaete range from 13 to 36.74 ng/L, whereas that for a mysid is 132 ng/L.
The acute-chronic ratios range from 15.38 to 238.5. Toxicity to macroalgae
was reported at 1,000 and 5,000 ng/L. Bioconcentration factors for chro-
mium (VI) range from 125 to 236 for bivalve molluscs and polychaetes.
Chromium (III) 1308-14-1
Acute values for chromium (III) are available for 20 freshwater animal spe-
cies in 18 genera ranging from 2,221 ^/L for a mayfly to 71,060 ng/L for a
caddisfly. Hardness has a significant influence on toxicity, with chromium
(III) being more toxic in soft water.
A life-cycle test with Daphnia magna in soft water gave a chronic value
of 66 ng/L. In a comparable test in hard water, the lowest test concentra-
tion of 44 ng/L inhibited reproduction of D. magna, but this effect may
have resulted from ingested precipitated chromium. In a life-cycle test
with the fathead minnow in hard water, the chronic value was 1,025 ng/L.
77

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Toxicity data are available for only two freshwater plant species: a con-
centration of 9,900 ng/L inhibited growth of roots of Eurasian
watermilfoil, and a freshwater green alga was affected by a concentration
of 397 ng/L in soft water. No bioconcentration factor has been measured
for chromium (III) with freshwater organisms.
Only two acute values are available for chromium (III) in saltwater
10,300 ng/L for the eastern oyster and 31,500 ng/L for the mummichog. In
a chronic test, effects were not observed on a polychaete worm at 50,400
jxg/L at pH=7.9, but acute lethality occurred at pH=4.5. Bioconcentration
factors for saltwater organisms and chromium (HI) range from 86 to 153,
similar to the bioconcentration factors for chromium (VI) and saltwater
species.
NATIONAL CRITERIA
Aquatic Life
Chromium (VI)
The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of chro-
mium (VI) does not exceed 11 ng/L more than once every three years on
the average, and if the one-hour average concentration does not exceed 16
Hg/L more than once every three years on the average.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of chromium (VI) does not exceed 50 ng/L
more than once every three years on the average, and if the one-hour aver-
age concentration does not exceed 1,100 jig/L more than once every three
years on the average. Data suggest that the acute toxicity of chromium (VI)
is salinity dependent; therefore, the one-hour average concentration might
be underprotective at low salinities.
Chromium (III)
The procedures described in the guidelines indicate that, except possibly
where a locally important species is very sensitive, freshwater aquatic or-
ganisms and their uses should not be affected unacceptably if the four-day
average concentration (in ng/L) of chromium (III) does not exceed the nu-
merical value given by
e(0.8190[ln(hardness)]+1.561)
more than once every three years on the average, and if the one-hour
average concentration (in ng/L) does not exceed the numerical value given
by
e(0.8190[ln (hardness)]+3.688)
more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaC03, the four-day average
concentrations of chromium (III) are 120, 210, and 370 ng/L, respectively,
and the one-hour average concentrations are 980,1,700, and 3,100 ng/L.
78

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No saltwater criterion can be derived for chromium (III), but 10,300
Hg/L is the EC50 for eastern oyster embryos, whereas 50,400 ng/L did not
affect a polychaete worm in a life-cycle test.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average time needed for an unstressed sys-
tem to recover from a pollution event in which exposure to chromium
exceeds the criterion. For example, a stressed system (one in which several
outfalls occur in a limited area) would be expected or require more time for
recovery. The resilience of ecosystems and their ability to recover differ
greatly, however, and site-specific criteria can be established if adequate
justification is provided.
In designing waste treatment facilities, criteria must be applied to an
appropriate wasteload allocation model; dynamic models are preferred.
Limited data or other factors may make use of these models impractical, in
which case one should rely on a steady-state model. The Agency recom-
mends the interim use of 1Q5 for IQIO for Criterion Maximum
Concentration design flow and 7Q5 or 7Q10 for the Criterion Continuous
Concentration design flow in steady-state models for unstressed and
stressed systems, respectively. These matters are discussed in more detail
in EPA's "Technical Support Document for Water Quality-Based Toxics
Control."
Human Health Human health criteria have been withdrawn for this compound (see 57
F.R. 60885, December 22, 1992). Although the human health criteria are
withdrawn, EPA published a document for this compound that may con-
tain useful human health information. This document was originally
noticed in 45 F.R. 79331, November 28,1980.
(45 F.R. 79318 November 28,1980) (50 F.R. 30784, July 29,1985)
(57 F.R. 60848, December 22,1992)
See Appendix A for Aquatic Life Methodology.
79

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80

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COLOR
CRITERIA
For aesthetic purposes, waters shall be virtually free from
substances producing objectionable color;
The source of supply should not exceed 75 color units on the
platinum-cobalt scale for domestic water supplies; and
Increased color (in combination with turbidity) should not reduce
the depth of the compensation point for photosynthetic activity by
more than 10 percent from the seasonally established norm for
aquatic life.
Introduction
Degradation processes in the natural environment are the principal con-
tributors to color in water. Although colloidal forms of iron and
manganese occasionally color water, the most common causes of color in
water are complex organic compounds originating from the decomposi-
tion of naturally occurring organic matter. Sources of organic matter
include materials in the soil that were generated by humans such as tan-
nins, human acid, humates, and decayed material from plankton and other
aquatic plants. Industrial discharges may contribute similar compounds
from, for example, the pulp and paper and tanning industries. Other in-
dustrial discharges, such as those from certain textile and chemical
processes, may contain brightly colored substances (see Table 1).
Table 1.—Maximum color of surface waters that have been used as industrial water
supplies.
INDUSTRY OR INDUSTRIAL USE
Boiler makeup water 	
Cooling water 	
Pulp and paper water 	
Chemical and allied products water
Petroleum products water	
COLOR UNITS
1,200
1,200
360
500
25
Surface waters may appear colored because of suspended matter,
which comprises turbidity. Such color is referred to as "apparent color"
and is differentiated from true color caused by colloidal human materials.
Natural color is reported in color "units" that generally are determined by
the platinum-cobalt method.
No general agreement exists as to the chemical composition of natural
color, and in fact, the composition may vary chemically from place to
place. Examined color-causing colloids have been characterized as aro-
matic, polyhydroxy, methoxy carboxylic acids. Color-causing constituents
were characterized as being dialyzable and composed of aliphatic, poly-
hydroxyl carboxylic acids with molecular weights varying from less than
200 to approximately 400. The colloidal fraction of color exists in the 3.5 to
10 mu diameter range. Other characteristics of color observed in laboratory
studies of natural waters were summarized as follows: color is caused by
81

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light scattering and fluorescence rather than absorption of light energy,
and pH affects both particle size of the color-causing colloids and the in-
tensity of color itself.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
82

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*COPPER
7440-50-8
CRITERIA
Aquatic Life Not to exceed 2.9 jig/Lin salt water.
Freshwater criteria are hardness dependent. See text.
Summary	Acute toxicity data are available for species in 41 genera of freshwater ani-
mals. At a hardness of 50 mg/L, the genera range in sensitivity from 16.74
M-g/L for Ptychocheilus to 10,240 ng/L for Acroneuria. Data for eight species
indicate that acute toxicity decreases as hardness increases. Additional
data for several species indicate that toxicity also decreases with increases
in alkalinity and total organic carbon.
Chronic values available for 15 freshwater species range from 3.873
Hg/L for brook trout to 60.36 ng/L for northern pike. Fish and invertebrate
species seem to be almost equally sensitive to the chronic toxicity of cop-
per.
Toxicity tests on copper conducted with a wide range of freshwater
plants indicate sensitivities similar to those of animals. Complexing effects
of the test media and a lack of good analytical data make it difficult to inter-
pret and apply these results. Protection of animal species, however, appears
to offer adequate plant protection. Bioconcentrations of copper are light in
edible portion of freshwater aquatic species.
Saltwater animals' acute sensitivities to copper range from 5.8 fig/L for
the blue mussel to 600 ng/L for the green crab. A chronic life-cycle test has
been conducted with a mysid; adverse effects were observed at 77 ng/L
but not at 38 ng/L, which resulted in an acute-chronic ratio of 3.346. Sev-
eral saltwater algal species have been tested, and effects were observed
between 5 and 100 ng/L. Oysters can bioaccumulate copper up to 28,200
times and become bluish green, apparently without significant mortality.
In long-term exposures, the bay scallop was killed at 5 M-g/L.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Uses" indicate that — except possibly where a locally important species is
very sensitive — freshwater aquatic organisms and their uses should not
be affected unacceptably if the four-day average concentration (in |xg/L) of
copper does not exceed the numerical value given by
g(0.8545[ln(hardness)]-1.465)
more than once every three years on the average, and if the one-hour
average concentration (in ng/L) does not exceed the numerical value given
by
e(0.9422pn(hardness)]-1.464)
•Indicates suspension, cancelation, or restriction by U.S. EPA Office of Pesticides and
Toxic Substances.
83

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more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaCO3, the four-day average
concentrations of copper are 6.5, 12, and 21 ng/L, respectively, and the
one-hour average concentrations are 9.2,18, and 34 ng/L.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the one-
hour average concentration of copper does not exceed 2.9 ng/L more than
once every three years on the average.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average time needed for an unstressed sys-
tem to recover from a pollution event in which exposure to copper exceeds
the criterion. For example, a stressed system (one in which several outfalls
occur in a limited area) would be expected to require more time for recov-
ery The resilience of ecosystems and their ability to recover differ greatly,
however, and site-specific criteria can be established if adequate justifica-
tion is provided.
In developing waste treatment facilities, criteria requires the selection
of an appropriate wasteload allocation model. Dynamic models are pre-
ferred for the application of these criteria. Limited data or other factors
may make their use impractical, in which case one should rely on a steady-
state model. The Agency recommends the interim use of 1Q5 or 1Q10 for
Criterion Maximum Concentration design flow and 7Q5 or 7Q10 for the
Criterion Continuous Concentration design flow in steady-state models
for unstressed and stressed systems respectively. These matters are dis-
cussed in more detail in EPA's "Technical Support Document for Water
Quality-Based Toxics Control."
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60890). The recal-
culated IRIS values for copper is 1,300 n-g/L for ingestion of contaminated
water and organisms. Using available organoleptic data, the estimated
level is 1 mg/L for controlling undesirable taste and odor quality of ambi-
ent water. Organoleptic data as a basis for establishing a water quality
criteria have limitations and no demonstrated relationship to potentially
adverse effects on human health.
(45 F.R. 79318 November 28,1980) (50 F.R. 30784, July 29, 1985)
(57 F.R. 60890, December 22,1992)
See Appendix A for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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CYANIDE
57-12-5
Freshwater— 4-day average of 52 ng/L
1-hour average of 22 fig/L
Saltwater — 1-hour average of 1.0 (ig/L
Data on the acute toxicity of free cyanide (the sum of cyanide present as
HCN and CN-, expressed as CN) are available for a wide variety of fresh-
water species that are involved in diverse community functions. In tests,
the acute sensitivities ranged from 44.73 |xg/L to 2,490 ng/L, but all of the
species with acute sensitivities above 400 ^g/L were invertebrates. A long-
term survival and a partial and life-cycle test with fish gave chronic values
of 13.57, 7.849, and 16.39 ng/L, respectively. Chronic values for two fresh-
water invertebrate species were 18.33 and 34.06 ng/L. Freshwater plants
were affected at cyanide concentrations ranging from 30 ng/L to 26,000
M-g/L-
Free cyanide's acute toxicity to saltwater species ranged from 4.893
jig/L to 10,000 ng/L; invertebrates were both the most and least sensitive
species. In an early life-stage test with the sheepshead minnow, long-term
survival gave a chronic value of 36.12 ng/L. Long-term survival in a mysid
life-cycle test resulted in a chronic value of 69.71 ng/L. Tests with the red
macroalga, Champia parvula, showed cyanide toxicity at 11 to 25 ng/L, but
other species were affected at concentrations up to 3,000 [ig/L.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of cya-
nide does not exceed 5.2 ng/L more than once every three years on the
average, and if the one-hour average concentration does not exceed 22
Hg/L more than once every three years on the average.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the one-
hour average concentration of cyanide does not exceed 1.0 ng/L more than
once every three years on the average.
EPA believes that a measurement such as free cyanide would provide a
more scientifically correct basis upon which to establish criteria for cya-
nide. The criteria were developed on this basis. However, at this time EPA
has approved no methods for such a measurement to implement the cri-
teria through Agency and State regulatory programs.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average amount of time it will take an un-
stressed system to recover from a pollution event in which exposure to
85
CRITERIA
Aquatic Life
Summary

-------
cyanide exceeds the criterion. A stressed system, for example — one in
which several outfalls occur in a limited area — would be expected to re-
quire more time for recovery. The resilience of ecosystems and their ability
to recover differ greatly, however, and site-specific criteria may be estab-
lished if adequate justification is provided.
In designing waste treatment facilities, criteria must be applied to an
appropriate wasteload allocation model; dynamic models are preferred.
Limited data or other factors may make use of these models impractical, in
which case one should rely on a steady-state model. The Agency recom-
mends the interim use of 1Q5 or 1Q10 for Criterion Maximum
Concentration design flow and 7Q5 or 7Q10 for the Criterion Continuous
Concentration design flow in steady-state models for unstressed and
stressed systems, respectively. These matters are discussed in more detail
in EPA's "Technical Support Document for Water Quality-Based Toxics
Control."
Human Health Published human health criteria were recalculated using Integrated Risk
Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60911). Recalculated IRIS values for cyanide are 700 (ig/L for in-
gestion of contaminated water and organisms and 220,000 ng/L for
ingestion of contaminated aquatic organisms only.
(45 F.R. 79318 November 28,1980) (50 F.R. 30784, July 29,1985)
(57 F.R. 60911, December 22,1992)
See Appendix A for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
86

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DDT AND METABOLITES
72-54-8
CRITERIA
Aquatic Life 24-hour average for freshwater and saltwater is 0.001 ng/L .
Not to exceed at anytime 1.1 ng/L in fresh water or 0.13 ng/L in salt water.
DDT	The criterion for DDT and its metabolites to protect freshwater aquatic life
as derived using the guidelines is 0.0010 jig/L as a 24-hour average. The
concentration should not exceed 1.1 [ig/L at any time.
The criterion for DDT and its metabolites to protect saltwater aquatic
life as derived using the guidelines is 0.0010 ng/L as a 24-hour average.
The concentration should not exceed 0.13 ng/L at any time.
TDE	The available data for TDE indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 0.6 ng/L and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning TDE's chronic toxicity to sensitive
freshwater aquatic life.
The available data for TDE indicate that acute toxicity to saltwater
aquatic life occurs at concentrations as low as 3.6 |xg/L and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning TDE's chronic toxicity to sensitive
saltwater aquatic life.
DDE	The available data for DDE indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 1,050 ng/L and would occur
at lower concentrations among species that are more sensitive than those
tested. No data are available concerning DDE's chronic toxicity to sensitive
freshwater aquatic life.
The available data for DDE indicate that acute toxicity to saltwater
aquatic life occurs in concentrations as low as 14 ng/L and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning DDE's chronic toxicity to sensitive
saltwater aquatic life.
Human Health The ambient water concentration should be zero, based on the non-
threshold assumption for this chemical, for the maximum protection of
human health from the potential carcinogenic effects of exposure to DDT
through ingestion of contaminated water and contaminated aquatic organ-
isms. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in incremental increase of cancer risk
over a lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 RR. 60848).
The recalculated IRIS value for DDT is 0.00059 ng/L for both ingestion of
water and organisms, and ingestion of contaminated aquatic organisms
only. IRIS values are based on a 10"6 risk level for carcinogens.
87

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The recalculated IRIS value for TDE is 0.00083 ng/L for both ingestion
of contaminated water and organisms and for ingestion of contaminated
aquatic organisms only.
The recalculated IRIS value for DDE is 0.00059 ng/L for both ingestion
of contaminated water and organisms and for ingestion of contaminated
aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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DEMETON
8065-48-3
CRITERIA
Aquatic Life 0.1 M-g/L for both freshwater and saltwater aquatic life.
Rationale	Static LC50 bioassays yielded toxicity values for the phosphorus pesticide
demeton and for carp, goldfish, fathead minnow, channel catfish, guppy,
rainbow trout, and bluegill ranging from 70 M-g/L to 15,000 fig/L. These
tests demonstrate a sharp division in species sensitivity, with the bluegill,
Lepomis macrochirus; rainbow trout, Oncorhynchus mykiss; and guppy,
Poecilia reticulata being susceptible to lower concentrations, while the re-
maining species were comparatively resistant.
Bluegills with a 24-hour LC50 of 70 ^g/L were the most sensitive fish.
Acute toxicity values reported for invertebrates range from 10 to 100,000
M-g/L.*
Static LC50 data for invertebrate Gammarus fasciatus yield 24- and 96-
hour LC50 values of 500 fig/L and 27 ng/L, respectively. Studies indicate
residual effects of AChE inhibition from exposure to demeton. The few
data on toxicity of demeton to marine organisms includes a 48-hour EC50
of 63 ng/L for the pink shrimp, Penaeus duorarum, and a 24-hour LC50 of
550 ng/L for the spot, Leiostomus xanthurus.
The criterion must be based partly on the fact that all organophos-
phates inhibit the production of the AChE enzyme. Demeton is unique,
however, in that the persistence of its AChE-inhibiting ability is greater
than that of 10 other common organophosphates. Because such inhibition
may be additive with repeated exposures and may be compounded by any
of the organophosphates, it is recommended that a criterion for demeton
be based primarily on its enzyme-inhibiting potential. A criterion of 0.1
Hg/L demeton for freshwater and marine aquatic life is recommended,
since that concentration will not be expected to significantly inhibit AChE
over a long period. In addition, the criterion recommendation is in close
agreement with the criteria for the other organophosphates.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
'Crustaceans and insect larvae were considerably more sensitive overall to demeton
than molluscs and tubifex worms.
89

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90

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DICHLOROBENZENES
25321-22-6
CRITERIA
Aquatic Life
Human Health
1.2-Dichlorobenzene	95-50-1
For the maximum protection of human health from the potential effects of
exposure to 1,2-dichlorobenzene, the human health criteria were recalcul-
ated using the Integrated Risk Information System (IRIS) to reflect
available information as of 12/92 (57 F.R. 60848). The recommended cri-
teria are 2,700 (ig/L for ingestion of contaminated water and organisms
and 17,000 |xg/L for ingestion of contaminated organisms only.
1.3-Dichlorobenzene	541-73-1
For the maximum protection of human health from the potential effects of
exposure to 1,3-dichlorobenzene, the recommended criteria are 400 ng/L
for ingestion of contaminated water and organisms and 2,600 |xg/L for in-
gestion of contaminated organisms only.
1.4-Dichlorobenzene	106-46-7
For the maximum protection of human health from the potential effects of
exposure to 1,4-dichlorobenzene, the recommended criteria are 400 ng/L
for ingestion of contaminated water and organisms and 2,600 (ig/L for in-
gestion of contaminated organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
The available data for dichlorobenzenes indicate that acute and chronic
toxicity to freshwater aquatic life occur at concentrations as low as 1,120
and 763 [ig/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested.
The available data for dichlorobenzenes indicate that acute toxicity to
saltwater aquatic life occurs at concentrations as low as 1,970 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
dichlorobenzenes to sensitive saltwater aquatic life.
-91

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92

-------
DICHLOROBENZIDINE
91-94-1
CRITERIA
Aquatic Life The data base available for dichlorobenzidines and freshwater organisms
is limited to one test on bioconcentration of 3,3-dichlorobenzidine; there-
fore, no statement can be made concerning acute or chronic toxicity.
No saltwater organisms have been tested with any dichlorobenzidine;
therefore, no statement can be made concerning acute or chronic toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to dichlorobenzidine through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for 3,3-dichlorobenzidine are 0.04 fig/L for inges-
tion of contaminated water and organisms and 0.077 ng/L for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
(45 F.R. 79318, November 28, 1980) (57 F.R. 60848, December 22, 1992)
See Appendix C for Human Health Methodology.
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DICHLOROETHYLENES
25323-30-3
CRITERIA
Aquatic Life The available data for dichloroethylenes indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 11,600 |xg/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No definitive data are available concerning the chronic
toxicity of dichloroethylenes to sensitive freshwater aquatic life.
The available data for dichloroethylenes indicate that acute and
chronic toxicity to saltwater aquatic life occurs at concentrations as low as
224,000 fig/L and would occur at lower concentrations among species that
are more sensitive than those tested. No data are available concerning the
chronic toxicity of dichloroethylenes to sensitive saltwater aquatic life.
Human Health
1.1-Dichloroethylene	75-35-4
For the maximum protection of human health from the potential carcino-
genic effects of exposure to 1, 1-dichloroethylene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for 1,1-dichloroethylene are 0.057 ng/L for inges-
tion of contaminated water and organisms and 3.2 fig/L for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
1.2-Dichloroethylene	156-60-5
Human health criteria were recalculated using IRIS to reflect available data
as of 12/92 (57 F.R. 60890). Recalculated IRIS values for 1,2-trans-
dichloroethylene are 700 pig/L for ingestion of contaminated water and
organisms. IRIS values are based on a 10'6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22,1992)
See Appendix C for Human Health Methodology.
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2,4-DICHLOROPHENOL
120-83-2
CRITERIA
The available data for 2,4-dichlorophenol indicate that acute and chronic
toxicity to freshwater aquatic life occurs at concentrations as low as 2,020
and 365 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested. Mortality to early
life stages of one species of fish occurs at concentrations as low as 70 ng/L.
Only one test has been conducted with saltwater organisms and 2,4-
dichlorophenol, and therefore, no statement can be made concerning acute
or chronic toxicity.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for 2,4-dichlorophenol are 93.0 ng/L for ingestion of
contaminated water and organisms and 790 ng/L for ingestion of contami-
nated aquatic organisms only.
Aquatic Life
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
97

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98

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DICHLOROPROPANE
26638-19-7
DICHLOROPROPENE
26952-23-8
CRITERIA
Aquatic Life The available data for dichloropropane indicate that acute and chronic tox-
icity to freshwater aquatic life occurs at concentrations as low as 23,000
and 5,700 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested. Acute and chronic
toxicity to saltwater aquatic life occur at concentrations as low as 10,300
and 3,040 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested.
The available data for dichloropropene indicate that acute and chronic
toxicity to freshwater aquatic life occurs at concentrations as low as 6,060
and 244 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested. Acute toxicity to
saltwater aquatic life occurs at concentrations as low as 790 fig/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
dichloropropene to sensitive saltwater aquatic life.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60890). Recalcul-
ated IRIS values for 1,2-dichloropropane is 0.52 jig/L per ingestion of
water and organisms and 39.0 ng/L per ingestion of organisms only.
For the protection of human health from the toxic properties of
dichloropropenes ingested through water and contaminated aquatic or-
ganisms, the ambient water criterion is 87 ng/L.
For the protection of human health from the toxic properties of
dichloropropenes ingested through contaminated aquatic organisms
alone, the ambient water criterion is 14.1 mg/L.
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22,1992)
See Appendix C for Human Health Methodology.
99

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100

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DIELDRIN
60-57-1
CRITERIA
Aquatic Life Freshwater— 24-hour average of 0.0019 ng/L
Never to exceed 2.5 ng/L
Saltwater — 24-hour average of 0.0019 |xg/L
Never to exceed 0.71 n-g/L
To protect freshwater aquatic life, the criterion for dieldrin is 0.0019
M-g/L as a 24-hour average. The concentration should not exceed 2.5 ng/L
at any time.
To protect saltwater aquatic life, the criterion as derived using the
guidelines is 0.0019 jig/L as a 24-hour average. The concentration should
not exceed 0.71 ng/L at any time.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to dieldrin through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tion should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in incremental increase of cancer risk
over the lifetime are estimated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for dieldrin are 0.00014 ng/L for ingestion of con-
taminated water and organisms and also for ingestion of contaminated
aquatic organisms only. IRIS values are based on a 10"6 risk level for carcin-
ogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
101

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102

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2,4-DIMETHYLPHENOL
105-67-9
CRITERIA
Aquatic Life The available data for 2,4-dimethylphenol indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 2,120 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
dimethylphenol to sensitive freshwater aquatic life.
No saltwater organisms have been tested with 2,4-dimethylphenol,
and therefore, no statement can be made concerning acute or chronic toxic-
ity.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect data available as of 12/92 (57 F.R. 60890). Recalcul-
ated IRIS values for 2,4-dimethylphenol are 540 ng/L for ingestion of
contaminated water and organisms and 2,300 |xg/L for ingestion of con-
taminated organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22,1992)
See Appendix C for Human Health Methodology.
103

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104

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DINITROTOLUENE
25321-14-6
CRITERIA
Aquatic Life
Human Health
-Dinitrotoluene 121-14-2
For the maximum protection of human health from the potential carcino-
genic effects of exposure to 2,4-dinitrotoluene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10"5,10"6, and 10"7. The corre-
sponding recommended criteria are 1.1 ptg/L, 0.11 ng/L, and 0.011 |xg/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 91 ng/L, 9.1
|xg/L, and 0.91 ng/L, respectively.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
The available data for 2,4-dinitrotoluene indicate that acute and chronic
toxicity to freshwater aquatic life occurs at concentrations as low as 330
and 230 fxg/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested.
The available data for 2,4-dinitrotoluene indicate that acute toxicity to
saltwater aquatic life occurs at concentrations as low as 590 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
dinitrotoluenes to sensitive saltwater aquatic life but a decrease in algal
cell numbers occurs at concentrations as low as 370 ng/L.
105

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106

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DIPHENYLHYDRAZINE
122-66-7
CRITERIA
Aquatic Life
1,2-Diphenylhydrazine
The available data for 1,2-diphenylhydrazine indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 270 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
1,2-diphenylhydrazine to sensitive freshwater aquatic life.
No saltwater organisms have been tested with 1,2-diphenylhydrazine,
and therefore, no statement can be made concerning its acute or chronic
toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to diphenylhydrazine through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10"5,10'6, and 10"7. The corre-
sponding recommended criteria are 422 ng/L, 42 ng/L, and 4 ng/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 5.6 [ig/L, 0.56
(ig/L, and 0.056 ng/L, respectively.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92
(57 E.R. 60913). Recalculated IRIS values for 1,2-diphenylhydrazine are
0.040 ng/L for ingestion of contaminated water and organisms and 0.54
Hg/L for ingestion of contaminated aquatic organisms only. IRIS values are
based on a 10"6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60913, December 22,1992)
See Appendix C for Human Health Methodology.
107

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108

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DI-2-ETHYLHEXYL PHTHALATE
117-81-7
CRITERIA
Aquatic Life The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" do not allow for the derivation of national criteria for di-2-
ethylhexyl phthalate (DEHP), based on the available test information.
Limited data indicate that acute toxicity occurs to freshwater aquatic
life at a concentration as low as 2,100 ^ig/L, which is above the reported
solubility limit for DEHP. Chronic toxicity occurs to one freshwater species
at a concentration as low as 160 ng/L.
Toxicity data for DEHP and saltwater life is limited. However, if their
chronic sensitivity to DEHP is similar to that of freshwater aquatic life, ad-
verse effects on individual species might be expected at sl60 ng/L. An
ecosystem process, ammonia flux, has been shown to be reduced at
15.5 n-g/L in summer months.
Human Health Refer to phthalate esters.
109

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110

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DISSOLVED OXYGEN
7782-44-7
National
Criteria		
The national criteria for ambient dissolved oxygen concentrations for the
protection of freshwater aquatic life are presented in Table 1. The criteria
are derived from the production impairment estimates that are based pri-
marily upon growth data and information on temperature, disease, and
pollutant stresses. The average dissolved oxygen concentrations selected
are values 0.5 mg/L above the slight production impairment values and
represent values between no production impairment and slight production
impairment. Each criterion may thus be viewed as an estimate of the
threshold concentration below which detrimental effects are expected.
Table 1 —Water quality criteria for ambient dissolved oxygen concentration

COLDWATER CRITERIA
warmwater criteria
EARLY LIFE OTHER LIFE
STAGES'1 STAGES
EARLY LIFE OTHER LIF-E
STAGES-' STAGES
30 Day Mean 	
7 Day Mean 	
7 Day Mean Minimum
1 Day Minimum4-5 ....
'NA'
9.5 (6.5)
NA
8.0 (5.0)
-P* Lft Z 9s
© © > I*
NA
6.0
NA
5.0
5.5
NA
4.0
3.0
'These are water column concentrations recommended to achieve the required intergravel dissolved oxygen con-
centrations shown in parentheses The 3 mg/L differential is discussed in the criteria document For species that have
earlv life stages exposed directly to the water column, the figures in parentheses apply
•Includes all embryonic and larval stages and all juvenile lonns to 10 days following hatching
*NA (not applicable)
4For highly manipulatable discharges, further restrictions apply
^All minimum values should be considered as instantaneous concentrations to be achieved at all times
Criteria for coldwater fish are intended to apply to waters containing a
population of one or more species in the family Salmonidae or to waters
containing other coldwater or coolwater fish deemed by the user to be
closer to salmonids in sensitivity than to most warmwater species. Al-
though the acute lethal limit for salmonids is at or below 3 mg/L, the
coldwater minimum has been established at 4 mg/L because a significant
proportion of the insect species common to salmonid habitats are less tol-
erant of acute exposures to low dissolved oxygen than are salmonids.
Some coolwater species may require more protection than that afforded by
the other life-stage criteria for warmwater fish, and protecting sensitive
coolwater species with the coldwater criteria may be desirable.
Many States have more stringent dissolved oxygen standards for
cooler waters that contain either salmonids, nonsalmonid coolwater fish,
or the sensitive Centrarchidae, the smallmouth bass. The warmwater cri-
teria are necessary to protect early life stages of warmwater fish as
sensitive as channel catfish and to protect other life stages of fish as sensi-
tive as largemouth bass. Criteria for early life stages are intended to apply
only where and when these stages occur. These criteria represent dissolved
oxygen concentrations that EPA believes provide a reasonable and ade-
quate degree of protection for freshwater aquatic life.
The criteria do not represent assured no-effect levels. However, be-
cause the criteria represent worst case conditions (i.e, for wasteload
111

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allocation and waste treatment plant design), conditions will be better than
the criteria nearly all of the time at most sites. In situations where criteria
conditions are just maintained for considerable periods, the proposed cri-
teria represent some risk of production impairment. This impairment
would depend on innumerable other factors. If slight production impair-
ment or a small but undefinable risk of moderate impairment is
unacceptable, then one should use the "no production impairment" values
given in the document as means and the "slight production impairment"
values as minimum. Table 2 presents these concentrations.
The criteria represent dissolved oxygen concentrations believed to pro-
tect the more sensitive populations of organisms against potentially
damaging production impairment. The dissolved oxygen concentrations in
the criteria are intended to be protective at typically high, seasonal envi-
ronmental temperatures for the appropriate taxonomic and life-stage
classifications, temperatures that are often higher than those used in the re-
search from which the criteria were generated, especially for other than
early life stages.
Where natural conditions alone create dissolved oxygen concentra-
tions less than 110 percent of the applicable criteria means or minima or
both, the minimum acceptable concentration is 90 percent of the natural
concentration. These values are similar to those presented graphically and
to those calculated from 1972 Water Quality Criteria. Absolutely no anthro-
Table 2.—Dissolved oxygen concentrations (MG/L) versus quantitative level of effect.
1.	Salmonid Waters
a.	Embryo and Larval Stages
No Production Impairment	=	11 * (8)
Slight Production Impairment	=	9* (6)
Moderate Production Impairment	=	8* (5)
Severe Production Impairment	=	7* (4)
Limit to Avoid Acute Mortality	=	6* (3)
*Notc: These are water column concentrations recommendtd to achieve the required
intergravel dissolved oxygen concentrations shown in parentheses. The 3 mg/L dif-
ference is discussed in the criteria document.
b.	Other Life Stages
No Production Impairment	=	8
Light Production Impairment	=	6
Moderate Production Impairment	=	5
Severe Production Impairment	=	4
Limit to Avoid Acute Mortality	=	3
2.	Nonsalmonid Waters
a.	Early Life Stages
No Production Impairment	=	6.5
Slight Production Impairment	=	5.5
Moderate Production impairment	=	5
Severe Production Impairment	=	4.5
Limit to Avoid Acute Mortality	=	4
b.	Other Life Stages
No Production Impairment	=	6
Slight Production Impairment	=	5
Moderate Production Impairment	=	4
Severe Production Impairment	=	3.5
Limit to Avoid Acute Mortality	=	3
3. Invertebrates
No Production Impairment	=	8
Some Production impairment	=	5
Acute Mortality Limit	=	4

-------
pogenic dissolved oxygen depression in the potentially lethal area below
the one-day minima should be allowed unless special care is taken to as-
certain the tolerance of resident species to low dissolved oxygen.
If daily cycles of dissolved oxygen are essentially sinusoidal, a reason-
able daily average is calculated from the day's high and low dissolved
oxygen values. A time-weighted average may be required if the dissolved
oxygen cycles are decidedly nonsinusoidal. Determining the magnitude of
daily dissolved oxygen cycles requires at least two appropriately timed
measurements daily, and characterizing the shape of the cycle requires sev-
eral more appropriately spaced measurements.
Once a series of daily mean dissolved oxygen concentrations are calcu-
lated, an average of these daily means can be calculated (Table 3). For
embryonic, larval, and early life stages, the averaging period should not
exceed seven days. This short time is needed to adequately protect these
often short-duration, most sensitive life stages. Other life stages can proba-
bly be adequately protected by 30-day averages. Regardless of the
averaging period, the average should be considered a moving average
rather than a calendar-week or calendar-month average.
Table 3.—Sample calculations for determining daily means and seven-day mean
dissolved oxygen concentrations (30-day averages are calculated in a similar fashion
using 30 days data).
DAY
DISSOLVED OXYGEN (MG/L)
DAILY MAX DAILY M1N
DAILY MEAN
1 	
9.0
7.0
8.0
2 	
10.0
7.0
8.5
3 	
11.0
8.0
9.5"
4 	
12.03
8.0
9.5
5 	
10.0
8.0
9.0
6 	
11 0
9.0
10.0
7 	
12.0a
10.0
10.5C


57.0
65.0
1-day Minimum 	
7.0

7-day Mean Minimum 	
8.1

7-Hav Mean


9.3


"Above air saturation concentration (assumed to be 11 0 mg/L Tor Hits example)
b(1l 0 + 8.0)2
<(1! 0 + 10.0)2
The criteria have been established on the basis that the maximum dis-
solved oxygen value actually used in calculating any daily mean should
not exceed the air saturation value. This consideration is based primarily
on analysis of studies of cycling dissolved oxygen and the growth of large-
mouth bass, which indicated that high dissolved oxygen levels (6 mg/L)
had no beneficial effect on growth.
During periodic cycles of dissolved oxygen concentrations, minima
lower than acceptable constant exposure levels are tolerable so long as:
1.	The average concentration attained meets or exceeds the criterion;
2.	The average dissolved oxygen concentration is calculated as
recommended in Table 3; and
3.	The minima are not unduly stressful and clearly are not lethal.
A daily minimum has been included to make certain that no acute
mortality of sensitive species occurs as a result of lack of oxygen. Because
repeated exposure to dissolved oxygen concentrations at or near the acute
lethal threshold will be stressful and because stress can indirectly produce
113

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mortality or other adverse effects (e.g., through disease), the criteria are de-
signed to prevent significant episodes of continuous or regularly recurring
exposures to dissolved oxygen concentrations at or near the lethal thresh-
old. This protection has been achieved by setting the daily minimum for
early life stages at the subacute lethality threshold, by the use of a seven-
day averaging period for early life stages, by stipulating a seven-day mean
minimum value for other life stages, and by recommending additional lim-
its for manipulatable discharges.
The previous EPA criterion for dissolved oxygen published in "Quality
Criteria for Water" (1976) was a minimum of 5 mg/L (usually applied as a
7Q10), which is similar to the current criterion minimum except for other
life stages of warmwater fish that now allows a seven-day mean minimum
of 4 mg/L. The new criteria are similar to those contained in the 1968
"Green Book" of the Federal Water Pollution Control Federation.
The Criteria and Monitoring and Design Conditions
The acceptable mean concentrations should be attained most of the time,
but some deviation below these values would probably not cause signifi-
cant harm. Deviations below the mean will probably be serially correlated
and hence apt to occur on consecutive days. The significance of deviations
below the mean will depend on whether they occur continuously or in
daily cycles, the former being more adverse than the latter. Current knowl-
edge regarding such deviations is limited primarily to laboratory growth
experiments and by extrapolation to other activity-related phenomena.
Under conditions where large daily cycles of dissolved oxygen occur, it
is possible to meet the criteria mean values and consistently violate the
mean minimum criteria. Under these conditions, the mean minimum cri-
teria will clearly be the limiting regulation unless alternatives, such as
nutrient control, can dampen the daily cycles.
The significance of conditions that fail to meet the recommended dis-
solved oxygen criteria depend largely upon five factors: (1) the duration of
the event; (2) the magnitude of the dissolved oxygen depression; (3) the fre-
quency of recurrence; (4) the proportional area of the site failing to meet the
criteria; and (5) the biological significance of the site where the event occurs.
Evaluation of an event's significance must be largely case- and site-
specific. Common sense would dictate that the magnitude of the
depression would be the single most important factor in general, especially
if the acute value is violated. A logical extension of these considerations is
that the event must be considered in the context of the resolution level of
the monitoring or modeling effort. Evaluating the extent, duration, and
magnitude of an event must be a function of the spatial and temporal fre-
quency of the data. Thus, a single deviation below the criterion takes on
considerably less significance where continuous monitoring occurs than
where sampling is comprised of once-a-week grab samples. This is so be-
cause, based on continuous monitoring, the event is provably small; but
with the much less frequent sampling, the event is not provably small and
can be considerably worse than indicated by the sample.
The frequency of recurrence is of considerable interest to those model-
ing dissolved oxygen concentrations because the return period, or period
between recurrences, is a primary modeling consideration contingent
upon probabilities of receiving water volumes, waste loads, temperatures,
and so forth. It should be apparent that the return period cannot be iso-
lated from the other four factors discussed above. Ultimately, the question
of return period may be decided on a site-specific basis, taking into ac-
114

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count the other factors (duration, magnitude, areal extent, and biological
significance) mentioned above. Future studies of temporal patterns of dis-
solved oxygen concentrations, both within and between years, must be
conducted to provide a better basis for selection of the appropriate return
period.
In conducting wasteload allocation and treatment plant design compu-
tations, the choice of temperature in the models will be important.
Probably the best option would be to use temperatures consistent with
those expected in the receiving water over the critical dissolved oxygen
period for the biota.
The Criteria and Manipuiatable Discharges
If daily minimum DOs are perfectly serially correlated (i.e, if the annual
lowest daily minimum dissolved oxygen concentration is adjacent in time
to the next lower daily minimum dissolved oxygen concentration, and one
of these two minima is adjacent to the third lowest daily minimum dis-
solved oxygen concentration, and so on), then, to meet the seven-day mean
minimum criterion, more than three or four consecutive daily minimum
values below the acceptable seven-day mean minimum will not likely
occur. Unless the dissolved oxygen pattern is extremely erratic, it is also
unlikely that the lowest dissolved oxygen concentration will be apprecia-
bly below the acceptable seven-day mean minimum, or that daily
minimum values below the seven-day mean minimum will occur in more
than one or two weeks each year.
For some discharges, the distribution of dissolved oxygen concentra-
tions can be manipulated to varying degrees. Applying the daily minimum
to manipuiatable discharges would allow repeated weekly cycles of mini-
mum acutely acceptable dissolved oxygen values, a condition of
unacceptable stress, and possible adverse biological effect. For this reason,
the application of the one-day minimum criterion to manipuiatable dis-
charges must limit either the frequency of occurrence of values below the
acceptable seven-day mean minimum or must impose further limits on the
extent of excursions below the seven-day mean minimum. For such con-
trolled discharges, the occurrence of daily minima below the acceptable
seven-day mean minimum should be limited to three weeks per year or
the acceptable one-day minimum should be increased to 4.5 mg/L for
coldwater fish and 3.5 mg/L for warmwater fish. Such decisions could be
site-specific based upon the extent of control and serial correlation.
(51F.R. 22978, June 24,1986)
See Appendix A for Aquatic Life Methodology.
115

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116

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DISSOLVED SOLIDS AND SALINITY
CRITERIA 	
250 mg/L for chlorides and sulfates in domestic water supplies (welfare).
Introduction Dissolved solids and total dissolved solids, terms generally associated
with freshwater systems, consist of inorganic salts, small amounts of or-
ganic matter, and dissolved materials. The equivalent terminology in
Standard Methods is "filtrable residue." Salinity, an oceanographic term, is
not precisely equivalent to the total dissolved salt content but is related.
For most purposes, the terms "total dissolved salt content" and "salinity"
are equivalent. The principal inorganic anions dissolved in water include
the carbonates, chlorides, sulfates, and nitrates (principally in ground wa-
ters); the principal cations are sodium, potassium, calcium, and
magnesium.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
117

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ENDOSULFAN
115-29-7
CRITERIA
Aquatic Life The criterion to protect freshwater aquatic life as derived using the guide-
lines is 0.056 ng/L as a 24-hour average; the concentration should not
exceed 0.22 (ig/L at any time.
The criterion to protect saltwater aquatic life as derived using the
guidelines is 0.0087 n-g/L as a 24-hour average; the concentration should
not exceed 0.034 ng/L at any time.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcu-
late IRIS values for alpha-endosulfan, beta-endosulfan and endosulfan
sulfate are 0.93 ng/L for ingestion of contaminated water and organisms
and 2.0 n-g/L for ingestion of contaminated aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Mehtodology.
119

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*ENDRIN
72-20-8
CRITERIA
Aquatic Life The criterion to protect freshwater aquatic life exposed to endrin, as de-
rived using the guidelines, is 0.0023 ng/L as a 24-hour average; the
concentration should not exceed 0.18 pig/L at any time.
The criterion to protect saltwater aquatic life exposed to endrin, as de-
rived using the guidelines, is 0.0023 ng/L as a 24-hour average; the
concentration should not exceed 0.037 ng/L at any time.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for both endrin and endrin aldehyde are 0.76 M-g/L for
ingestion of contaminated water and organisms and 0.81 |o.g/L for inges-
tion of contaminated aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Mehtodology.
•Indicates suspension, cancelation, or restriction by U.S. EPA Office of Pesticides and
Toxic Substances.
121

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122

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ETHYLBENZENE
100-41-4
CRITERIA
Aquatic Life The available data for ethylbenzene indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 32,000 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No definitive data are available concerning the chronic
toxicity of ethylbenzene to sensitive freshwater aquatic life.
The available data for ethylbenzene indicate that acute toxicity to salt-
water aquatic life occurs at concentrations as low as 430 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of ethyl-
benzene to sensitive saltwater aquatic life.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for ethylbenzene are 3,100 ng/L for ingestion of
contaminated water and organisms and 29,000 ng/L for ingestion of con-
taminated aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60912, December 22, 1992)
See Appendix C for Human Health Methodology.
123

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124

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FLUORANTHENE
206-44-0
CRITERIA
Aquatic Life The available data for fluoranthene indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 3,980 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of
fluoranthene to sensitive freshwater aquatic life.
The available data for fluoranthene indicate that acute and chronic tox-
icity to saltwater aquatic life occur at concentrations as low as 40 and 16
Hg/L, respectively, and would occur at lower concentrations among spe-
cies that are more sensitive than those tested.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for fluoranthene are 300 ng/L for ingestion of
contaminated water and organisms and 370 ng/L for ingestion of contami-
nated organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22, 1992)
See Appendix C for Human Health Methodology.
125

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126

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GASES, TOTAL DISSOLVED
CRITERIA
Aquatic Life To protect freshwater and saltwater aquatic life, the total dissolved gas
concentrations in water should not exceed 110 percent of the saturation
value for gases at the existing atmospheric and hydrostatic pressures.
Rationale	Fish in water containing excessive dissolved gas pressure or tension are
killed when dissolving gases in their circulatory system come out of solu-
tion to form bubbles (emoli) that block the flow of blood through the
capillary vessels. In aquatic organisms this is commonly referred to as "gas
bubble disease." External bubbles (emphysema) also appear in the fins, on
the opercula, in the skin, and in other body tissues. Aquatic invertebrates
are also affected by gas bubble disease but usually at supersaturation lev-
els higher than those lethal to fish.
Percent saturation of water containing a given amount of gas varies
with the absolute temperature and the pressure. Because of the pressure
changes, percent saturation with a given amount of gas changes with the
water depth. Gas supersaturation decreases by 10 percent per meter of in-
crease in water depth due to hydrostratic pressure; a gas that is at 130
percent saturation at the surface would be at 100 percent saturation at 3
meters' depth. Compensation for altitude may be needed because a reduc-
tion in atmostpheric pressure changes the water/gas equilibrium,
resulting in changes in solubility of dissolved gases.
Total dissolved gas supersaturation can occur in several ways:
1.	Excessive biological activity: dissolved oxygen (DO)
concentrations can reach supersaturation as a result of excessive
algal photosynthesis. Algal blooms are often accompanied by
increased water temperatures, which further contribute to
supersaturation.
2.	Water spillage from hydropower dams causes supersaturation.
3.	Gas bubble disease may be induced by discharges from
power-generating and other thermal sources. Discharged water
becomes supersaturated with gases.
In recent years, gas bubble disease has been identified as a major prob-
lem affecting valuable stocks of salmon and trout in the Columbia River
system. The disease is caused by high concentrations of dissolved atmo-
spheric gas which enters the river's water during heavy spilling at
hydroelectric dams.
Field and laboratory reports result in several conclusions:
1. When either juvenile or adult salmonids are confined to shallow
water (1M), substantial mortality occurs at and above 115 percent
total dissolved gas saturation.
127

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2.	When either juvenile or adult salmonids are free to sound and
obtain hydrostatic compensation either in the laboratory or in the
field, substantial mortality still occurs when saturation levels (of
total dissolved gases) exceed 120 percent saturation.
3.	Using survival estimates made in the Snake River from 1966 to
1975, it is concluded that juvenile fish losses ranging from 40 to 95
percent do occur. A major portion of this mortality can be
attributed to fish exposure to supersaturation by atmospheric
gases during years of high flow.
4.	Juvenile salmonids subjected to sublethal periods of exposure to
supersaturation can recover when returned to normally saturated
water, but adults do not recover and generally die from direct and
indirect effects of the exposure.
5.	Some species of salmon and trout can detect and avoid
supersaturated waters; others may not.
6.	Higher survival was observed during periods of intermittent
exposure than during continuous exposure.
7.	In general, in acute bioassays, salmon and trout were less tolerant
than the nonsalmonids.
Interested individuals should review the original document for associ-
ated references and reports. This document is available from the National
Technical Information Service (NTIS). See Appendix F for ordering infor-
mation.
No differences are proposed in the criteria for freshwater and marine
aquatic life, as available data indicate little difference in overall tolerance
between saltwater and freshwater species.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
128

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GUTHION
86-50-0
CRITERIA
Aquatic Life .01 ng/L for freshwater and saltwater aquatic life.
Rationale	Four-day LC50 values for fish exposed to the organophosphate pesticide
range from 4 to 4,270 ng/L. Decreased spawning was documented in fat-
head minnows exposed to low levels during complete life-cycle exposures.
The estimated "safe" long-term concentration for this species is 0.3 |ig/L to
0.5 |xg/L.
Organophosphate pesticides inhibit the enzyme acetylcholinesterase
(AChE), which is essential to nerve impulse transport. Inhibition of 40 to
70 percent of fish brain AChE is usually fatal. Centrarchids are considered
one of the most sensitive fish to guthion.
Four-day LC50 values for aquatic invertebrates range from 0.10 ng/L
to 22.0 ng/L, indicating an overall greater sensitivity than fish and a nar-
rower spectrum of tolerance across species.
Results of toxicity studies with marine organisms indicate similar re-
sponses, with saltwater invertebrates exhibiting LC50s as low as 0.33
Hg/L.
A criterion level of 0.01 ng/L for guthion is based on the use of 0.1 as
an application factor, applied to the 96-hour LC50 of 0.1 (ig/L (for the am-
phipod, Gammarus) and a similar value of 0.3 (xg/L, exhibited by the
European shrimp.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
129

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130

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HALOETHERS
CRITERIA
Aquatic Life The available data for haloethers indicate that acute and chronic toxicity to
freshwater aquatic life occurs at concentrations as low as 360 and 122
M-g/L, respectively, and would occur at lower concentrations among spe-
cies that are more sensitive than those tested.
No saltwater organisms have been tested with any haloether, and
therefore, no statement can be made concerning acute or chronic toxicity.
Human Health Using the present guidelines, a satisfactory criterion cannot be derived at
this time because of insufficient available data for haloethers. See also
chloroalkyl ethers.
(45 F.R. 79318, November 28,1980)
See Appendix B for Aquatic Life Methodology.
131

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132

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HALOMETHANES
The available data for halomethanes indicate that acute toxicity to freshwa-
ter aquatic life occurs at concentrations as low as 11,000 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of
halomethanes to sensitive freshwater aquatic life.
The available data for halomethanes indicate that acute and chronic
toxicity to saltwater aquatic life occurs at concentrations as low as 12,000
and 6,400 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested. A decrease in algal
cell numbers occurs at concentrations as low as 11,500 ng/L.
For the maximum protection of human health from the potential carcino-
genic effects of exposure through the ingestion of contaminated water and
aquatic organisms for bromoform, dichlorobromomethane, and methylene
chloride, the ambient water concentration should be zero. However, zero
levels may not be attainable at the present time. Therefore, the levels that
may result in incremental increase of cancer risk over a lifetime are esti-
mated at 10"5,10"6, and 10"7.
Bromoform (Tribromomethane) 75-25-2
Human health criteria were recalculated using the Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for bromoform are 4.3 ng/L for ingestion of con-
taminated water and organisms and 360 fig/L for ingestion of
contaminated organisms only. IRIS values are based on a 10"6 risk level for
carcinogens.
Dichlorobromomethane 75-27-4
Human health criteria were recalculated using the IRIS to reflect available
data as of 12/92 (57 F.R. 60848). Recalculated IRIS values for dichlorobro-
momethane are 0.27 ng/L for ingestion of contaminated water and
organisms and 22 ng/L for ingestion of contaminated organisms only. IRIS
values are based on a 10"6 risk level for carcinogens.
Methylene Chloride 75-09-2
Human health criteria were recalculated using the IRIS to reflect available
data as of 12/92 (57 F.R. 60848). Recalculated IRIS values for methylene
chloride are 4.7 ng/L for ingestion of contaminated water and organisms
and 1,600 ng/L for ingestion of contaminated organisms only. IRIS values
are based on a 10"6 risk level for carcinogens.
Methyl Chloride (Chloromethane) 74-87-3
Human health criteria have been withdrawn for this compound
(57 F.R. 60848, December 22, 1992). However, EPA published a document,
CRITERIA
Aquatic Life
Human Health
133

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"Ambient Water Quality Criteria for Halomethanes," which includes this
compound. This document may contain useful information on human
health and was originally noticed on December 22,1992.
Methyl Bromide (Bromomethane) 74-83-9
Human health criteria were recalculated using the IRIS to reflect available
data as of 12/92 (57 F.R. 60848). Recalculated IRIS values for methyl bro-
mide are 48 ng/L for ingestion of contaminated water and organisms and
4000 ng/L for ingestion of contaminated organisms only.
(45 F.R. 79318, November 28, 1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.

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HARDNESS
CRITERIA
Introduction Water hardness is caused by polyvalent metallic ions dissolved in water. In
fresh water, these metallic ions are primarily calcium and magnesium, al-
though other metals such as iron, strontium, and manganese can also be
present in appreciable concentrations. Commonly, hardness is reported as
an equivalent concentration of calcium carbonate (CaC03).
The concept of hardness comes from water supply practices: it is mea-
sured by soap requirements for adequate lather formation and as an
indicator of the rate of scale formation in hot water heaters and low pres-
sure boilers. A commonly used classification is given in Table 1.
Table 1.—Classification of water by hardness content.
MAXIMUM CONCENTRATION
WATER DESCRIPTION
MG/L AS CACO,
0-75 	
soft
75-150 	
moderately hard
150-300 	
hard
300 and up	
very hard
The principal natural source of hardness is limestone, which is dis-
solved by percolating rainwater made acid by carbon dioxide. Industrial
and industrially related sources of hardness include the inorganic chemical
industry and discharges from operating and abandoned mines.
Hardness in fresh water frequently is distinguished in carbonate and
noncarbonate fractions: carbonate fractions are chemically equivalent to
the bicarbonates present in water. Since bicarbonates generally are mea-
sured as alkalinity, the carbonate hardness is usually considered equal to
the alkalinity.
Rationale	The determination of hardness in raw waters subsequently treated and
used for domestic water supplies is useful as a parameter to characterize
the total dissolved solids present and for calculating dosages where lime-
soda softening is practiced. Because hardness concentrations in water have
not been proven health related, the final level achieved is principally a
function of economics. Since hardness in water can be removed with treat-
ment by such processes as lime-soda softening and zeolite or ion exchange
systems, a criterion for raw waters used for public water supply is not
practical.
The effects of hardness on freshwater fish and other aquatic life appear
to be related to the ions causing the hardness rather than hardness. Both
the National Technical Advisory Committee (NTAC) and The National
Academy of Sciences (NAS) panels have recommended against using the
term hardness but suggest including the concentrations of the specific ions.
This procedure should avoid confusion in future studies but is not helpful
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in evaluating previous studies. For most existing data, it is difficult to de-
termine whether toxicity of various metal ions is reduced because of the
formation of metallic hydroxides and carbonates caused by the associated
increases in alkalinity, or because of an antagonistic effect of one of the
principal cations contributing to hardness — e.g., calcium — or a combina-
tion of both effects. One theory presented, without proof, that if cupric ions
were the toxic form of copper, whereas copper carbonate complexes were
relatively non-toxic, then the observed difference in toxicity of copper be-
tween hard and soft waters can be explained by the difference in alkalinity
rather than hardness. A review of the literature on toxicity presented data
showing that increasing calcium, in particular, reduced the toxicity of other
heavy metals. Under usual conditions in fresh water and assuming that
other bivalent metals behave similarly to copper, we can assume that both
effects occur simultaneously and explain the observed reduction of toxicity
of metals in waters containing carbonate hardness. The amount of reduced
toxicity related to hardness, as measured by a 40-hour LC50 for rainbow
trout, has been estimated to be about four times for copper and zinc when
the hardness was increased from 10 to 100 mg/L as CaC03.
Limits on hardness for industrial uses are quite variable. Table 2 lists
maximum values accepted by various industries as a source of raw water.
Subsequent treatment generally can reduce harness to tolerable limits, al-
though costs of such treatment are an important factor in determining its
desirability for a particular water source.
Hardness is not a determination of concern for irrigation water use.
The concentrations of the cations calcium and magnesium, which comprise
hardness, are important in determining the exchangeable sodium in a
given water. This particular calculation will be discussed under total dis-
solved solids rather than hardness.
Table 2.—Maximum hardness levels accepted by industry as a raw water source.*
INDUSTRY
MAXIMUM CONCENTRATION
MG/L AS CACO,
Electric utilities
5,000
Textile
120
Pulp and paper
475
Chemical
1,000
Petroleum
900
Primary metals
1,000
'Requirements for final use within e process may be essentially zero, whtch requires treatment for concentration
reductions
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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HEPTACHLOR
76-44-8
CRITERIA
Aquatic Life The criterion to protect freshwater aquatic life for heptachlor and hepta-
chlor epoxide, as derived using the guidelines, is 0.0038 ng/L as a 24-hour
average, and the concentration should not exceed 0.52 ng/L at any time.
The criterion to protect saltwater aquatic life for heptachlor and hepta-
chlor expoxide, as derived using the guidelines, is 0.0036 ng/L as a
24-hour average. The concentration should not exceed 0.053 ng/L at any
time.
Human Health For the maximum protection of human health from potential carcinogenic
effects from exposure to heptachlor through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tion should be zero, based on the nonthreshold assumption for this
chemical; however, zero level may not be attainable. Therefore, the levels
that may result in incremental increase of cancer risk over a lifetime are es-
timated at 10"5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 RR. 60848).
Recaculated IRIS values for heptachlor is 0.00021 ng/L for ingestion of
contaminated water and organisms and for ingestion of contaminated
aquatic organisms only. IRIS values are based on a 10*6 risk level for carcin-
ogens.
Published human health criteria were recalculated using IRIS to reflect
available data as of 12/92 (57 RR. 60848). Recalculated IRIS values for hep-
tachlor epoxide are 0.00010 ng/L for ingestion of contaminated water and
organisms and 0.00011 M-g/L for ingestion of contaminated aquatic organ-
isms only. IRIS values are based on a 10"6 risk level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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138

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HEXACHLOROBENZENE
118-74-1
CRITERIA
Aquatic Life As of 8/88, a draft Ambient Water Quality Criteria (AWQC) document for
hexachlorobenzene became available. Final rulemaking will eventually be
promulgated, but as of this writing, aquatic life criteria for
hexachlorobenzene has not been finalized.
Human Health Refer to Chlorinated Benzenes.
139

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140

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HEXACHLOROBUTADIENE
86-68-3
CRITERIA
Aquatic Life The available data for hexachlorobutadiene indicate that acute and chronic
toxicity to freshwater aquatic life occur at concentrations as low as 90 and
9.3 ng/L, respectively, and would occur at lower concentrations among
species that are more sensitive than those tested.
The available data for hexachlorobutadiene indicate that acute toxicity
to saltwater aquatic life occurs at concentrations as low as 32 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
hexachlorobutadiene to sensitive saltwater aquatic life.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to hexachlorobutadiene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical; however, zero level may not be attainable. Therefore,
the levels that may result in incremental increase of cancer risk over a life-
time are estimated at 10'3,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect avaialable data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for hexachlorobutadiene are 0.44 ng/L for inges-
tion of contaminated water and organisms and 50 jxg/L for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
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142

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HEXACHLOROCYCLOHEXANE
58-89-9
CRITERIA
Aquatic Life
Gamma-hexachlorocyclohexane (Lindane) 58-89-9
For gamma-hexachlorocyclohexane (lindane), the criterion to protect fresh-
water aquatic life as derived using the guidelines is 0.080 ng/L as a
24-hour average. The concentration should not exceed 2.0 ng/L at any
time.
For saltwater aquatic life the concentration of lindane should not ex-
ceed 0.16 ng/L at any time. No data are available for lindane's chronic
toxicity to sensitive saltwater aquatic life.
BHC 680-73-1
The available data for a mixture of isomers of benzene hexachloride (BHC)
indicate that acute toxicity to freshwater aquatic life occurs at concentra-
tions as low as 100 ng/L and would occur at lower concentrations among
species that are more sensitive than those tested. No data are available con-
cerning the chronic toxicity of a mixture of isomers of BHC to sensitive
freshwater aquatic life.
The available data for a mixture of isomers of BHC indicate that acute
toxicity to saltwater aquatic life occurs at concentrations as low as 0.34
H-g/L and would occur at lower concentrations among species that are
more sensitive than those tested. No data are available concerning the
chronic toxicity of a mixture of isomers of BHC to sensitive saltwater
aquatic life.
Human Health
Alpha-hexachlorocyclohexane 319-84-6
For the maximum protection of human health from the potential carcino-
genic effects of exposure to alpha-hexachlorocyclohexane through
ingestion of contaminated water and contaminated aquatic organisms, the
ambient water concentrations should be zero, based on the nonthreshold
assumption for this chemical; however, zero level may not be attainable.
Therefore, the levels that may result in incremental increase of cancer risk
over the lifetime are estimated at 10~5,10"6, and 10"7.
Human health criteria were recalculated using Integrated Risk Infor-
mation System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848).
Recalculated IRIS values for hexachlorocyclohexane-alpha are 0.0039 for
ingestion of contaminated water and organisms and 0.013 for ingestion of
contaminated aquatic organisms only. IRIS values are based on a 10"6 risk
level for carcinogens.
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Beta-hexachlorocyclohexane 319-85-7
For the maximum protection of human health from the potential carcino-
genic effects of exposure to beta-hexachlorocyclohexane through ingestion
of contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical; however, zero level may not be attainable. Therefore,
the levels that may result in incremental increase of cancer risk over the
lifetime are estimated at 10"5, 10"6, and 10'7. The corresponding recom-
mended criteria are 163 ng/L, 16.3 ng/L, and 1.63 ng/L, respectively. If
these estimates are made for consumption of aquatic organisms only, ex-
cluding consumption of water, the levels are 547 ng/L, 54.7 ng/L, and 5.47
ng/L, respectively.
Published human health criteria were recalculated using IRIS to reflect
available data as of 3/91. Recalculated IRIS values for hexachlorocyclohex-
ane-beta are 0.014 ng/L for ingestion of contaminated water and
organisms and 0.046 ng/L for ingestion of contaminated aquatic organ-
isms only. IRIS values are based on a 10"6 risk level for carcinogens.
Gamma-hexachlorocyclohexane (Lindane) 58-89-9
For the maximum protection of human health from the potential carcino-
genic effects due to exposure of gamma-hexachlorocyclohexane through
ingestion of contaminated water and contaminated aquatic organisms, the
ambient water concentrations should be zero, based on the nonthreshold
assumption for this chemical; however, zero level may not be attainable.
Therefore, the levels that may result in incremental increase of cancer risk
c.	£.	n
over a lifetime are estimated at 10" , 10 , and 10 . The corresponding rec-
ommended criteria are 186 ng/L, 18.6 ng/L, and 1.86 ng/L, respectively. If
these estimates are made for consumption of aquatic organisms only, ex-
cluding consumption of water, the levels are 625 ng/L, 62.5 ng/L, and 6.25
ng/L, respectively.
Published human health criteria were recalculated using IRIS to reflect
available data as of 3/91. Recalculated IRIS values for hexachlorocyclohex-
ane-gamma (Lindane) are 0.019 ng/L for ingestion of contaminated water
and organisms and 0.063 for ingestion of contaminated aquatic organisms
only. IRIS values are based on a 10"6 risk level for carcinogens.
Technical-hexachlorocyclohexane 319-86-8
Using the present guidelines, satisfactory criteria cannot now be derived
for delta and epsilon hexachlorocyclohexane because of insufficient data.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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HEXACHLOROCYCLOPENTADIENE
77-47-4
CRITERIA
Aquatic Life The available data for hexachlorocyclopentadiene indicate that acute and
chronic toxicity to freshwater aquatic life occurs at concentrations as low
as 7.0 and 5.2 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested.
The available data for hexachlorocyclopentadiene indicate that acute
toxicity to saltwater aquatic life occurs at concentrations as low as 7.0 ng/L
and would occur at lower concentrations among species that are more sen-
sitive than those tested. No data are available concerning the chronic
toxicity of hexachlorocyclopentadiene to sensitive saltwater aquatic life.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848)..Recalcul-
ated IRIS values for hexachlorocyclopentadiene are 240.0 ng/L for
ingestion of contaminated water and organisms and 17,000. ng/L for in-
gestion of contaminated aquatic organisms only.
Using available organoleptic data, the estimated level is 1 ng/L to con-
trol undesirable taste and odor quality of ambient water. Organoleptic data
have limitations as a basis for establishing water quality criteria but no
demonstrated relationship to potentially adverse effects on human health.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
145

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146

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IRON
7439-89-6
CRITERIA
Aquatic Life 0.3 mg/L for domestic water supply (health).
1.0 mg/L for freshwater aquatic life.
Introduction Of the elements that make up the earth's crust, iron is the fourth most
abundant by weight. Common in many rocks, iron is an important compo-
nent of many soils but especially clay, where it is usually a major
constituent. Iron may be present in varying quantities in water, depending
upon the area's geology and the waterway's other chemical components.
Iron is an essential trace element required by both plants and animals.
In some waters, it may be a limiting factor for the growth of algae and
other plants, especially in some marl lakes where it is precipitated by the
highly alkaline conditions. Also, iron is a vital oxygen transport mecha-
nism in the blood of all vertebrate and some invertebrate animals.
The ferrous, or bivalent (Fe++), and the ferric, or trivalent (Fe+++), irons
are of primary concern in the aquatic environment, although other forms
of iron may occur in organic and inorganic wastewater streams. The fer-
rous (Fe++) form can persist in waters void of dissolved oxygen and
originates usually from groundwaters or mines that have been pumped or
drained. For practical purposes, the ferric (Fe+++) form is insoluble. Iron
can exist in natural organometallic or humic compounds and colloidal
forms. Black or brown swamp waters can contain iron concentrations of
several mg/L in the presence or absence of dissolved oxygen, but this form
of iron has little effect on aquatic life.
Soluble ferrous iron occurs in the deep waters of stratified lakes with
anaerobic hypolimnia. During the autumnal or vernal overturns and with
aeration of these lakes, it is oxidized rapidly to the ferric ion that precipi-
tates to the bottom sediments as a hydroxide, Fe(OH)3, or with other
anions. If hydrogen sulfide (H2S) is present in anaerobic bottom waters or
muds, ferrous sulfide (FeS) may be formed. Ferrous sulfide is a black com-
pound and produces black mineral muds.
Prime iron pollution sources are industrial wastes, mine drainage wa-
ters, and iron-bearing groundwaters. In the presence of dissolved oxygen,
iron in water from mine drainage is precipitated as a hydroxide, Fe(OH)3.
These yellowish or ochre precipitates produce "yellow boy" deposits
found in many streams draining coal mining regions of Appalachia. Occa-
sionally, ferric oxide (Fe203) is precipitated, forming red waters. Both of
these precipitates form as gels or floes that may be detrimental to fishes
and other aquatic life when suspended in water. These precipitates can set-
tle to form flocculent materials that cover stream bottoms, thereby
destroying bottom-dwelling invertebrates, plants, or incubating fish eggs.
With time these floes can consolidate to form cement-like materials, thus
consolidating bottom gravels into pavement-like areas unsuitable as
spawning sites for nest-building fishes. This is particularly detrimental to
147

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trout and salmon populations whose eggs are protected in the interstices of
gravel and incubated with oxygen-bearing waters passing through the
gravel.
Rational	Iron is an objectional constituent in water supplies for both domestic and
industrial use. Iron appreciably affects the taste of beverages and can stain
laundered clothes and plumping fixtures. A study by the Public Health
Service (see original document) indicates that the taste of iron may be de-
tected readily at 1.8 mg/L in spring water and at 3.4 mg/L in distilled
water.
The daily nutritional requirement for iron is 1 to 2 mg, but intake of
larger quantities is required as a result of poor absorption. Diets contain 7
to 35 mg per day and average 16 mg. The iron criterion in water to prevent
objectionable tastes or laundry staining (0.3 mg/L) constitutes only a small
fraction of the iron normally consumed and is of aesthetic rather than toxi-
cological significance.
Studies obtain 96-hour LC50 values of 0.32 mg/L iron for mayflies,
stoneflies, and caddisflies; all are important fish food organisms. Other
studies found iron toxic to carp (Cyrinus carpio) at concentrations of
0.9 mg/L when the pH of the water was 5.5. Pike (Esox lucius) and trout
(species unknown) died at iron concentrations of 1 to 2 mg/L. In an iron
polluted Colorado stream, neither trout nor other fish were found until the
waters were diluted or the iron had precipitated to effect a concentration of
less than 1.0 mg/L, even though other water quality constituents were
suitable for the presence of trout.
Ferric hydroxide floes have been observed to coat the gills of white
perch (Morone americanus), minnows, and silversides (Menidia sp). The
smothering effects of settled iron precipitates may be particularly detri-
mental to fish eggs and bottom-dwelling fish food organisms. Iron
deposits in the Brule River, Michigan and Wisconsin, were found to have a
residual long-term effect on fish food organisms even after the pumping of
iron-bearing waters from deep shaft iron mines had ceased. Settling iron
floes have also been reported to trap and carry diatoms downward in wa-
ters.
In 69 of 75 study sites with good fish fauna, the iron concentrations
were less than 10.0 mg/L. The European Inland Fisheries Commission rec-
ommended that iron concentrations not exceed 1.0 mg/L in waters to be
managed for aquatic life.
Based principally on field observations, a criterion of 1 mg/L iron for
freshwater aquatic life is believed to be adequately protective. As noted,
data obtained under laboratory conditions suggest a greater toxicity for
iron than that obtained in natural ecosystems. Ambient natural waters will
vary according to alkalinity, pH, hardness, temperature, and the presence
of ligands, which change the valence state and solubility and therefore the
toxicity of the metal.
The effects of iron on marine life have not been investigated ade-
quately to determine a water quality criterion. Dissolved iron readily
precipitates in alkaline seawaters. Fears have been expressed that these set-
tled iron floes may have adverse effects on important benthic commercial
mussels and other shellfish resources.
Iron has not been reported to have a direct effect on the recreational
uses of water, other than its effect on aquatic life. Suspended iron precipi-
tates may interfere with swimming and be aesthetically objectionable with
water deposits as yellow ochre or reddish iron oxides.
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Iron at exceedingly high concentrations has been reported to be toxic to
livestock and interfere with the metabolism of phosphorus. Dietary sup-
plements of phosphorus can be used to overcome this metabolic deficiency.
In aerated soils, iron in irrigation waters is not toxic. Precipitated iron may
be complex phosphorus and molybdenum, making them less available as
plant nutrients. In alkaline soils, iron may be so insoluble as to be deficient
as a trace element and result in chlorosis, an objectionable plant nutrient
deficiency disease. A reported reduction in the quality of tobacco was due
to precipitate iron oxides on the leaves when the crop was spray irrigated
with water containing 5 mg/L of soluble iron.
For some industries, iron concentrations in process waters lower than
that required for public water supplies are required or desirable. Examples
include high pressure boiler feed waters; scouring, bleaching, and dyeing
of textiles; certain types of paper production; some chemicals; some food
processing; and leather finishing industries.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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ISOPHORONE
78-59-1
CRITERIA
Aquatic Life The available data for isophorone indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 117,000 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of
isophorone to sensitive freshwater aquatic life.
The available data for isophorone indicate that acute toxicity to salt-
water aquatic life occurs at concentrations as low as 12,900 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
isophorone to sensitive saltwater aquatic life.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for isophorone are 8.4 ng/L for ingestion of contaminated
water and organisms and 600 for ingestion of contaminated aquatic organ-
isms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
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LEAD
7439-92-1
Saltwater — 1-hour average of 220 |ig/L
4-day average of 8.5 |xg/L
Freshwater criteria are hardness dependent. See text.
The acute toxicity of lead to several species of freshwater animals has been
shown to decrease as the hardness of water increases. At a hardness of 50
mg/L, the acute sensitivities of 10 species range from 142.5 M-g/L for an
amphipod to 235,900 ng/L for a midge. Data on the chronic effects of lead
on freshwater animals are available for two fish and two invertebrate spe-
cies. The chronic toxicity of lead also decreases as hardness increases, and
the lowest and highest available chronic values (12.26 and 128.1 (ig/L) are
both for a cladoceran, but in soft and hard water, respectively. Acute-
chronic ratios are available for three species and range from 18 to 62.
Freshwater algae are affected by concentrations of lead above 500 ng/L,
based on data for four species. Bioconcentration factors are available for
four invertebrate and two fish species and range from 42 to 1,700.
Acute values are available for 13 saltwater animal species and range
from 315 ng/L for the mummichog to 27,000 ng/L for the soft shell clam. A
chronic toxicity test was conducted with a mysid; unacceptable effects
were observed at 37 ng/L but not at 17 ng/L; the acute-chronic ratio for
this species is 124.8. A species of macroalgae was affected at 20 ng/L.
Available bioconcentration factors range from 17.5 to 2,570.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration (in
M-g/L) of lead does not exceed the numerical value given by
_{1.273[1n (hardness)]-4.705)
more than once every three years on the average, and if the one-hour
average concentration (in ng/L) does not exceed the numerical value given
by
_(1.273[ln(hardness)]- 1.460)
v
more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaCC^ the four-day average
concentrations of lead are 1.3, 3.2, and 7.7 (xg/L, respectively, and the
one-hour average concentrations are 34, 82, and 200 ng/L.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of lead does not exceed 8.5* ng/L more than
CRITERIA
Aquatic Life
Summary
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once every three years on the average and if the one-hour average concen-
tration does not exceed 220* ng/L more than once every three years on the
average.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average time needed for an unstressed sys-
tem to recover from a pollution event in which lead exposure exceeds the
criterion. A stressed system — for example one, in which several outfalls
occur in a limited area — would be expected to require more recovery
time. The resilience of ecosystems and their ability to recover differ greatly,
however, and site-specific criteria may be established if adequate justifica-
tion is provided.
The use of criteria in designing waste treatment facilities requires the
selecting of an appropriate wasteload allocation model. Dynamic models
are preferred for applying these criteria. Limited data or other factors may
make their use impractical, in which case one should rely on a steady-state
model. The Agency recommends the interim use of 1Q5 or 1Q10 for Crite-
rion Maximum Concentration design flow, and 7Q5 or 7Q10 for the
Criterion Continuous Concentration design flow in steady-state models
for unstressed and stressed systems, respectively. These matters are dis-
cussed in more detail in EPA's "Technical Support Document for Water
Quality-Based Toxics Control."
Human Health Human health criteria have been withdrawn for this compound (see
57 F.R. 60885, December 22, 1992). Although the human health criteria are
withdrawn, EPA published a document for this compound that may con-
tain useful human health information. This document was originally
noticed in 45 F.R. 79331, November 28,1980.
(45 F.R. 79318 November 28,1980) (50 F.R. 30784, July 29, 1985)
See Appendix A for Aquatic Life Methodology.
* Saltwater lead concentrations are based on a recalculation. See 57 F.R. 60882,
December 22,1992, Comment #45.
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MALATHION
121-75-5
CRITERIA
Aquatic Life 0.1 ng/L for freshwater and saltwater aquatic life.
Rationale	Salmonids and centrarachids appear to be the most sensitive freshwater
fish to malathion. Documented 96-hour LC50s ranged from 120 to 265
Hg/L for four different salmonid species and 101 to 285 ng/L for three spe-
cies of centrarchids. Reported 96-hour LC50s for rainbow trout
(Onchorhynchus mykiss), largemouth bass (Micropterus salmoides), and
chinook salmon, (Onchorhynchus tshawytcha) were 86, 50, and 28 ng/L, re-
spectively.
Freshwater invertebrates are generally even more sensitive to mala-
thion than fish. Reported 96-hour LC50s for amphipod Gammarus lacustris
and G. fasciutus were 1.0 and 0.76 ng/L, respectively. The 48-hour LC50s
for the cladocerans Simocephalus serrulatus and Daphnia pulex ranged from
1.8 to 3.5 n-g/L, while the 24-hour LC50s for two midge larvae species aver-
aged just over two ng/L in tests with malathion.
In flow-through exposures to malathion with saltwater fish (Lagodon
rhombides), 575 ]u.g/L result in a 50 percent mortality rate in 3.5 hours and
caused about 75 percent brain acetylcholinesterase (AChE) inhibition. Sim-
ilar mortality and AChE inhibition is documented with other saltwater fish
as well. Static 96-hour tests for saltwater teleosts exposed to malathion in-
dicates a broad spectrum of species sensitivities, with LC50 values ranging
from 27 to 3,250 jig/L for several different species. For the commercially
and economically important striped bass, Morone saxatilis, a flow-through
96-hour LC50 of 14 ng/L is documented.
Saltwater invertebrates are comparably sensitive to the less-resistant
fish species with 96-hour LC50s ranging from 33 ng/L for sand shrimp
(Crangon septemspinosa) to 83 |i.g/L for the hermit crab (Pagurus
longicorpus).
Malathion enters the aquatic environment primarily as a result of its
application as an insecticide. Because it degrades quite rapidly in most wa-
ters, depending on pH, its occurrence is sporadic rather than continous.
Because malathion's toxicity is exerted through inhibition of the enzyme
AChE and because such inhibition may be additive with repeated expo-
sures and may be caused by any of the organophosphorus insecticides,
inhibition of AChE by more than 35 percent may be expected to result in
damage to aquatic organisms.
An application factor of 0.1 is applied to the 96-hour LC50 data for
Gammarus lacustris, G. fasciatus, and Daphnia, which are all approximately
1.0 ng/L, yielding a criterion of 0.1ng/L.
(Quality Criteria for Water, July 1976) PB-263943
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MANGANESE
7439-96-5
CRITERIA 	
50 ng/L for domestic water supply (health).
100 (xg/L for protection of consumers of saltwater molluscs.
Introduction Manganese does not occur naturally as a metal but is found in various salts
and minerals that are frequently in association with iron compounds. The
principal manganese-containing substances are manganese dioxide
(Mn02), pyrolusite, manganese carbonate (rhodocrosite), and manganese
silicate (rhodonite).
The oxides are the only important minerals mined. Manganese is not
mined in the United States, except when contained in iron ores that are de-
liberately used to form ferro-manganese alloys.
The primary uses of manganese are in metal alloys, dry cell batteries,
micro-nutrient fertilizer additives, organic compounds used in paint dri-
ers, and as chemical reagents. Permanganates are very strong oxidizing
agents of organic materials.
Manganese is a vital micro-nutrient for both plants and animals. When
manganese is not present in sufficient quantities, plants exhibit chlorosis (a
yellowing of the leaves) or failure of proper leaf development. Inadequate
quantities of manganese in domestic animal food results in reduced repro-
ductive capabilities and deformed or poorly maturing young. Livestock
feeds usually have sufficient manganese, but beef cattle on a high corn diet
may require a supplement.
Rationale	Although inhaled manganese dusts have been reported to be toxic to hu-
mans, manganese normally is ingested as a trace nutrient in food. The
average human intake is approximately 10 mg per day. Very large doses of
ingested manganese can cause some disease and liver damage, but these
are not known to occur in the United States. Only a few manganese toxicity
problems have been found throughout the world, and these have occurred
under unique circimstances (i.e., a well in Japan near a deposit of buried
batteries).
It is possible to partially sequester manganese with special treatment,
but manganese is not removed in the conventional treatment of domestic
waters. Consumer complaints arise when manganese exceeds a concentra-
tion of 150 ng/L in water supplies. These complaints are concerned
primarily with the brownish staining of laundry and objectionable tastes in
beverages. The presence of low concentrations of iron may intensify the
adverse effects of manganese. Manganese at concentrations of about 10 to
20 ng/L is acceptable to most consumers. A criterion for domestic water
supplies of 50 ng/L should minimize the objectionable qualities.
Ions of manganese are found rarely at concentrations above 1 mg/L in
freshwater. The reported tolerance values range from 1.5 mg/L to over
1,000 mg/L. Thus, manganese is not considered to be a problem in fresh
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waters. Permanganates have been reported to kill fish in 8 to 18 hours at
concentrations of 2.2 to 4.1 mg/L. Permanganates are not persistent be-
cause they rapidly oxidize organic materials and are thereby reduced and
rendered nontoxic.
Few data are available on the toxicity of manganese to marine organ-
isms. The ambient concentration of manganese is about 2 ng/L. The
material is rapidly assimulated and bioconcentrated into nodules that are
deposited on the sea floor. The major problem with manganese may be
concentration in the edible portions of mollusks, as bioaccumulation fac-
tors as high as 12,000 have been reported. In order to protect against a
possible health hazard to humans by manganese accumulation in shellfish,
a criterion of 100 ng/L is recommended for marine water.
Manganese is not known to be a problem in water consumed by live-
stock. At concentrations of slighly less than 1 mg/L to a few milligrams per
liter, manganese may be toxic to plants from irrigation water applied to
soils with pH values lower than 6.0. The problem may be rectified by lim-
ing soils to increase the pH. Problems may develop with long-term
(20-year) continuous irrigation on other soils with water containing about
10 mg/L of manganese. But as stated previously, manganese rarely is
found in surface waters at concentrations greater than 1 mg/L. Thus, no
specific criterion for manganese in agricultural waters is proposed. In se-
lect areas and where acidophilic crops are cultivated and irrigated, a
criterion of 200 |xg/L is suggested for consideration.
Most indrustrial users of water can operate successfully where the cri-
terion proposed for public water supplies is observed. However, a more
restrictive criterion may be needed to protect or ensure product quality.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.

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*MERCURY
7439-97-6
CRITERIA
Not to exceed 2.4 ng/L in fresh water or 2.1 jxg/L in salt water.
0.012 and 0.025 fig/L for freshwater and saltwater aquatic life, respectively.
AQUATIC LIFE
SUMMARY Data are available on the acute toxicity of mercury (II) to 28 genera of
freshwater animals. Acute values for invertebrate species range from 2.2
Hg/L for Daphnia pulex to 2,000 ng/L for three insects. Acute values for
fishes range from 30 ng/L for the guppy to 1,000 ng/L for the Mozambique
tilapia. Few data are available for various organomercury compounds and
mercurous nitrate, and they all appear to be 4 to 31 times more acutely
toxic than mercury (II).
Available chronic data indicate that methylmercury is the most chroni-
cally toxic of the tested mercury compounds. Tests on methylmercury with
Daphnia magna and brook trout (Salvelinus fontinalis) produced chronic val-
ues less than 0.07 ng/L. For mercury (II) the chronic value obtained with
D. magna was about 1.1 fig/L and the acute-chronic ratio was 4.5. In both a
life-cycle test and an early life-stage test on mercuric chloride with the fat-
head minnow (Pinnephales promelas), the chronic value was less than 0.26
M-g/L, and the acute-chronic ratio was over 600.
Freshwater plants show a wide range of sensitivities to mercury, but
the most sensitive appear to be less affected than the most sensitive fresh-
water animals to both mercury (II) and methylmercury. A bioconcentration
factor of 4,994 is available for mercury (II), but the bioconcentration factors
for methylmercury range from 4,000 to 85,000.
Data on the acute toxicity of mercuric chloride are available for 29 gen-
era of saltwater animals, including annelids, molluscs, crustaceans,
echinoderms, and fishes. Acute values range from 3.5 ng/L for a mysid to
1,678 ng/L for winter flounder (Pseudo pleuroneotes americanus). Fishes tend
to be more resistant, and molluscs and crustaceans tend to be more sensi-
tive to the acute toxic effects of mercury (II). Results of a life-cycle test with
the mysid show that mercury (II) at a concentration of 1.6 ng/L signifi-
cantly affected time of first spawn and productivity; the resulting
acute-chronic ratio was 3.1.
Concentrations of mercury that affected growth and photosynthetic ac-
tivity of one saltwater diatom and six species of brown algae range from 10
to 160 ng/L. Bioconcentration factors of 10,000 and 40,000 have been ob-
tained for mercuric chloride and methylmercury with an oyster.
National Criteria Derivation of a water quality criterion for mercury is more complex than
for most metals because of methylation of mercury in sediments, fish, and
their food chain. Apparently, almost all mercury currently being
•Indicates suspension, cancelation, or restriction by U.S. EPA Office of Pesticides and
Toxic Substances.
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discharged is mercury (II). Thus mercury (II) should be the only important
possible cause of acute toxicity, and the Criterion Maximum Concentra-
tions can be based on its acute values.
The best available data on long-term exposure of fish to mercury (II)
indicates that concentrations above 0.23 ng/L statistically affected fathead
minnows significantly, causing the concentration of total mercury in the
whole body to exceed 1.0 mg/kg. Although it is not known what percent
of the mercury in the fish was methylmercury, it is also not known whether
uptake from food would increase the concentration in natural situations.
Species such as rainbow trout (Oncorhynchus mykiss), coho salmon
(Oncorhynchus kisutch), and especially the bluegill (Lepomis macrochirus)
might suffer chronic effects and accumulate high residues of mercury as
did the fathead minnow.
With regard to long-term exposure to methylmercury, scientists found
that brook trout can exceed the FDA action level without suffering statisti-
cally significant adverse effects on survival, growth, or reproduction. Thus
for methylmercury, the Final Residue Value would be substantially lower
than the Final Chronic Value.
Basing a freshwater criterion on the Final Residue Value of 0.012 ng/L
derived from the bioconcentration factor of 81,700 for methylmercury with
the fathead minnow essentially assumes that all discharged mercury is
methylmercury. On the other hand, in field situations uptake from food
might add to the uptake from water. Similar considerations apply to the
derivation of the saltwater criterion of 0.025 ng/L using the BCF of 40,000
obtained for methylmercury with the eastern oyster. Because the Final Res-
idue Values for methylmercury are substantially below the Final Chronic
Values for mercury (II), of lesser importance is that many fishes, including
the rainbow trout, coho salmon, bluegill, and haddock (Melanogrammus
aeglefinus), might not be adequately protected by the freshwater and salt-
water Final Chronic Values for mercury (II).
In contrast to the complexities of deriving numerical criteria for mer-
cury, monitoring for unacceptable environmental effects should be
relatively straightforward. The most sensitive adverse effect will probably
be to exceed the FDA action level. Therefore, existing discharges should be
acceptable if the concentration of methylmercury in the edible portion of
exposed consumed species does not exceed the FDA action level.
The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms
and Their Uses" indicate that, except possibly where a locally important
species is very sensitive, freshwater aquatic organisms and their uses
should not be affected unacceptably if the four-day average concentration
of mercury does not exceed 0.012 ng/L more than once every three years
on the average and if the one-hour average concentration does not exceed
2.4 |xg/L more than once every three years on the average. If the four-day
average concentration exceeds 0.012 ng/L more than once in a three-year
period, the edible portion of consumed species should be analyzed to de-
termine whether the concentration of methylmercury exceeds the FDA
action level.
The procedures described in the guidelines indicate that, except possi-
bly where a localy important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of mercury does not exceed 0.025 [ig/L more
than once every three years on the average and if the one-hour average
concentration does not exceed 2.1 ng/L more than once every three years
on the average. If the four-day average concentration exceeds 0.025 ng/L

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more than once in a three-year period, the edible protion of consumed spe-
cies should be analyzed to determine whether the concentration of
mathylmercury exceeds the FDA action level.
The recommended exceedence frequency of three years is the Agency's
best scientific judgment of the average time needed for an unstressed sys-
tem to recover from a pollution event in which exposure to mercury
exceeds the criterion. A stressed system, for example — one in which sev-
eral outfalls occur in a limited area — would be expected to require more
time for recovery. The resilience of ecosystems and their ability to recover
differ greatly, however, and site-specific criteria can be established if ade-
quate justification is provided.
The use of criteria in designing waste treatment facilities requires the
selection of an appropriate wasteload allocation model. Dynamic models
are preferred for the application of these criteria. Limited data or other fac-
tors may make their use impractical, in which case one should rely on a
steady-state model. The Agency recommends the interim use of 1Q5 or
1Q10 for Criterion Maximum Concentration design flow and 7Q5 or 7Q10
for the Criterion Continuous Concentration design flow in steady-state
models for unstressed and stressed systems, respectively. These matters
are discussed in more detail in EPA's "Technical Support Document for
Water Quality-Based Toxics Control."
Human Health
Criteria	The ambient water criterion is 144 ng/L for the protection of human health
from the toxic properties of mercury ingested through water and contami-
nated aquatic organisms.
For the protection of human health from the toxic properties of mer-
cury ingested through contaminated aquatic organisms alone, the ambient
water criterion is 146 ng/L. These values include the consumption of fresh-
water, estuarine, and saltwater species.
(45 F.R. 79318 November 28,1980) (50 F.R. 30784, July 29,1985)
See Appendix A for Aquatic Life Methodology.
See Appendix C for Human Health Mehtodology.
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162

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METHOXYCHLOR
72-435
CRITERIA 	
100 ng/L for domestic water supply (health).
0.03 ng/L for freshwater and saltwater aquatic life.
Rationale	Where adequate human data are available for corraboration of the animal
results, the total "safe" drinking intake level is assumed to be 1/100 of the
"no effect" or "minimal effect" level reported for the most sensitive animal
tested — in this case, humans.
Applying the available data and assuming that 20 percent of the toal
intake of methoxychlor is from drinking water and that the average person
weighs 70 kg and consumes two liters of water per day, the formual for cal-
culating a criterion is 2.0 mg/kg x 0.2 x 70 kg x 1/100 x 1/2 = 0.14 ng/L. A
criterion level for domestic water supply of 100 ng/L is recommended.
In tests with aquatic organisms exposed to methoxychlor, reduced
hatchability of fathead minnow (Pimephales promelas) embryos at 0.125
Hg/L and lack of spawning at 2.0 fig/L was observed. Yellow perch (Perca
flavescens) exposed to 0.6 ng/L methoxychlor for 8 months exhibited re-
duced growth. The four-day LC50s for the fathead minnow, the yellow
perch, and economically important striped bass (Morone saxatilis) were 7.5,
22, and 3.3 ng/L, respectively.
The four-day LC50s for aquatic invertebrates were as low as 0.61 (ig/L
for the scud (Gammarus pseudolimnaeus) and ranged from 1.6 to 7 ng/L for
three insect larvae and a crayfish. In 28-day tests, reduction in emergence
or pupation of aquatic insects was observed at 0.25 to 0.5 ng/L of methoxy-
chlor.
The data indicate that 0.1 ng/L methoxychlor would be just below the
chronic effect level for the fathead minnow and Vb the acute toxicity level
in crayfish. Thus, a criterion level of 0.03 ng/L is recommended. The 96-
hour LC50 for the marine pink shrimp (Penaeus duorarum) exposed to
methoxychlor is 3.5 ng/L. The 30-day LC50 was 1.3 ng/L. Applying an ap-
plication factor of 0.01 with the pink shrimp acute toxicity of 3.5 ng/L, the
recommended criterion for a saltwater environment is also 0.03 ng/L.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
163

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164

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MIREX
2385-85-5
CRITERIA 	
0.001 jig/L for freshwater and saltwater aquatic life.
Rationale	Mirex is generally limited in its use to control the imported fire ant in the
southeastern United States, and it is always presented in bait.
In experiments with field-collect crayfish, juvenile Procumbarus
blandingi were exposed to 1 or 5 ng/L mirex for six to 144 hours and then
transferred to clean water and observed for 10 days. After five days, 95
percent of the crayfish exposed to 1 (xg/L mirex for 144 hours were dead.
Exposure to 5 ng/L for 6, 24, and 58 hours resulted in 26, 50, and 98 percent
mortality within the 10-day observation period in clean water. Several sim-
ilar experiments with other crayfish species revealed comparable mortality
levels in exposures to low levels of mirex. For Promcambarus hayi, mirex tis-
sue residue accumulations ranged from 940- to 27,210-fold above water
concentrations after 48-hour exposures to 0.1 and 0.5^g/L.
In feeding experiments with 108 crayfish, one particle of mirex bait re-
sulted in a 77 percent mortality rate after six days. Comparable
experiments yielded similar results, indicating that mirex is extremely
toxic to the tested species of crayfish. Mortality and accumulation in-
creased with exposure time. Field studies have shown that mirex is
accumulated through the food chain, while additional data reveals that
mirex is transported from treated land into marsh and estuarine areas.
Mirex residues were found to increase with trophic levels of animals sam-
pled. In addition, further studies documented mirex in areas not treated
with the insecticide. Mirex has been reported in fish from Lake Ontario,
Canada, although mirex is not registered for use in Canada.
A summary of data available shows a mosaic of effects. Crayfish and
channel catfish survival is affected by mirex in the water or by ingestion of
the bait particles. Bioaccumulation is well established for a wide variety of
organisms. Mirex is very persistent in bird tissue. Considering the extreme
toxicity and potential for bioaccumulation, every effort should be made to
keep mirex bait particles out of water containing aquatic organisms, and
water concentrations should not exeed 0.001 ng/L mirex. This value is
based upon an application factor of 0.01 applied to the lowest levels at
which effects on crayfish have been observed.
Data on which to base a marine criterion involve several estuarine and
marine crustaceans. A concentration of 0.1 ng/L mirex was lethal to juve-
nile pink shrimp (Penaeus duorarum) in a three-week exposure. Reduced
survival of the mud crab (Rhithropanopeus harrisii) was observed in 0.1
M-g/L mirex. In three of four 28-day seasonal flow-through experiments, re-
duced survival of Callinectes sapidus, Penaeus duorarum, and grass shrimp
(Palaemonetes pugio), at levels of 0.12 M-g/L in summer, 0.06 ng/L in fall, and
0.09 M-g/L in winter, was observed.
Effects of mirex on estuarine and marine crustaceans were observed
after considerable time had elapsed, so that length of exposure is an
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important consideration for this chemical. This may not be the case in fresh
water since the crayfish were affected within 48 hours. Therefore, a three-
to four-week exposure might be considered acute; and by applying an ap-
plication factor of 0.01 to reasonable average of toxic effect levels as
summarized above, a recommended marine criterion of 0.001 ng/L results.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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NAPHTHALENE
91-20-3
CRITERIA
Aquatic Life The available data for naphthalene indicate that acute and chronic toxicity
to freshwater aquatic life occurs at concentrations as low as 2,300 and 620
fig/L, respectively, and would occur at lower concentrations among spe-
cies that are more sensitive than those tested.
The available data also indicate that acute toxicity to saltwater aquatic
life occurs at concentrations as low as 2,350 ng/L and would occur at lower
concentrations among species that are more sensitive than those tested. No
data are available concerning the chronic toxicity of naphthalene to sensi-
tive saltwater aquatic life.
Human Health Using the present guidelines, a satisfactory criterion cannot be derived at
this time because of insufficient available data.
(45 F.R. 79318, November 28,1980)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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168

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NICKEL
7440-02-0
CRITERIA 	
Not to exceed 75 ng/L in salt water.
8.3 (xg/L for saltwater aquatic life.
Freshwater criteria are hardness dependent. See text.
Summary	Acute values with 21 freshwater species in 18 genera range from 1,101
Hg/L for a cladoceran to 43,240 ng/L for a fish. Fishes and invertebrates
are both spread throughout the range of sensitivity. Acute values with four
species are significantly correlated with hardness. Data are available con-
cerning the chronic toxicity of nickel to two invertebrates and two fishes in
fresh water. Data available for two species indicate that chronic toxicity de-
creases as hardness increases. The measured chronic values ranged from
14.77 M-g/L for Daphnia magna in soft water to 526.7 ng/L for the fathead
minnow (Pimephales promelas) in hard water. Five acute-chronic ratios are
available for two species in soft and hard water and range from 14 to 122
Hg/L-
Nickel appears to be quite toxic to freshwater algae, with concentra-
tions as low as 50 ng/L producing significant inhibition. Bioconcentration
factors for nickel range from 0.8 for fish muscle to 193 for a cladoceran.
Acute values for 23 saltwater species in 20 genera range from 151.7
Hg/L for mysid juveniles to 1,100,000 |xg/L for juveniles and adults of a
clam. The acute values for the four species of fish range from 7,598 to
350,000 ng/L. Nickel's acute toxicity appears to be related to salinity and is
species-dependent.
An acceptable chronic test on nickel has been conducted on only one
saltwater species, Mysidopsis bahia. In it, chronic exposure to 141 pig/L and
greater resulted in reduced survival and reproduction; the measured
acute-chronic ratio was 5.478.
Bioconcentration factors in saltwater range from 261.8 for an oyster to
675 for a brown alga.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of
nickel (in (ig/L) does not exceed the numerical value given by
e(0.8460[ln (hardness)]+l.1645)
more than once every three years on the average and if the one-hour
average concentration (in ng/L) does not exceed the numerical value given
by
e(0.8460[ln (hardness)]+3.3612)
more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaC03, the four-day average
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concentrations of nickel are 88, 160, and 280 ng/L, respectively, and the
one-hour average concentrations are 790,1400, and 2500 ng/L.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of nickel does not exceed 8.3 ng/L more than
once every three years on the average and if the one-hour average concen-
tration does not exceed 75 ng/L more than once every three years on the
average.
Three years is the Agency's best scientific judgment of the average
amount of time aquatic ecosystems should be provided between excur-
sions. The resiliences of ecosystems and their abilities to recover differ
greatly, however, and site-specific, allowed excursion frequencies can be
established if adequate justification is provided.
When developing water quality-based permit limits and designing
waste treatment facilities, criteria must be applied to an appropriate
wasteload allocation model. Dynamic models are preferred, but limited
data or other considerations might make their use impractical; therefore,
regulatory programs must rely on a steady-state model.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60911). Recalcul-
ated IRIS values for nickel are 610 ^g/L for ingestion of contaminated
water and organisms and 4,600 ng/L for ingestion of contaminated aquatic
organisms only.
(45 F.R. 79337, November 28,1980) (51 F.R. 43665, December 3,1986)
(57 F.R. 60911, December 22,1992)
See Appendix A for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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NITRATES/NITRITES
14797-55-8
CRITERIA 	
10 mg/L nitrate nitrogen (N) for domestic water supply (health).
Introduction Two gases — molecular nitrogen and nitrous oxide — and five forms of
nongaseous, combined nitrogen — amino and amide groups, ammonium,
nitrite, and nitrate — are important in the nitrogen cycle. The amino and
amide groups are found in soil organic matter and as constituents of plant
and animal protein. The ammonium ion either is released from protein-
aceous organic matter and urea or is synthesized in industrial processes
involving atmospheric nitrogen fixation. The nitrite ion is formed from the
nitrate or the ammonium ions by certain microorganisms found in soil,
water, sewage, and the digestive tract. The nitrate ion is formed by the
complete oxidation of ammonium ions by soil or water microorganisms.
This process, known as denitrification, takes place when nitrate-containing
soils become anaerobic and the conversion to nitrite, molecular nitrogen,
or nitrous oxide occurs; in some instances, ammonium ions are produced.
In oxygenated natural water systems, nitrite is rapidly oxidized to nitrate.
Growing plants assimilate nitrate or ammonium ions and convert them to
protein.
Among the major point sources of nitrogen entry into waterbodies are
municipal and industrial wastewaters, septic tanks, and feed lot dis-
charges. Diffuse sources of nitrogen include farm-site fertilizer and animal
wastes, lawn fertilizer, leachate from waste disposal in dumps or sanitary
landfills, atmospheric fallout, nitric oxide and nitrite discharges from auto-
mobile exhausts and other combustion processes, and losses from natural
sources such as mineralization of soil organic matter. Water reuse systems
in some fish hatcheries employ a nitrification process for ammonia reduc-
tion; this can result in exposure of hatchery fish to elevated levels of nitrite.
Rationale	In quantities normally found in food or feed, nitrates become toxic only
under conditions in which they are, or may be, reduced to nitrites. Other-
wise, at "reasonable" concentrations, nitrates are rapidly excreted in the
urine. High intake of nitrates constitutes a hazard primarily to
warmblooded animals under conditions favorable to their reduction to ni-
trite. Under certain circumstances, nitrate can be reduced to nitrite in the
gastrointestinal tract, which then reaches the bloodstream and reacts di-
rectly with hemoglobin to produce methemoglobin, with consequent
impairment of oxygen transport.
The reaction of nitrite with hemoglobin can be hazardous in infants
under 3 months of age. Serious and occasionally fatal poisonings in infants
have occurred following ingestion of untreated well waters shown to con-
tain nitrate at concentrations greater than 10 mg/L nitrate /nitrogen (N).
High nitrate concentrations frequently are found in shallow farm and rural
community wells, often as the result of inadequate protection from barn-
yard drainage or from septic tanks. Increased concentrations of nitrates
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also have been found in streams from farm tile drainage in areas of intense
fertilization and farm crop production.
The differences in susceptibility to methemoglobina are not yet under-
stood. They appear, however, to be related to a combination of factors
including nitrate concentration, enteric bacteria, and the lower acidity
characteristic of the digestive system of baby mammals. Methemoglobine-
mia symptoms and other toxic effects were observed when high nitrate
well waters containing pathogenic bacteria were fed to laboratory mam-
mals. Conventional water treatment has no significant effect on nitrate
removal from water.
Because of the potential risk of methemoglobinemia to bottle-fed in-
fants, and in view of the absence of substantiated physiological effects at
nitrate concentrations below 10 mg/L nitrate/nitrogen, this level is the cri-
terion for domestic water supplies. Waters with nitrite/nitrogen
concentrations over 1 mg/L should not be used for infant feeding. Waters
with a significant nitrite concentration usually would be heavily polluted
and probably bacteriologically unacceptable.
Quality Criteria for Water, July 1976, provides data for exposed fishes.
This data concludes that
1.	Levels of nitrate/nitrogen at or below 90 mg/L would have no
adverse effects on warmwater fish.
2.	Nitrite/nitrogen at or below 5 mg/ L should be protective of most
warmwater fish.
3.	Nitrite/nitrogen at or below 0.06 mg/L should be protective of
salmonid fishes.
These levels are not known to occur or would be unlikely to occur in
natural surface waters. Recognizing that concentrations of nitrate or nitrite
that would exhibit toxic effects on warmwater fish could rarely occur in
nature, restrictive criteria are not recommended.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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NITROBENZENE
98-95-3
CRITERIA
Aquatic Life The available data for nitrobenzene indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 27,000 (ig/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No definitive data are available concerning the chronic
toxicity of nitrobenzene to sensitive freshwater aquatic life.
The available data for nitrobenzene indicate that acute toxicity to salt-
water aquatic life occurs at concentrations as low as 6,680 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No definitive data are available concerning the chronic toxic-
ity of nitrobenzene to sensitive saltwater aquatic life.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for nitrobenzene are 17.0 ng/L for ingestion of
contaminated water and organisms and 1,900 ng/L for ingestion of con-
taminated aquatic organisms only.
Using available organoleptic data, the estimated level is 30 fig/L to
control undesirable taste and odor qualities of ambient water. Organolep-
tic data do have limitations as a basis for establishing a water quality
criterion, however, but no demonstrated relationship to potentially ad-
verse effects on human health.
The U.S. EPA is currently developing Acceptable Daily Intake (ADI) or
Verified Reference Dose (RfD) values for agencywide use for this chemical.
The new value should be substituted when it becomes available. The Janu-
ary 1986 draft Verified Reference Dose document cites an RfD of .0005
mg/kg/day for nitrobenzene.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
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174

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NITROPHENOLS
CRITERIA
Aquatic Life The available data for nitrophenols indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 230 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of nitro-
phenols to sensitive freshwater aquatic life but toxicity to one species of
algae occurs at concentrations as low as 150 |xg/L.
The available data for nitrophenols indicate that acute toxicity to salt-
water aquatic life occurs at concentrations as low as 4,850 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of nitro-
phenols to sensitive saltwater aquatic life.
Human Health
2,4-Dinitro-O-Cresol (2-Methyl-4,6-Dinitrophenol) 534-52-1
Values to protect human health from exposure to 2,4-dinitro-o-cresol are
13.4 M-g/L through ingestion of contaminated water and organisms and
765 ng/L through ingestion of contaminated organisms only.
2,4-Dinitrophenol 51-28-5
Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for 2,4-dinitrophenol are 70 ng/L for ingestion of
contaminated water and organisms and 14,000 ng/L for ingestion of con-
taminated aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
175

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176

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NITROSAMINES
35576-91-1
CRITERIA
Aquatic Life The available data for nitrosamines indicate that acute toxicity to fresh-
water aquatic life occurs at concentrations as low as 5,850 ng/Land would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of nitro-
samines to sensitive freshwater aquatic life.
The available data for nitrosamines indicate that acute toxicity to salt-
water aquatic life occurs at concentrations as low as 3,300,000 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
nitrosamines to sensitive saltwater aquatic life.
Human Health
N-nitrosodiethylamine 55-18-5
For the maximum protection of human health from the potential carcino-
genic effects of exposure to N-nitrosodiethylamine and all other
nitrosamines, except those listed below, through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tions should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that can result in incremental increase of cancer risk
over a lifetime are estimated at 10"5,10"6, and 10"7. The corresponding rec-
ommended criteria are 8.0 ng/L, 0.8 ng/L, and 0.08 ng/L, respectively. If
these estimates are made for consumption of aquatic organisms only, ex-
cluding consumption of water, the levels are 12,400 ng/L, 1,240 ng/L, and
124 ng/L, respectively.
N-nitrosodimethylamine 62-75-9
For the maximum protection of human health from the potential carcino-
genic effects of exposure to N-nitrosodimethylamine through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Published human health criteria were recalculated using In-
tegrated Risk Information System (IRIS) to reflect available data as of
12/92 (57 F.R. 60914). Therefore, the levels that may result in incremental
increase of cancer risk over the lifetime are estimated at 10'5,10"6, and 10'7.
The corresponding recommended criteria are 0.0069 ng/L, 0.00069 ng/L,
and 0.000069 ng/L, respectively. If these estimates are made for consump-
tion of aquatic organisms only, excluding consumption of water, the levels
are 81 ng/L, 8.1 ng/L, and 0.81 ng/L, respectively.
177

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N-nitrosodibutylamine 924-16-3
For the maximum protection of human health from the potential carcino-
genic effects of exposure to N-nitrosodibutylamine through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10'5,10"6, and 10"7. The cor-
responding recommended criteria are 64 ng/L, 6.4 ng/L, and 0.64 ng/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 5,868 ng/L, 587
ng/L, and 58.7 ng/L, respectively.
N-nitrosopyrrolidine 930-55-2
For the maximum protection of human health from the potential carcino-
genic effects of exposure to N-nitrosopyrrolidine through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over a lifetime are estimated at 10"5,10"6, and 10"7, The corre-
sponding recommended criteria are 160 ng/L, 16 ng/L, and 1.6 ng/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 919,000 ng/L,
91,900 ng/L, and 9,190 ng/L, respectively.
N-nitrosodiphenylamine 86-30-6
For the maximum protection of human health from the potential carcinogenic
effects of exposure to N-nitrosodiphenylamine through ingestion of contami-
nated water and contaminated aquatic organisms, the ambient water
concentrations should be zero, based on the nonthreshold assumption for this
chemical. However, zero level may not be attainable at the present time.
Therefore, the levels that may result in incremental increase of cancer risk
over the lifetime are estimated at 10"5' 10"6, and 10"7. Published human
health criteria were recalculated using IRIS to reflect available data as of
12/92 (57 F.R. 60914). Recalculated IRIS values for N-nitrosodiphenylam-
ine are 5.0 ng/L for ingestion of contaminated water and organisms and
16.0 (xg/L for ingestion of contaminated aquatic organisms only.
N-nitrosodi-n-propylamine 62-164-7
For the maximum protection of human health from the potential carcino-
genic effects of exposure to N-nitrosodi-n-propylamine through ingestion
of contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10"5,10"6, and 10"7. Human
health criteria were calculated using IRIS to reflect available data as of
12/92 (57 F.R. 60890). Calculated values are based on 10"6 risk level for
N-nitrosodi-n-propylamine are 0.005 ng/L for ingestion of contaminated
water and organisms and 1.4 ng/L for ingestion of contaminated aquatic
organism only.
178
(45 F.R. 79318, November 28,1980) (57 F.R. 60890, December 22,1992.
See Appendix C for Human Health Methodology.

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OIL AND GREASE
CRITERIA
Domestic Water
Supply
Virtually free from oil and grease, particularly from the tastes and odors
that emanate from petroleum products.
Aquatic Life
1. 0.01 of the lowest continuous flow 96-hour LC50 to several
important freshwater and marine species, each having a
demonstrated high susceptibility to oils and petrochemicals.
2. Levels of oils or petrochemicals in the sediment that cause
deleterious effects to the biota should not be allowed.
3. Surface waters shall be virtually free from floating nonpetroleum
oils of vegetable or animal origin as well as petroleum-derived oils.
INTRODUCTION
An estimated 5 to 10 million metric tons of oil enter the marine environ-
ment annually. A major difficulty encountered in setting criteria for oil and
grease is categorization. Chemicals are not divided into categories but in-
clude thousands of organic compounds with varying physical, chemical,
and toxicological properties. They can be either volatile or nonvolatile, sol-
uble or insoluble, and persistent or easily degraded.
Rationale	Field and laboratory evidence have demonstrated both acute lethal toxicity
and long-term sublethal toxicity of oils to aquatic organisms. Events such
as the Tampico Maru wreck of 1957 in Baja, California, and the No. 2 fuel oil
spill in West Falmouth, Massachusetts, in 1969, both of which caused im-
mediate death to a wide variety of organisms, illustrate the lethal toxicity
that may be attributed to oil pollution. Similarly, a gasoline spill in South
Dakota in November 1969 was reported to have caused immediate death
to the majority of freshwater invertebrates and 2,500 fish, 30 percent of
which were native species of trout. Because of the wide range of com-
pounds included in the category of oil, establishing meaningful 96-hour
LC50 values for oil and grease without specifying the product involved is
impossible. The most susceptible category of organisms, the marine larvae,
appear to be intolerant of petroleum pollutants, particularly the water sol-
uble compounds, at concentrations as low as 0.1 mg/L.
The long-term sublethal effects of oil pollution refer to interferences
with cellular and physiological processes such as feeding and reproduction
and do not lead to immediate death of the organism. Disruption of such
behavior apparently can result from petroleum product concentrations as
low as 10 to 100 ng/L-
179

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Summaries of some of the sublethal toxicities for various petroleum
pollutants and aquatic species are contained in the 1976 criteria. In addi-
tion to sublethal effects reported at the 10 to 100 ng/L level, petroleum
products can harm aquatic life at concentrations as low as 1 ng/L.
Bioaccumulation of petroleum products presents two especially impor-
tant public health problems: (1) the tainting of edible, aquatic species, and
(2) the possibility of edible marine organisms incorporating the high boil-
ing, carcinogenic polycyclic aromatics in their tissues. Research shows that
O.Olmg/L of crude oil caused tainting in oysters. Concentrations as low as
1 to 10 jig/L could lead to tainting within very short periods of time.
Chemicals responsible for cancer in animals and humans (such as 3,4-
benzopyrene) occur in crude oil. Also, marine organisms are capable of
incorporating potentially carcinogenic compounds into their body fat
where the compounds remain unchanged.
Oil pollutants may also be incorporated into sediments. Evidence
shows that once this occurs in the sediments below the aerobic surface
layer, petroleum oil can remain unchanged and toxic for long periods,
since its rate of baterial degradation is slow. For example, No. 2 fuel oil in-
corporated into the sediments after the West Falmouth spill persisted for
over a year, and even began spreading in the form of oil-laden sediments
to more distant areas that had remained unpolluted immediately after the
spill. The persistence of unweathered oil within the sediment could have a
long-term effect on the structure of the benthic community or cause the de-
mise of specific sensitive important species.
Reports show that 0.01 mg/L oil produced deformed and inactive flat-
fish larvae and inhibition or delay of cellular division in algae by oil
concentrations of 10"4 to 10"1 mg/L. A reduction in the chemotactic percep-
tion of food by the snail, Nassarius obsoletus, at kerosene concentrations of
0.001 to 0.004 mg/L was also reported. Decreased survival and fecundity
in worms were reported at concentrations of 0.01 to 10 mg/L of detergent.
Because of the great variability in the toxic properties of oil, establish-
ing a numerical criterion applicable to all types of oil is difficult. Thus, an
application factor of 0.01 of the 96-hour LC50 as determined by using con-
tinuous flow with a sensitive resident species should be employed for
individual petrochemical components.
Toxicological data is sparse on the ingestion of the components of re-
finery wastewaters by humans or by test animals. Any tolerable health
concentrations for petroleum-derived substances far exceed the limits of
taste and odor. Since petroleum derivatives become organoleptically objec-
tionable at concentrations far below the human chronic toxicity, hazards to
humans will not likely arise from drinking oil-polluted waters. Oils of ani-
mal or vegetable origin generally are nontoxic to humans and aquatic life.
In view of the problem of petroleum oil incorporated in sediments, its
persistence and chronic toxic potential, and the present lack of sufficient
toxicity data to support specific criteria, oil concentrations in sediments
should not approach levels that cause deleterious effects to important spe-
cies or the bottom community as a whole.
Petroleum and nonpetroleum oils share some similar physical and
chemical properties. Because they share common properties, they may
cause similar harmful effects in the aquatic environment by forming a
sheen, film, or discoloration on the water surface. Like petroleum oils, non-
petroleum oils may occur at four levels of the aquatic environment: (a)
floating on the surface, (b) emulsified in the water column, (c) solubilized,
and (d) settled on the bottom as a sludge. Analogous to the grease balls

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from vegetable oil and animal fats are the tar balls of petroleum origin that
have been found in the marine environment or washed ashore on beaches.
Oils of any kind can cause (a) drowning of water fowl because of loss
of buoyancy, exposure because of loss of insulating capacity of feathers,
and starvation and vulnerability to predators because of lack of mobility;
(b) lethal effects on fish by coating epithelial surface of gills, thus prevent-
ing respiration; (c) potential fishkills resulting from biochemical oxygen
demand; (d) asphyxiation of benthic life forms when floating masses be-
come engaged with surface debris and settle on the bottom; an (e) adverse
aesthetic effects of fouled shorelines and beaches. These and other effects
have been documented in the U.S. Department of Health, Education and
Welfare report on Oil Spills Affecting the Minnesota and Mississippi Rivers
in the 1975 Proceedings of the Joint Conference on Prevention and Control
of Oil Spills.
Oils of animal or vegetable origin generally are chemically nontoxic to
humans or aquatic life; however, floating sheens of such oils result in dele-
terious environmental effects described in this criterion. Thus, surface
waters should be virtually free from floating nonpetroleum oils of vegeta-
ble or animal origin.This same recommendation applies to floating oils of
petroleum origin, since they, to, may produce similar effects.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
181

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182

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PARATHION
56-38-2
Freshwater — 1-hour average of 0.065 ng/L
4-day average of 0.013 ng/L
The acute values for 37 freshwater species in 31 genera range from 0.04
Hg/L for an early instar of a crayfish (Orconectes nais) to 5,230 ng/L for two
species of tubificid worms. For Daphnia magna, the chronic value and
acute- chronic ratio are 0.0990 ng/L and 10.10, respectively. Chronic toxic-
ity values are available for two freshwater fish species, the bluegill
(Lepomis macrochirus) and the fathead minnow (Pimephales promelas), with
chronic values of 0.24 (ig/L and 6.3 ng/L, and acute-chronic ratios of 2,121
and 79.45, respectively. Two freshwater algae were affected by toxaphene
concentrations of 30 and 390 ng/L, respectively. Bioconcentration factors
determined with three fish species ranged from 27 to 573.
The acute values available for saltwater species are 11.5 and 17.8 fig/L
for the Korean shrimp, Palaemon macrodactylus, and 17.8 n-g/L for the
striped bass (Morone saxatilis). No data are available concerning the
chronic toxicity of parathion to saltwater species, toxicity to saltwater
plants, or bioaccumulation by saltwater species. Some data indicate that
parathion is acutely lethal to commercially important saltwater shrimp at
concentrations as low as 0.24 ng/L. Measurement of acetylcholinesterase
(AChE) in fish tissue might be useful for diagnosing fish kills caused by
parathion.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of para-
thion does not exceed 0.013 ng/L more than once every three years on the
average and if the one-hour average concentration does not exceed 0.065
M-g/L more than once every three years on the average.
The procedures described in the guidelines require the availability of
specified data for the derivation of a criterion. A saltwater criterion for
parathion cannot be derived because very few of the required data are
available.
In the Agency's best scientific judgment, three years is the average time
aquatic ecosystems should be provided between excursions. The resil-
iences of ecosystems and their abilities to recover differ greatly, however,
and site-specific allowed excursion frequencies may be established if ade-
quate justification is provided.
When developing water quality-based permit limits and for designing
waste treatment facilities, criteria must be based upon an appropriate
CRITERIA
Aquatic Life
Summary
183

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wasteload allocation model. Dynamic models are preferred; but if limited
data or other considerations might make their use impractical, rely on
steady-state models.
(51 F.R. 43665, December 3,1986)
See Appendix A for Aquatic Life Methodology.
184

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PENTACHLOROPHENOL (PCP)
87-86-5
CRITERIA 	
Not to exceed 13.0 ng/L in salt water.
7.9 (ig/L for saltwater aquatic life.
Freshwater criteria are pH dependent. See text.
Summary	The acute and chronic toxicity of PCP to freshwater animals increased as
pH and dissolved oxygen concentration of the water decreased. Generally,
toxicity also increases with increased temperature. The estimated acute
sensitivities of 36 species at pH = 6.5 ranged from 4.355 |xg/L for larval
common carp to 43,920 pig/L for a crayfish. At pH = 6.5, the lowest and
highest estimated chronic values of < 1.835 and 79.66 ng/L, respectively,
were obtained with different cladoceran species. Chronic toxicity to fish
was affected by the presence of impurities, with industrial-grade PCP
being more toxic than purified samples. Mean acute-chronic ratios for six
freshwater species ranged from 0.8945 to 15.79, but the mean ratios for the
four most acutely sensitive species only ranged from 0.8945 to 5.035. Fresh-
water algae were affected by concentrations as low as 7.5 ng/L, whereas
vascular plants were affected at 189 ng/L and above. Bioconcentration fac-
tors ranged from 7.3 to 1,066 for three species of fish.
Acute toxicity values from tests with 18 species of saltwater animals,
representing 17 genera, range from 22.63 ng/L for late yolk-sac larvae of
the Pacific herring (Clupea harengus pallasi) to 18,000 ng/L for adult blue
mussels (Mytllus edulis). The embryo and larval stages of invertebrates and
the late larval premetamorphosis stage of fish appear to be the most sensi-
tive life stages to PCP. With few exceptions, fish are more sensitive than
invertebrates to PCP. Salinity, temperature, and pH have a slight effect on
the toxicity of PCP to some saltwater animals.
Life-cycle toxicity tests have been conducted with the sheepshead min-
now (Cyprinodon variegatus) and the polychaete worm (Ophryotrocha
diadema). The chronic value for the minnow is 64.31 jig/L and the acute-
chronic ratio is 6.873. Unfortunately, no statistical analysis of the worm test
data is available.
The EC50s for saltwater plants range from 17.40 ng/L for the diatom,
Skeletonema costatum, to 3,600 ng/L for the green alga (Dunaliella ter-
tiolecta). Apparent steady-state BCFs are available for the eastern oyster
(Crassostrea virginica) and two saltwater fishes and range from 10 to 82.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration (in
Hg/L) of pentachlorophenol does not exceed the numerical value given by
g[1.005(pH)-5.290]
185

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more than once every three years on the average and if the one-hour
average concentration (in fig/L) does not exceed the numerical value given
by
g[1,005(pH)-4.830]
more than once every three years on the average. For example, at pH = 6.5,
7.8, and 9.0, the four-day average concentrations of pentachlorophenol are
3.5,13, and 43 ng/L, respectively, and the one-hour average concentrations
are 5.5,20, and 68 fig/L. At pH = 6.8, a pentachlorophenol concentration of
1.74 ng/L caused a 50 percent reduction in the growth of yearling sockeye
salmon (Oncorhynchus nerka) in a 56-day test.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of pentachlorophenol does not exceed 7.9 (ig/L
more than once every three years on the average and if the one-hour aver-
age concentration does not exceed 13 ng/L more than once every three
years on the average.
In the Agency's best scientific judgment, three years is the average time
aquatic ecosystems should be provided between excursions. The resil-
iences of ecosystems and their abilities to recover differ greatly, however,
and site-specific allowed excursion frequencies may be established if ade-
quate justification is provided.
When developing water quality-based permit limits and designing
waste treatment facilities, criteria must be based upon an appropriate
wasteload allocation model. Dynamic models are preferred, but if limited
data or other considerations might make their use impractical, rely on
steady-state models.
(51 F.R. 43665, December 3,1986)
See Appendix A for Aquatic Life Methodology.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect data available as of 12/92 (57 F.R. 60840). Recalcul-
ated IRIS values for pentachlorophenol are 0.28 ng/L for ingestion of
contaminated water and organisms and 8.2 ng/L for ingestion of contami-
nated aquatic organisms only.
Using available organoleptic data, the estimated level is 30 jig/L to
control undesirable taste and odor qualities of ambient water. Organolep-
tic data do have limitations as a basis for establishing a water quality
criterion but no demonstrated relationship to potentially adverse effects on
human health.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
186

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PH
CRITERIA
pH Range
Introduction
•	5-9 for domestic water supplies.
•	6.5-9.0 for freshwater aquatic life.
•	6.5-8.5 for marine aquatic life (but not more than 0.2 units outside of
normally occurring range).
"pH" measures the hydrogen ion activity in a water sample. It is mathe-
matically related to hydrogen ion activity according to the expression:
pH = -Log10 (H+), where (H+) is the hydrogen ion activity.
The pH of natural waters is a measure of acid base equilibrium
achieved by the various dissolved compounds, salts, and gases. The princi-
pal system regulating pH in natural waters is the carbonate system, which
is composed of carbon dioxide (CO2), carbonic acid (H2CO3), bicarbonate
ion (HC03), and carbonate ions (CO3). The interactions and kinetics of this
system have been described by scientists.
pH is an important factor in the chemical and biological systems of
natural waters. The degree of dissociation of weak acids or bases is af-
fected by changes in pH, which is important because the toxicity of many
compounds is affected by the degree of dissociation. One such example is
hydrogen cyanide (HCN): cyanide toxicity to fish increases as the pH is
lowered because the chemical equlibrium is shifted toward an increased
concentration of HCN. Similar results have been shown for hydrogen sul-
fide (H2S).
pH also affects the solubility of metal compounds contained in bottom
sediments or as suspended material. For example, laboratory equilibrium
studies under anaerobic conditions indicated that pH was an important
parameter involved in releasing manganese from bottom sediments.
The pH of a waterbody does not indicate ability to neutralize additions
of acids or bases without appreciable change. This characteristic, termed
"buffering capacity," is controlled by the amounts of alkalinity and acidity
present.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
187

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188

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PHENOL
108-95-2
CRITERIA
Aquatic Life
Human Health
The available data for phenol indicate that acute and chronic toxicity to
freshwater aquatic life occurs at concentrations as low as 10,200 ng/L and
2,560 |ig/L, respectively, and would occur at lower concentrations among
species that are more sensitive than those tested.
The available data for phenol indicate that toxicity to saltwater aquatic
life occurs at concentrations as low as 5,800 ng/L and would occur at lower
concentrations among species that are more sensitive than those tested. No
data are available concerning the chronic toxicity of phenol to sensitive
saltwater aquatic life.
Published human health criteria were recalculated using Integrated Risk
Information System (IRIS) to reflect data available as of 12/92
(57 F.R. 60912). Recalculated IRIS values for phenol are 21,000 ng/L for in-
gestion of contaminated water and organisms and 4,600,000 ng/L for
ingestion of contaminated aquatic organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60912, December 22,1992)
See Appendix C for Human Health Methodology.
189

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190

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PHOSPHORUS
7723-14-0
CRITERION 	
0.10 ng/L yellow (elemental) phosphorus for estuarine and saltwater
aquatic life.
Introduction Phosphorus in its elemental form is particularly toxic and subject to
bioaccumulation in much the same way as mercury. Phosphorus as phos-
phate, however, is one of the major nutrients required for plant nutrition
and essential for life. In excess of a critical concentration, phosphates stim-
ulate plant growths.
During the past 30 years, the belief has developed that increasing
standing crops of aquatic plants, which often interfere with water uses and
are nuisances to humans, frequently are caused by increasing phosphorus
supplies. Such phenomena are associated with a condition of accelerated
eutrophication or aging of waters. Phosphorus is not the sole cause of eu-
trophication, but evidence suggests that frequently it is the key element of
all elements required by freshwater plants. Generally, it is present in the
least amount relative to need. Therefore, an increase in phosphorus allows
use of other already present nutrients for plant growth. Further, of all ele-
ments required for plant growth in the water environment, phosphorus is
most easily controlled by humans.
Large deposits of phosphate rock are found near the western shore of
central Florida, as well as in a number of other States. Deposits in Florida
are found in the form of pebbles embedded in a matrix of clay and sand
that vary in size from fine sand to about the size of a human foot. The
phosphate rock beds lie within a few feet of the surface and are mined by
using hydraulic water jets and a washing operation that separates the
phosphates from waste materials, a process similar to strip-mining. Flor-
ida, Idaho, Montana, North Carolina, South Carolina, Tennessee, Utah,
Virginia, and Wyoming all mine phosphate.
Phosphates enter waterways from several different sources. The
human body excretes about one pound per year of phosphorus, expressed
as "P." The use of phosphate detergents and other domestic phosphates in-
creases the per capita contribution to about 3.5 pounds per year of
phosphorus as P. Some industries, such as potato processing, have
wastewaters high in phosphates. Cropland, forestland, and idle and urban
land contribute varying amounts of phosphorus-diffused sources in drain-
age to watercourses: surface runoff of rainfall, effluent from tile lines, or
return flow from irrigation. Other contributing sources are cattle feedlots,
concentrations of domestic or wild ducks, tree leaves, and fallout from the
atmosphere.
Evidence indicates that high phosphorus concentrations are associated
with accelerated eutrophication of waters when other growth-promoting
factors are present; aquatic plant problems develop in reservoirs and other
standing waters with phosphorus values lower than those critical in flow-
ing streams; reservoirs and lakes collect phosphates from influent streams
191

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and store a portion of them within consolidated sediments, thus serving as
a phosphate sink; and phosphorus concentrations critical to noxious plant
growth vary. Therefore, nuisance growths may result from a particular
concentration of phosphate in one geographical area but not in another.
The amount or percentage of inflowing nutrients that may be retained
by a lake or reservoir is variable and will depend upon the following:
•	The nutrient loading to the lake or resevoir;
•	The volume of the euphotic zone;
•	The extent of biological activities;
•	The detention time within a lake basin or the time available for
biological activities; and
•	The level of discharge from the lake or of the penstock from the
reservoir.
Once nutrients are combined within the aquatic ecosystem, their re-
moval is tedious and expensive. Phosphates are used by algae and higher
aquatic plants and excess may be stored within the plant cell. With decom-
position of the plant cell, some phosphorus may be released immediately
through bacterial action for recycling within the biotic community, while
the remainder may be deposited with sediments. Much of the material that
combines with the consolidated sediments within the lake bottom is bound
permanently and will not be recycled into the system.
Rationale
Elemental Phosphorus
Isom (1960) reported and LC50 of 0.105 mg/L at 48 hours and 0.025 mg/L
at 160 hours for bluegill sunfish (Lepomis macrochirus) exposed to yellow
phosphorus in distilled water at 26°C and pH 7. The 125- and 195-hour
LCsos of yellow phosphorus to Atlantic cod (Gadus morhua) and Atlantic
salmon (Salmo salar) smolts in continuous-exposure experiments was 1.89
and 0.79 ng/L, respectively. No evidence of an incipient lethal level was
observed since the lowest concentration of P4 tested was 0.79 fig/L.
Salmon that were exposed to elemental phosphorus concentration of 40
Hg/L or less developed a distinct external red color and showed signs of
extensive hemolysis. The predominant features of P4 poisoning in salmon
were external redness, hemolysis, and reduced hematocrits.
Following the opening of an elemental phosphorus production plant in
Long Harbour, Placentia Bay, Newfoundland, divers observed dead fish
upon the bottom throughout the harbor. Mortalities were confined to a
water depth of less than 18 meters. Visual evidence showed selective mor-
tality among benthos. Live mussels were found within 300 meters of the
effluent pipe, while all scallops within this area were dead.
Fish will concentrate elemental phosphorus from water containing as
little as 1 ng/L. In one set of experiments, a cod swimming in water con-
taining 1 ng/L elemental phosphorus for 18 hours concentrated
phosphorus to 50 ng/L in muscle, 150 ng/kg in fatty tissue, and 25,000
Hg/kg in the liver. The experimental findings showed that phosphorus is
quite stable in the fish tissues.
The criterion of 0.10 jxg/L elemental phosphorus for marine or estua-
rine waters in 1/10 of demonstrated lethal levels to important marine
organisms and of levels found to result in significant bioaccumulation.

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Phosphate Phosphorus
Although a total phosphorus criterion to control nuisance aquatic growths
is not presented, the following rationale to support such a criterion, which
currently is evolving, should be considered.
Total phosphate phosphorus concentrations in excess of 100 jig/L P
may interfere with coagulation in water treatment plants. When such con-
centrations exceed 25 ng/L at the time of the spring turnover on a
volume-weighted basis in lakes or reservoirs, they may occasionally stimu-
late excessive (nuisance) growths of algae and other aquatic plants. Algal
growths inpart undesirable tastes and odors to water, interfere with water
treatment, become aesthetically unpleasant, and alter the chemistry of the
water supply. They contribute to the phenomenon of cultural eutrophica-
tion.
To prevent the development of biological nuisances and to control ac-
celerated or cultural eutrophication, total phosphates as phosphorus (P)
should not exceed 50 ng/L in any stream at the point where it enters any
lake or reservoir, nor 25 ng/L within the lake or reservoir. A desired goal
for the prevention of plant nuisances in streams or other flowing waters
not discharging directly to lakes or impoundments is 100 |xg/L total P.
Most relatively uncontaminated lake districts are known to have surface
waters that contain from 10 to 30 ^/L total phosphorus as P.
The majority of the Nation's eutrophication problems are associated
with lakes or reservoirs. Currently, more data support establishing a limit-
ing phosphorus level in those waters than in streams or rivers that do not
directly impact such water. Some natural conditions, also, would dictate
whether a more or less stringent phosphorus level should be considered.
Eutrophication problems may occur in waters where the phosphorus con-
centration is less than that indicated previously. Obviously, such waters
would need more stringent nutrient limits. Likewise, in some waters phos-
phorus is not now a limiting nutrient and the need for phosphorus limits is
substantially diminished.
Establishing a phosphorus criterion for flowing waters requires two
basic needs: one is to control the development of plant nuisances within
the flowing water and, in turn, to control and prevent animal pests that
may become associated with such plants; the other is to protect the down-
stream receiving waterway, regardless of its proximity in linear distance. A
portion of the phosphorus that enters a stream or other flowing waterway
eventually will reach a receiving lake or estuary either as a component of
the fluid mass, as bed load sediments carried downstream, or as floating
organic materials drifting just above the streambed or floating on the
water's surface. Superimposed on the loading from the inflowing water-
way, a lake or estuary can receive additional phosphorus as fallout from
the air shed or as a direct introduction from shoreline areas.
Another method to control the inflow of nutrients, particularly phos-
phates, into a lake is that of prescribing an annual loading to the receiving
water. Vollenweider suggests total phosphorus loadings in grams per
square meter of surface area per year as a critical level for eutrophic condi-
tions within the receiving waterway for a particular water volume where
the mean depth of the lake in meters is divided by the hydraulic detention
time in years. Vollenweider's data suggest a range of loading values that
should result in oligotrophic lake water quality (see Table 1).
In some waterways, higher concentrations or loadings of total phos-
phorus do not produce eutrophy or lower concentrations or loadings of
total phosphorus may produce nuisance organisms. Therefore, waters now
193

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Table 1.—Annual loadings.
MEAN DEPTH/HYDRAUUC
DETENTION TIME
(meters/year)
OLIGOTROPHY OR
PERMISSIBLE LOADING
(grams/meter^/year)
EUTROPHIC OR CRITICAL
LOADING
(grams/meter^/year)
0.5
0.07
0.14
1.0
0.10
0.20
2.5
0.16
0.32
5.0
0.22
0.45
7.5
0.27
0.55
10.0
0.32
0.63
25.0
0.50
1.00
50.0
0.71
1.41
75.0
0.87
1.73
100.0
1.00
2.00
Source: Vollenweider (1973).
containing less than the specified amounts of phosphorus should not be
degraded by the introduction of additional phosphates.
The following specific exceptions can reduce the threat of phosphorus
as a contributor to lake eutrophy:
•	Naturally occurring phenomena may limit the development of
plant nuisances;
•	Technological or cost-effective limitations may help control
introduced pollutants;
•	Waters may be highly laden with natural silts or colors that reduce
the penetration of sunlight needed for plant photosynthesis;
•	Some waters' morphometric features — steep banks, great depth,
and substantial flows — contribute to a history of no plant
problems;
•	Waters may be managed primarily for waterfowl or other wildlife;
•	In some waters, a nutrient other than phosphorus limits plant
growth; the level and nature of such a limiting nutrient would not
be expected to increase to an extent that would influence
eutrophication; and
•	In some waters, phosphorus control cannot be sufficiently effective
under present technology to make phosphorus the limiting
nutrient.
No national criterion is presented for phosphate phosphorus for the
control of eutrophication.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.

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PHTHALATE ESTERS
CRITERIA
Aquatic Life The available data for phthalate esters indicate that acute and chronic tox-
icity to freshwater aquatic life occurs at concentrations as low as and
Hg/L, respectively, and would occur at lower concentrations among
species that are more sensitive than those tested.
The available data for phthalate esters indicate that acute toxicity to
saltwater aquatic life occurs at concentrations as low as Hg/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
phthalate esters to sensitive saltwater aquatic life, but toxicity to one spe-
cies of algae occurs at concentrations as low as	M-g/L.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60913). Recalcul-
ated IRIS values for diethyl phthalate are 23,000 ^ig/L for ingestion of
contaminated water and organisms and 120,000 ng/L for ingestion of con-
taminated aquatic organisms only.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60913). Recalculated IRIS values for dibutyl
phthalate are 2,700 y.g/L for ingestion of contaminated water and organ-
isms and 12,000 ng/L for ingestion of contaminated aquatic organisms
only.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60890). Recalculated IRIS values for di-2-
ethylhexyl phthalate are 1.8 ng/L for ingestion of contaminated water and
organisms and 5.9 |Ag/L for ingestion of contaminated aquatic organisms
only. IRIS values are based on a 10"6 risk level for carcinogens.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60913). Recalculated IRIS values for dimethyl
phthalate are 313,000 |xg/L for ingestion of contaminared water organisms
and 2,900,000 ng/L for ingestion of contaminated organisms only. IRIS val-
ues are based on a 10'6 risk level for carcinogens.
Human health criteria were recalculated using IRIS to reflect available
data as of 12/92 (57 F.R. 60890). Recalculated IRIS values for butylbenzyl
phthalate are 3,000 ng/L for ingestion of contaminated water and organ-
isms and 5,200 |xg/L for ingestion of contaminated organisms only. IRIS
values are based on a 10"6 risk level for carcingens.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
195

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196

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POLYCHLORINATED BIPHENYLS (PCBs)
1336-36-3
CRITERIA 	
Not to exceed 2.0 ng/L in fresh water or 10.0 ng/L in salt water.
0.014 and 0.03 mg/L for freshwater and saltwater aquatic life, respec-
tively.
Aquatic Life For polychlorinated biphenyls, the criterion to protect freshwater aquatic
life as derived using the guidelines is 0.014 (ig/L as a 24-hour average. This
concentration is probably too high because it is based on bioconcentration
factors measured in laboratory studies; however, field studies produce fac-
tors at least 10 times higher for fishes. The available data indicate that
acute toxicity to freshwater aquatic life probably will occur only at concen-
trations above 2.0 ng/L; therefore, the 24-hour average should provide
adequate protection against acute toxicity.
The criterion to protect saltwater aquatic life for polychlorinated bi-
phenyls as derived using the guidelines is 0.030 |xg/L as a 24-hour average.
This concentration is probably too high because it is based on
bioconcentration factors measured in laboratory studies; however, field
studies produce factors at least 10 times higher for fishes. The available
data indicate that acute toxicity to saltwater aquatic life probably will only
occur at concentrations above 10 ng/L; therefore, 24-hour average criterion
should provide adequate protection against acute toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to polychlorinated biphenyls through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time.
Published human health criteria were recalculated using Integrated
Risk Information System (IRIS) to reflect available data as of 12/92
(57 F.R. 60915). Recalculated IRIS values for PCBs 1016, 1221, 1232, 1242,
1248,1254, and 1260 are estimated at 10"6 risk level. The recommended cri-
teria for consumption of contaminated water and organisms is 0.000044
Hg/L and 0.000045 ng/L for organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60915, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
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198

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POLYNUCLEAR AROMATIC
HYDROCARBONS
CRITERIA
Aquatic Life The limited freshwater database, available mostly from short-term
bioconcentration studies with two compounds, does not permit a state-
ment concerning acute or chronic toxicity for polynuclear aromatic
hydrocarbons.
The data that are available indicate that acute toxicity to saltwater
aquatic life occurs at concentrations as low as 300 ng/L and would occur at
lower concentrations among species that are more sensitive than those
tested. No data are available concerning these hydrocarbons' chronic toxic-
ity to sensitive saltwater aquatic life.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to polynuclear aromatic hydrocarbons through
ingestion of contaminated water and. contaminated aquatic organisms, the
ambient water concentration should be zero, based on the nonthreshold as-
sumption for this chemical. However, zero level may not be presently
attainable. Therefore, the levels that may result in incremental increase of
cancer risk over a lifetime are estimated at 10"5,10"6, and 10"7, with the cor-
responding recommended criteria 28.0 ng/L, 2.8 ng/L, and 0.28 ng/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 311.0 ng/L, 31.1
ng/L, and 3.11 ng/L, respectively.
Human health criteria for three polynuclear aromatic hydrocarbons —
acenaphylene, phenanthrene, and benzo (g,h,i) perylene — have been de-
leted (see 57 F.R. 60887, December 22, 1992). Although the water quality
criteria for these compounds have been deleted, information in the 1980
document may be useful.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
199

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200

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SELENIUM
7782-49-2
CRITERIA 	
Not to exceed 20 ng/L in fresh water or 300 jig/L in salt water.
5.0 ng/L and 71 ng/L for freshwater and saltwater aquatic life, respec-
tively.
Implementation Because of the variety of forms of selenium in ambient water and the lack
of definitive information about their relative toxicities to aquatic species,
no available analytical measurement is known to be ideal for expressing
aquatic life criteria for selenium. Previous aquatic life criteria for metals
and metalloids were expressed in terms of the total recoverable measure-
ment, but newer criteria for metals and metalloids were expressed in terms
of the acid-soluble measurement. Acid-soluble, selenium — operationally
defined as the selenium that passes through a 0.45 um membrane filter
after the sample has been acidified to a pH between 1.5 and 2.0 with nitric
acid — is probably the best measurement at the present time.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of sele-
nium does not exceed 5.0 ng/L more than once every three years on the
average, and if the one-hour average concentration does not exceed 20
jig/L more than once every three years on the average.
The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms
and Their Uses" indicate that, except possibly where a locally important
species is very sensitive, saltwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of sele-
nium does not exceed 71 |xg/L more than once every three years on the
average and if the one-hour average concentration does not exceed 300
Hg/L more than once every three years on the average. If selenium is as
toxic to saltwater fishes as it is to freshwater fishes in the field, the status of
the fish community should be monitored whenever the concentration of
selenium exceeds 5 fig/L in salt water.
Human Health Human health criteria have been withdrawn for this compound (see 57
F.R. 60885, December 22, 1992). Although the human health criteria are
withdrawn, EPA published a document for this compound that may con-
tain useful human health information. This document was originally
noticed in 45 F.R. 79331, November 28, 1980.
(45 F.R. 79331, November 28,1980) (53 F.R. 177, January 5,1988)
(57 F.R. 60911, December 22,1992)
See Appendix A for Aquatic Life Methodology.
201

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202

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SILVER
7440-22-4
CRITERIA 	
Not to exceed 2.3 ng/L in salt water.
Freshwater values are hardness dependent. See text.
0.12 ng/L for freshwater aquatic life.
Aquatic Life For freshwater aquatic life, the concentration (in pig/L) of total recoverable
silver should not exceed the numerical value given by
(1.72[ln(hardness)]-6.52)
V
at any time. For example, at hardnesses of 50, 100, and 200 mg/L as
CaCC^ the concentration of total recoverable silver should not exceed 1.2,
4.1, and 13 M-g/U respectively, at any time. The available data indicate that
chronic toxicity to freshwater aquatic life may occur at concentrations as
low as 0.12 ng/L.
For saltwater aquatic life, the concentration of total recoverable silver
should not exceed 2.3 ng/L at any time. No data are available concerning
the chronic toxicity of silver to sensitive saltwater aquatic life.
Human Health Human health criteria have been withdrawn for this compound (see
57 F.R. 60885, December 22, 1992). Although the human health criteria are
withdrawn, EPA published a document for this compound that may con-
tain useful human health information. This document was originally
noticed in 45 F.R. 79318, November 28,1980.
(45 F.R. 79318, November 28,1980) (57 F.R. 60911, December 22,1992)
See Appendix B for Aquatic Life Methodology.
See Appendix C for Human Health Methodology.
203

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204

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SOLIDS (SUSPENDED, SETTLEABLE)
AND TURBIDITY
CRITERIA
Freshwater Fish
and Other	Settleable and suspended solids should not reduce the depth of the com-
Aquatic Life pensation point for photosynthetic activity by more than 10 percent from
the seasonally established norm for aquatic life.
Introduction The term "suspended and settleable solids" describes the organic and inor-
ganic particulate matter in water. The equivalent terminology used for
solids in "Standard Methods" is "total suspended matter" for suspended
solids, "settleable matter" for settleable solids, "volatile suspended mat-
ter"' for volatile solids, and "fixed suspended matter" for fixed suspended
solids. The term "solids" is used in this discussion because of its more
common use in the water pollution control literature.
Rationale	Suspended solids and turbidity are important parameters in both munici-
pal and industrial water supply practices. Finished drinking waters have a
maximum limit of 1 turbidity unit where the water enters the distribution
system. This limit is based on health considerations as they relate to effec-
tive chlorine disinfection. Suspended matter provides areas where
microorganisms do not come into contact with the chlorine disinfectant.
The ability of common water treatment processes (i.e., coagulation, sedi-
mentation, filtration, and chlorination) to remove suspended matter to
achieve acceptable final turbidities is a function of the material's composi-
tion as well as its concentration. Because of the variability of such removal
efficiency, general raw water criterion for these uses cannot be delineated.
Turbid water interferes with recreational use and aesthetic enjoyment
of water. It can be dangerous for swimming, especially if diving facilities
are provided, because of the possibility of unseen submerged hazards and
the difficulty in locating swimmers in danger of drowning. The less turbid
the water, the more desirable it becomes for swimming and other water
contact sports. Other recreational pursuits, such as boating and fishing,
will be adequately protected by suspended solids criteria developed for
protection of fish and other aquatic life.
Fish and other aquatic life requirements concerning suspended solids
can be divided into those whose effect occurs in the water column and
those whose effect occurs following sedimentation to the bottom of the
waterbody. Noted effects are similar for both fresh and marine waters.
The effects of suspended solids on fish have been reviewed by the Eu-
ropean Inland Fisheries Advisory Commission. This 1965 review identified
four effects on fish and fish food populations:
1. By acting directly on the fish swimming in water in which solids
are suspended,-and either killing them or reducing their growth
rate, resistance to disease, etc.;
205

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2.	By preventing the successful development of fish eggs and larvae;
3.	By modifying natural movement and migrations of fish; and
4.	By reducing the abundance of food available to the fish.
Settleable materials that blanket the bottom of waterbodies damage the
invertebrate populations, block gravel spawning beds, and, if organic, re-
move dissolved oxygen from overlying waters. In a study downstream
from the discharge of a rock quarry where inert suspended solids were in-
creased to 80 mg/L, the density of macroinvertebrates decreased by 60
percent; in areas of sediment accumulation, benthic invertebrate popula-
tions also decreased by 60 percent regardless of the suspended solid
concentrations. Similar effects have been reported downstream from an
area that was intensively logged. Major increases in stream suspended sol-
ids (25 ppm turbidity upstream, versus 390 ppm downstream) caused
smothering of bottom invertebrates, reducing organism density to only 7.3
per square foot, versus 25.5 per square foot upstream.
When settleable solids block gravel spawning beds that contain eggs,
high mortalities result although evidence suggests that some species of sal-
monids will not spawn in such areas.
Silt attached to the eggs may prevent sufficient exchange of oxygen
and carbon dioxide between the egg and the overlying water. The impor-
tant variables are particle size, stream velocity, and degree of turbulence.
Deposition of organic materials to the bottom sediments can cause im-
balances in stream biota by increasing bottom animal density, principally
worm populations; diversity is reduced as pollution-sensitive forms disap-
pear. Algae likewise flourish in such nutrient-rich areas, although forms
may become less desirable.
Plankton and inorganic suspended materials reduce light penetration
into the waterbody, reducing the depth of the photic zone. This reduces
primary production and decreases fish food. In 1974 the National Acad-
emy of Sciences recommended that the depth of light penetration not be
reduced by more than 10 percent. Additionally, the near surface waters are
heated because of the greater heat absorbency of the particulate material,
which tends to stabilize the water column and prevents vertical mixing.
Such mixing reductions decrease the dispersion of dissolved oxygen and
nutrients to lower portions of the waterbody.
One partially offsetting benefit of suspended inorganic material in
water is the sorption of organic materials such as pesticides. Following this
sorption process, subsequent sedimentation may remove these materials
from the water column into the sediments.
Identifiable effects of suspended solids on irrigation use of water in-
clude the formation of crusts on top of the soil, which inhibit water
infiltration and plant emergence and impeded soil aeration; the formation
of films on plant leaves that blocks sunlight and impedes photosynthesis
and that may reduce the marketability of some leafy crops like lettuce; and
finally, the adverse effect on irrigation reservoir capacity, delivery canals,
and other distribution equipment.
The criteria for freshwater fish and other aquatic life are essentially
that proposed by the National Academy of Sciences and the Great Lake
Water Quality Board.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.

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SULFIDE - HYDROGEN SULFIDE
7783-06-4
CRITERIA 	
2 (ig/L undissociated H2S for fish and other aquatic life in either fresh
water or salt water.
Introduction Hyrogen sulfide is a soluble, highly poisonous, gaseous compound that
smells like rotten eggs. Humans can detect it in air at a dilution of 0.002
ppm. It will dissolve in water at 4,000 mg/L at 20°C and one atmosphere of
pressure. Biologically, hydrogen sulfide is an active compound found pri-
marily as an anaerobic degradation product of both organic sulfur
compounds and inorganic sulfates. Sulfides are constituents of many in-
dustrial wastes such as those from tanneries, paper mills, chemical plants,
and gas works. The anaerobic decomposition of sewage, sludge beds,
algae, and other naturally deposited organic material is a major source of
hydrogen sulfide.
When soluble sulfides are added to water, they react with hydrogen
ions to form HS" or H2S, the proportion of each depending on the pH. The
toxicity of sulfides derives primarily from H2S rather than from the hydro-
sulfide (HS") or sulfide (S=) ions.
When hydrogen sulfide dissolves in water it dissociates according to
the following reactions:
H2S*>HS" + H+ and HS" ~S'2 + H+
At pH 9, about 99 percent of the sulfide is in the form of HS", at pH 7
the sulfide is equally divided between HS' and H2S; and at pH 5 about 99
percent of the sulfide is present as H2S. Investigators have minimized the
toxic effects of H2S on fish and other aquatic life because H2S is oxidized in
well-aerated water by natural biological systems to sulfates or is biologi-
cally oxidized to elemental sulfur.
Rationale	The degree of hazard exhibited by sulfide to aquatic animal life is depen-
dent on the temperature, pH, and dissolved oxygen. AT lower pH values, a
greater proportion is in the form of the toxic undissociated H2S. In winter
when the pH is neutral or below or when dissolved oxygen levels are low
but not lethal to fish, the hazard from sulfides is exacerbated. Fish exhibit a
strong avoidance reaction to sulfide. Based on data from experiments with
the stickleback, if fish encounter a lethal concentration of sulfide, reason-
able chance exists that they will be repelled by it before they are harmed.
This, of course, assumes that an escape route is open.
Many past data on the toxicity of hydrogen sulfide to fish and other
aquatic life have been based on extremely short exposure periods. Conse-
quently, these early data have indicated that concentrations between 0.3
and 0.4 mg/L permit fish to survive. Recent long-term data, both in field
situations and under controlled laboratory condition, demonstrate hydro-
gen sulfide toxicity at lower concentrations.
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Concentrations as high as 0.7 mg/L have been found within 20 mm of
the bottom of sludge beds, and the levels of 0.1 to 0.02 mg/L were common
within the first 20 mm of water above this layer. Walleye (Stizostedion vit-
reum) eggs held in trays in this zone did not hatch. The hatchability of
northern pike (Esox lucius) eggs was substantially reduced at 25 ng/L at
H2S; at 47 ng/L, mortality was almost complete. Northern pike fry had 96-
hour LC50 values that varied from 17 to 32 ng/L at normal oxygen levels of
6.0 mg/L. The highest concentration of hydrogen sulfide that had no ob-
servable effect on eggs and fry was 14 and 4 (j.g/L, respectively. Eggs, fry,
and juveniles of walleyes and white suckers (Catostomus commersoni) and
safe levels in working on walleyes and fathead minnows (Pimephales pro-
melas) were found to vary from 2.9 ng/L to 12 ng/L, with eggs being the
least sensitive and juveniles being the most sensitive in short-term test. In
96-hour bioassays, fathead minnows and goldfish (Carassius auratus) var-
ied greatly in tolerance to hydrogen sulfide with changes in temperature.
They were more tolerant at low temperatures (6 to 10°C). In addition, 1.0
mg/L sulfide caused 100 percent mortality in 72 hours with Pacific salmon.
On the basis of chronic tests evaluating growth and survival, the safe
H2S level for bluegill (Lepomis macrochirus) juveniles and adults was 2
Hg/L. Egg deposition in bluegills was reduced after 46 days in 1.4 ng/L
H2S. White sucker eggs were hatched at 15 (ig/L, but juveniles showed
growth reductions at 1 ng/L. Safe level for fathead minnows were between
2 and 3 M-g/L. Studies showed that safe levels for Gammarus Pseudolimnaeus
and Hexagenia limbata were 2 and 15 ng/L, respectively. Some species typi-
cal of normally stressed habitats (Asellus spp.) were much more resistant.
Sulfide criteria for domestic or livestock use have not been established
because the unpleasant odor and taste would preclude such use at hazard-
ous concentrations.
The hazard from hydrogen sulfide to aquatic life is often localized and
transient. Available data indicate that water containing concentrations of
2.0 |xg/L undissociated H2S would not be hazardous to most fish and other
aquatic wildlife, but concentrations in excess of 2.0 ng/L would constitute
a long-term hazard.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
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TAINTING SUBSTANCES
CRITERIA
Aquatic Life Materials should not be present in concentrations that individually or in
combination produce undesirable flavors that are detectable by organolep-
tic tests performed on the edible portions of aquatic organisms.
Rationale	Fish or shellfish with abnormal flavors, colors, tastes, or odors are either
not marketable or elicit consumer complaints and possible rejection of the
food source, even though subsequent lots of organisms may be acceptable.
In some areas, poor product quality can and has seriously affected or elim-
inated the commercial fishing industry. Recreational fishing also can be
affected adversely by off-flavor fish. For the majority of sport fishers, con-
suming the catch is part of their recreation; off-flavored catches can divert
the angler to other waterbodies. This can have serious economic impact on
established recreation industries such as tackle and bait sales and boat and
cottage rental.
A number of wastewaters and chemical compounds have been found
to lower the palatability of fish flesh. Implicated wastewaters included
those from 2,4-D manufacturing plants, kraft and neutral sulfite pulping
processes, municipal wastewater treatment plants, and slaughterhouses, as
well as oily, refinery, and phenolic wastes. The list of implicated chemical
compounds is long; it includes cresol and phenol compounds, kerosene,
naphthol, styrene, toluene, and exhaust outboard motor fuel.
The susceptibility of fishes to the accumulation of tainting substances
is variable and depends on the species, length of exposure, and the pollu-
tant. As little as 0.1 fig/L o-chlorophenol can cause tainting of fish flesh. As
little as 5 ^g/L of gasoline can impart off-flavors to fish.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
209

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210

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TEMPERATURE
CRITERIA 	
Freshwater
Aquatic Life For any time of year, a location has two upper limiting temperatures
(based on the important sensitive species present at that time).
1.	One limit, a maximum temperature for short exposures, is time
dependent and given by the following species-specific equation:
Temperature (°C) = (1 /b) [Logio [time (min)] -a) - 2°C
where:
Logio = logarithm to base 10 (common logarithm)
a = intercept on the " y" or logarithmic axis of the line fitted to
experimental data, which is available for some species from
Appendix II-C of the National Academy of Sciences (1974)
b = slope of the line fitted to experimental data, available for
some species from Appendix II-C of the National
Academy of Sciences (1974)
2.	The second value is a limit on the weekly average temperature that
a.	in cooler months — mid-October to mid-April in the North and
December to February in the South — will protect against
mortality of important species if the elevated plume temperature
is suddenly dropped to the ambient temperature, with the limit
being the acclimation temperature minus 2°C, when the lower
lethal threshold temperature equals the ambient water
temperature (in some regions this limitation may also be
applicable in summer);
b.	in the warmer months — April through October in the North and
March through November in the South — is determined by
adding to the physiological optimum temperature (usually for
growth) a factor calculated as one-third of the difference between
the ultimate upper incipient lethal temperature and the optimum
temperature for the most sensitive important species (and
appropriate life state) that normally is found at that location and
time;
c.	during reproductive seasons — generally April through June and
September through October in the North and March through
May and October through November in the South — the limit is a
temperature that meets site-specific requirements for successful
migration, spawning, egg incubation, fry rearing, and other
reproductive functions of important species. These local
requirements should supersede all other requirements when they
are applicable; or
d.	is site-specific and found necessary to preserve normal species
diversity or prevent appearance of nuisance organisms.
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Marine
Aquatic Life To assure protection of the characteristic indigenous marine community of
a waterbody segment from adverse thermal effects
1.	The maximum acceptable increase in the weekly average
temperature resulting from artificial sources is 10°C (1.8°F) during
all seasons of the year, providing the summer maxima are not
exceeded; and
2.	Daily temperature cycles characteristic of the waterbody segment
should not be altered in either amplitude or frequency.
Summer thermal maxima, which define the upper thermal limits for the
communities of the discharge area, should be established on a site-specific
basis. Existing studies suggest the following regional limits as shown in
Table 1.
Table 1.—Regional Limits.

SHORT-TERM
MAXIMUM
MAXIMUM
TRUE DAILY
MEAN*
Sub-tropical regions (south of Cape Canaveral
and Tampa Bay. Fla., and Hawaii)
Cape Hatteras. N.C.. to Cape Canaveral. Fla.
Long island (south shore) to Cape Hatteras,
N.C.
32.2°C (90°F)
32.2°C (90°F)
30.6°C (87°F)
29.4°C (85°F)
29.4°C (85°F)
27.8°C (82°F)
~True daily mean = average or 24 hourly temperature readings.
Baseline thermal conditions should be measured at a site without un-
natural thermal addition from any source, which is in reasonable
proximity to the thermal discharge (within 5 miles), and which has similar
hydrography to that of the receiving waters at the discharge.
Introduction Human uses of water in and out of its natural situs in the environment are
affected by its temperature. Offstream domestic uses and in-stream recre-
ation are both partially temperature-dependent. Likewise, species
composition and activity of life in any aquatic environment is regulated by
water temperature. Since essentially all of these are so-called "cold
blooded" or poikilotherm organisms, the temperature of the water regu-
lates their metabolism and ability to survive and reproduce effectively.
Industrial uses for process water and cooling are likewise regulated by the
water temperature.
Temperature, therefore, is an important physical parameter that, to
some extent, regulates many of the beneficial uses of water. In 1967, the
Federal Water Pollution Control Administration called temperature "a cat-
alyst, a depressant, an activator, a restrictor, a stimulator, a controller, a
killer — one of the most important and most influential water quality char-
acteristics to life in water."
Rationale	The suitability of water for total body immersion is greatly affected by tem-
perature. In temperate climates, dangers from exposure to low
temperatures is more prevalent than exposure to elevated water tempera-
tures. Depending on the amount of activity by the swimmer, comfortable
temperatures range from 20°C to 30°C. Short durations of lower and higher
temperatures can be tolerated by most individuals. For example, for a 30-
minute period, temperatures of 10°C or 35°C can be tolerated without harm.
212

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Temperature also affects the self-purification phenomenon in
waterbodies, and therefore, the aesthetic and sanitary qualities that exist.
Increased temperatures accelerate the biodegradation of organic material
both in the overlying water and in bottom deposits, which makes in-
creased demands on the dissolved oxygen resources of a given system. The
typical situation is exacerbated by the fact that oxygen becomes less solu-
ble as water temperature increases. Thus, greater demands are exerted on
an increasingly scarce resource that may lead to total oxygen depletion and
obnoxious septic conditions.
Indicator enteric bacteria, and presumably enteric pathogens, are like-
wise affected by temperature. Both total and fecal coliform bacteria die
away more rapidly in the environment with increasing temperatures. Like-
wise, changes from a coldwater fishery to a warmwater fishery can occur
because temperature may be directly lethal to adults or fry and cause a re-
duction of activity or limit reproduction.
Upper and lower limits for temperature have been established for
many aquatic organisms. Considerably more data exist for upper than
lower limits. Tabulations of lethal temperatures for fish and other organ-
isms are available. Factors such as diet, activity, age, general health,
osmotic stress, and even weather contribute to the lethality of temperature.
The aquatic species, thermal accumulation state, and exposure time are
considered the critical factors.
The effects of sublethal temperatures on metabolism, respiration, be-
havior, distribution and migration, feeding rate, growth, and reproduction
have been summarized. Another study has illustrated that the tolerance
zone contains a more restrictive temperature range in which normal activ-
ity and growth occur; and an even more restrictive zone exists inside that
in which normal reproduction will occur.
Data on the combined effects of increased temperature and toxic mate-
rials on fish indicate that toxicity generally increases with increased
temperature an that organisms subjected to stress from toxic materials are
less tolerant of temperature extremes.
An organisms tolerance to temperature extreme is a function of its ge-
netic ability to adapt to thermal changes.
Temperature effects have been shown for water treatment processes.
Lower temperatures reduce the effectiveness of coagulation with alum and
subsequent rapid sand filtration. In one study, difficulty was especially
pronounced below 5°C. Decreased temperature also decreases the effec-
tiveness of chlorination. Based on studies relating chlorine dosage to
temperature, and with a 30-minute contact time, dosages required for
equivalent disinfective effect increased by as much as a factor of 3 when
temperatures were decreased from 20°C to 10°C. Increased temperature
may increase the water's odor because of the increased volatility of odor-
causing compounds. Odor problems associated with plankton may also be
aggravated.
The effects of temperature on aquatic organisms have been the subject
of comprehensive literature reviews and annual literature reviews pub-
lished by the Water Pollution Control Federation. Only highlights from the
thermal effects on aquatic life are presented here.
Temperature changes in waterbodies can alter the existing aquatic
community. The dominance of various phytoplankton groups in specific
temperature ranges has been shown. For example, from 20°C to 25°C, dia-
toms predominated; green algae predominated from 30°C to 35"C; and
blue-greens predominated above 35°C within their characteristic
213

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temperature range, the acclimation temperature prior to exposure, and the
time of exposure to the elevated temperature. The upper incipient lethal
temperature or the highest temperature than 50 percent of a sample of or-
ganisms can survive is determined on the organism at the highest
sustainable acclimation temperature. The lowest temperature that 50 per-
cent of the warm acclimated organisms can survive in is the ultimate lower
incipient lethal temperature. True acclimation to changing temperatures
requires several days. The lower end of the temperature accommodation
range for aquatic life is 0°C in fresh water and somewhat less for saline wa-
ters. However, organisms acclimated to relatively warm water, when
subjected to reduced temperatures that under other conditions of acclima-
tion would not be detrimental, may suffer a significant mortality caused by
thermal shock.
Through the natural changes in climatic conditions, the temperatures
of waterbodies fluctuate daily, as well as seasonally. These changes do not
eliminate indigenous aquatic populations, but affect the existing commu-
nity structure and the geographic distribution of species. Such temperature
changes are necessary to induce the reproductive cycles of aquatic organ-
isms and to regulate other life factors.
Artificially induced changes, such as the return of cooling water or the
release of cool hypolimnetic waters from impoundments, may alter indige-
nous aquatic ecosystems. Entrained organisms may be damaged by
temperature increases across cooling water condensers if the increase is
sufficiently great or the exposure period sufficiently long. Impingement
upon condenser screens, chlorination for slime control, or other physical
insults damage aquatic life. However, data has shown that algae passing
through condensers are not injured if the temperature of the outflowing
water does not exceed 345°C.
In open waters elevated temperatures may affect periphyton, benthic
invertebrates, and fish, in addition to causing shifts in algae dominance.
Studies of the Delaware River downstream from a power plant concluded
that the periphyton population was considerably altered by the discharge.
The number and distribution of bottom organisms decrease as water
temperature increase. The upper tolerance limit for a balanced benthic
population structure is approximately 32°C. A large number of these inver-
tebrate species are able to tolerate higher temperatures than those required
for reproduction.
In order to define criteria for fresh waters, the following was cited as
currently defineable requirements:
1.	Maximum sustained temperatures that are consistent with
maintaining desirable levels of productivity;
2.	Maximum levels of metabolic acclimation to warm temperatures
that will permit return to ambient winter temperatures should
artificial sources of heat cease;
3.	Time-dependent temperature limitations for survival of brief
exposures to temperature extremes, both upper and lower;
4.	Restricted temperature ranges for various states of reproduction,
including (for fish) gametogenesis, spawning migration, release of
gametes, development of the embryo, commencement of
independent feeding (and other activities) by juveniles, and
temperatures required for metamorphosis, emergence, or other
activities of lower forms;

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5.	Thermal limits for diverse species compositions of aquatic
communities, particularly where reduction in diversity creates
nuisance growths of certain organisms, or where important food
sources (food chains) are altered;
6.	Thermal requirements of downstream aquatic life (in rivers) where
upstream diminution of a coldwater resource will adversely affect
downstream temperature requirements.
The major portion of such information that is available, however, is for
freshwater fish species rather than lower forms of marine aquatic life.
The temperature-time duration for short-time exposures, such that 50
percent of a given population will survive and extreme temperature, fre-
quently is expressed mathematically by fitting experimental data with a
straight line on a semi-logarithmic plot with time on the logarithmic scale
and temperature on the linear scale. In equation form, this 50 percent mor-
tality relationship is
Logio (time^minutes^) = a + b (Temperature (°C))
where:
Logio = logarithm to base 10 (common logarithm)
a = intercept on the "y" or logarithmic axis of the line fitted to
experimental data and which is available for some species
from Appendix II-C of the National Academy of Sciences
document
b = slope of the line fitted to experimental data and which is
available for some species from Appendix Il-C of the
National Academy of Sciences document
To provide a safety factor so that none or only a few organisms will
perish, experiments found that a criterion of 2°C below maximum temper-
ature is usually sufficient. To provide safety for all the organisms, the
temperature causing a median mortality for 50 percent of the population
would be calculated and reduced by 2°C in the case of an elevated temper-
ature. Available scientific information includes upper and lower incipient
lethal temperatures, coefficients "a" and "b" for the thermal resistance
equation, and information of size, life stage, and geographic source of the
particular test species.
Maximum temperatures for an extensive exposure (e.g., more than one
week) must be divided into those for warmer periods and winter. Other
than for reproduction, the most temperature-sensitive life function appears
to be growth. A satisfactory estimate of a limiting maximum weekly mean
temperature may be an average of the optimum temperature for growth
and the temperature for zero net growth.
Because of the difficulty in determining the temperature of zero net
growth, essentially the same temperature can be derived by adding to the
optimum essentially to temperature (for growth or other physiological
functions) a factor calculated as one-third of the difference between the ul-
timate upper incipient lethal temperature and the optimum temperature.
In equation form
Maximum weekly = optimum + Vs (ultimate upper incipient lethal
average temperature temperature	temperature - optimum
temperature)
Since temperature tolerance varies with various states of development
of a particular species, the criterion for a particular location would be
215

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calculated for the most important life form likely to be present during a
particular month. One caveat in using the maximum weekly mean temper-
ature is that the limit for short-term exposure must not be exceeded.
Example calculations for predicting the summer maximum temperatures
for short-term survival and for extensive exposure for various fish species
are presented in Table 2. These calculations use the above equations and
data from EPA's Environmental Research Laboratory in Duluth.
Table 2.— Example calculated values for maximum weekly average temperatures for
growth and short-term maxima for survival for juveniles and adults during the
summer (centigrade and Fahrenheit).
SPECIES
GROWTH'
MAXIMA"
Atlantic salmon
20 (68)
23 (73)
Bigmouth buffalo
—
—
Black crappie
27(81)
—
Bluegill
32 (90)
35 (95)
Brook trout
19 (66)
24 (75)
Carp
—
—
Channel catfish
32 (90)
35 (95)
Coho salmon
18 (64)
24 (75)
Emerald shiner
30 (86)
—
Freshwater drum
—
—
Lake herring (Cisco)
17 (63)c
25 (77)
Largemouth bass
32 (90)
34 (93)
Northern pike
28 (82)
30 (86)
Rainbow trout
19 (66)
24 (75)
Sauger
25 (77)
—
Smalimouth bass
29 (84)
—
Smallmouth buffalo
—
—
Sockcye salmon
18 (64)
22 (72)
Striped bass
—
—
Thrcadfin shad
—
—
White bass
—
—
White crappie
28 (82)
—
While sucker
28 (82)c
—
Yellow perch
29 (84)
—
•Calculated according to (he equation (using optimum temperature for gnmth) maximum weekly average temperature
for growth = optimum temperature + lA (ultimate incipient lethal temperature — optimum temperature).
''Based on temperature (°C) "= l/b |togur
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Table 3.— Summary of reported values for maximum weekly average temperatures
for spawning and short-term maxima for embryo survival during the spawning
season (centigrade and Fahrenheit).
SPECIES
SPAWNING"
EMBRYO
SURVIVAL"
Atlantic salmon
5 (41)
11 (52)
Biemouth buffalo
17 (63)
27 (81 )c
Black crappie
17 (63)
20 (68)c
Blucgill
25 (77)
34 (93)
Brook trout
9 (48)
13 (55)
Carp
21 (70)
33 (91)
Channel catfish
27 (81)
29 (84)c
Coho salmon
10 (50)
13 (55)<
Emerald shiner
24 (75)
28 (82)c
Freshwater drum
21 (70)
26 (79)
Lake herrinc (Cisco)
3 (37)
8 (46)
Largcmouth bass ,.
21 (70)
27 (81)c
Northern pike
11 (52)
19 (66)
Rainbow trout
8 (46)
15 (59)
Saugcr
12 (54)
18 (64)
Smallmouth bass
17 (63)
23 (73)c
Smallmouth buffalo
21 (70)
28 (82)c
Sockeye salmon
10 (50)
13 (55)
Striped bass
18 (64)
24 (75)
Threadfin shad
18 (64)
34 (93)
White bass
17 (63)
26 (79)
White crappic
18 (64)
23 (73)
White sucker
10 (50)
20 (68)
Yellow perch
12 (54)
20 (68)
•The oplimum or mean of ihe range of spawning temperatures reported for the species (CRL-Duluih. 1976).
*The upper temperature for successful incubation and hatching reported for the species (ERl-Duluth. 1976).
cUpper temperature for spawning.
unproductive area. Reported temperature data on maximum temperatures
for spawning and embryo survival have been summarized in Table 3.
Although the limiting effects of thermal addition to estuarine and ma-
rine waters are not as conspicuous in the fall, winter, and spring as during
the summer season of maximum heat stress, nonetheless crucial thermal
limitations do exist. Hence, thermal additions to the receiving waters must
be minimized during all seasons. Size of harvestable stocks of commercial
fish and shellfish, particularly near geographic limits of the fishery, appear
to be markedly influenced by slight changes in the long-term temperature
regime.
Studies on the relationship between temperature and annual variation
in seven-year catch data for winter flounder (Pseudopleuronected amer-
icanus) in Narragansett Bay, Rhode Island, revealed that a 78 percent
decrease in annual catch correlated closely with a 0.5°C increase in the av-
erage temperature over the 30-month period between spawning and
recruitment in to the fishery. Sissenwine's 1874 model predicts a 68 percent
reduction of recruitment in yellowtail flounder (Limanda ferruginea) with a
1°C long-term elevation in southern New England waters.
Community balance can be influenced strongly by such temperature-
dependent factors as rates of reproduction, recruitment, and growth of
each component population. A few degrees elevation in average monthly
temperature can appreciably alter a community through changes in inter-
species relationships. A 50 percent reduction in the softshell clam fishery in
Maine by the green crab (Carcinus maenus) illustrates how an increase in
winter temperatures can establish new predator-prey relationships. Over a
period of four years, temperature naturally ameliorated and the monthly
mean for the coldest month of each year did not fall below 2°C. This appar-
ently precluded appreciable ice formation and winter cold kill of the green
crab and permitted a major expansion of its population, with resulting in-
creased predation of the softshell clam.
217
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Temperature is a primary factor controlling reproduction and can in-
fluence many events of the reproductive cycle from gametogenesis to
spawning. Among marine invertebrates, initiation of reproduction (game-
togenesis) is often triggered during late winter by attainment of a
minimum environmental threshold temperature. In some species, avail-
ability of adequate food is also a requisite. Elevated temperature can limit
gametogenesis by preventing accumulation of nutrients in the gonads.
This problem could be acute during the winter if food availability and
feeding activity is reduced. Most marine organisms spawn during the
spring and summer; gametogenesis is usually initiated during the previ-
ous fall. Some species spawn only during the fall (herring), while others
spawn during the winter and early spring. At the higher latitudes, winter
breeders include such estuarine community dominants as acorn barnacles
(Balanus balanus and B. balanoides), the edible blue mussel (Mytilus edulis),
sea urchin Strongylocentrotus drobachiensis), sculpin, and the winter floun-
der (Pseudopleuronectes americanus). The two boreal barnacles require
temperatures below 10°C before egg production will be initiated. Adapta-
tions for reproduction exist that are dependent on temperature condition
closes to the natural cycle.
Juvenile and adult fish usually thermoregulate behaviorally by moving
to water having temperatures closest to their thermal preference. This pro-
vides a thermal environment that approximates the optimal temperature
for many physiological functions, including growth. As a consequence,
fishes usually are attracted to heated water during the fall, winter, and
spring. Avoidance will occur as warmer temperature exceeds the pre-
ferendum by 1 to 3°C. This response precludes problems of heat stress for
juvenile and adult fishes during the summer, but several potential prob-
lems exist during the other seasons. The possibility of cold shock and
death of plume-entrained fish resulting from winter plant shutdown is
well recognized. Also, increased incidence of disease and a deterioration of
physiological condition has been observed among plume-entrained fishes,
perhaps because of insufficient food. A weight loss of approximately 10
percent for each 1°C rise in water temperature has been observed in fish
when food is absent. Indirect adverse effects may also occur on the indige-
nous community because of increased predation pressure if thermal
addition leads to a concentration of fish that are dependent on this com-
munity for their food.
Fish migration is often linked to natural environmental temperature
cycles. In early spring, fish employ temperature as their environmental cue
to migrate northward (e.g., menhaden, bluefish) or to move inshore (win-
ter flounder). Likewise, water temperature strongly influences timing of
spawning runs of anadromous fish into rivers. In the autumn, a number of
juvenile marine fishes and shrimp are dependent on a drop in temperature
to trigger their migration from estuarine nursery grounds for oceanic dis-
persal or southward migration.
Thermal discharges should not alter diurnal and tidal temperature
variations normally experienced by marine communities. Laboratory stud-
ies show thermal tolerance to be enhanced when animals are maintained
under a diurnally fluctuating temperature regime rather than at a constant
temperature. A daily cyclic regime can be protective additionally as it re-
duces duration of exposure to extreme temperatures.
Summer thermal maxima should be established to protect the various
marine communities within each biogeographic region. During the sum-
mer, naturally elevated temperatures may be of sufficient magnitude to
cause death or emigration. This more commonly occurs in tropical and
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warm temperate zone waters, but has been reported for enclosed bays and
shallow waters in other regions as well. Summer heat stress also can con-
tribute to increased incidence of disease or parasitism; reduced or blocked
sexual maturation; inhibited or blocked embryonic cleavage of larval de-
velopment; reduced feeding and growth of juveniles and adults; increased
predation; and reduced productivity of macroalgae and seagrasses. The
general ceilings set forth here are derived from studies delineating limiting
temperatures for the more thermally sensitive species or communities of a
biogeographic region.
(Quality Criteria for Water, July 1976) PB-263943
See Appendix D for Methodology.
219
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220
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2,3,7,8-TETRACHLORODIBENZO- P-DIOXIN
(TCDD) (DIOXIN)
CRITERIA
Aquatic Life Not enough data are available concerning the effects of 2,3,7,8-TCDD on
aquatic life and its uses to derive national criteria. The available informa-
tion indicates that acute values for some freshwater animal species are 1.0
Hg/L; some chronic values are M-g/L; and the chronic value for rain-
bow trout (Oncorhynchus mykiss) is 0.001 ng/L. Because exposures of some
species of fishes to 0.01 M-g/L for < 6 days resulted in substantial mortality
several weeks later, derivation of aquatic life criteria for 2,3,7,8-TCDD may
require special consideration.
Predicted bioconcentration factors (BCFs) for 2,3,7,8-TCDD range from
3,000 to 900,000, but the available measured BCFs range from 390 to 13,000.
If the BCF is 5,000, concentrations 0.00001 ng/L should result in concentra-
tions in edible freshwater and saltwater fish and shellfish that exceed
levels identified in a Food and Drug Administration health advisory. If the
BCF is 5,000 or if uptake in a field situation is greater than in laboratory
tests, the value of 0.00001 M-g/L will be too high.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of 2,3,7,8-TCDD exposure through ingestion of contaminated
water and contaminated aquatic organisms, the ambient water concentra-
tion should be zero. This criterion is based on the nonthreshold
assumption for 2,3,7,8-TCDD. However, zero may not be an attainable
level at this time. Therefore, the levels that may result in an increase of can-
cer risk over the lifetime are estimated at 10"5, 10"6, and 10"7, and the
corresponding recommended criteria are 1.3 x lO'^g/L, 1.3 x 10"8 ng/L,
and 1.3 x 10"9 ng/L, respectively.
If the above estimates are made for consumption of aquatic organisms
only, excluding consumption of water, the levels are 1.4 x 10"7 ng/L, 1.4 x
10 ng/L, and 1.4 x 10"9 jig/L, respectively. If these estimates are made for
comsumption of water only, the levels are 2.2 x 10"6 ng/L, 2.2 x 10"7ng/L,
and 2.2 x 10"8 ng/L, respectively.
(49 F.R. 5831, February 15,1984)
See Appendix A for Aquatic Life Methodology.
221
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222
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TETRACHLOROETHYLENE
127-18-4
CRITERIA
Aquatic Life For freshwater aquatic life, the available data for tetrachloroethylene indi-
cate that acute and chronic toxicity occurs at concentrations as low as 5,280
and 840 ng/L, respectively, and would occur at lower concentrations
among species that are more sensitive than those tested.
For saltwater aquatic life, the available data for tetrachloroethylene in-
dicate that acute and chronic toxicity occurs at concentrations as low as
10,200 and 450 ng/L, respectively, and would occur at lower concentra-
tions among species that are more sensitive than those tested.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to tetrachloroethylene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable.
Therefore, the levels that may result in incremental increase of cancer risk
over the lifetime are estimated at 10-5, 10"6, and 10"7. The corresponding
recommended criteria are 8.0 ng/L, 0.80 M-g/L, and 0.08 ng/L, respectively.
If these estimates are made for consumption of aquatic organisms only, ex-
cluding consumption of water, the levels are 88.5 ng/L, 8.85 fig/L, and 0.88
M-g/L, respectively.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
223
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224
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THALLIUM
7440-28-0
CRITERIA
Aquatic Life For freshwater aquatic life, the available data for thallium indicate that
acute and chronic toxicity occurs at concentrations as low as 1,400 and 40
Hg/L, respectively, and would occur at lower concentrations among spe-
cies that are more sensitive than those tested. Toxicity to one species of fish
occurs at concentrations as low as 20 ng/L after 2,600 hours of exposure.
For saltwater aquatic life, the available data for thallium indicate that
acute toxicity occurs at concentrations as low as 2,130 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of thal-
lium to sensitive saltwater aquatic life.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for thallium are 1.7 ng/L for ingestion of contaminated
water and organisms and 6.3 ng/L for ingestion of contaminated aquatic
organisms only.
(45 F.R. 79318, November 28,1980) (57 F.R. 60848, December 22,1992)
See Appendix C for Human Health Methodology.
225
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226
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TOLUENE
108-88-3
CRITERIA
Aquatic Life The available data for toluene indicate that acute toxicity to freshwater
aquatic life occurs at concentrations as low as 17,500 ng/L and would
occur at lower concentrations among species that are more sensitive than
those tested. No data are available concerning the chronic toxicity of tolu-
ene to sensitive freshwater aquatic life.
The available data for toluene indicate that acute and chronic toxicity
to saltwater aquatic life occurs at concentrations as low as 6,300 and 5,000
Hg/L, respectively, and would occur at lower concentrations among spe-
cies that are more sensitive than those tested.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for toluene are 6,800 p.g/L for ingestion of contaminated
water and organisms and 200,000 fig/L for ingestion of contaminated or-
ganisms only.
(45 F.R. 79318, November 28, 1980) (57 F.R. 60848, December 22, 1992)
See Appendix C for Human Health Methodology.
227
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228
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TOXAPHENE
8001-35-2
CRITERIA
Aquatic Life Freshwater — 4-day average of 0.0002 ng/L
1-hour average of 0.73 ng/L
Saltwater — 4-day average of 0.0002 jig/L
1-hour average of 0.21 ng/L
Summary The acute sensitivities of 36 freshwater species in 28 genera ranged from
0.8 ng/L to 500 jig/L. Such important fish species as the channel catfish,
largemouth bass, chinook and coho salmon, brook, brown, and rainbow
trout, striped bass, and bluegill had acute sensitivities ranging from 0.8
Hg/L to 10.8 ug/L. Chronic values for four freshwater species range from
less than 0.039 jxg/L for the brook trout (Salvelinus fontitialis) to 0.1964
Ug/L for the channel catfish (Ictalurus punctatus).
The growth of algae was affected at 100 to 1,000 ug/L, and
bioconcentration factors from laboratory tests ranged from 3,100 to 90,000.
Concentrations in lake trout in the Great Lakes have frequently exceeded
the Food and Drug Administration (FDA) action level of 5 mg/kg, even
though the concentrations in the water seem to be only 1 to 4 ng/L. These
concentrations in lake water are thought to have resulted from toxaphene
being transported to the Great Lakes from remote sites, the locations of
which are not well known.
The acute toxicity of toxaphene to 15 species of saltwater animals
ranges from 0.53 for pinfish (Lagodon rhomoides) to 460,000 ng/L for the
adults of the clam (Rangia cuneata). Except for resistant species tested at
concentrations greater than toxaphene's water solubility, acute values for
most species were within a factor of 10. The toxicity of toxaphene was
found to decrease slightly with increasing salinity for adult blue crabs
(Callinectes sapidus), whereas no relationship between toxicity and salinity
was observed with the threespine stickleback (Gasterosteus aculeatus). Tem-
perature significantly affected the toxicity of toxaphene to blue crabs.
Early life-stage toxicity tests have been conducted with the sheepshead
minnow (Cyprinodon variegatus) and the longnose killifish (Fundulus
similis), whereas life-cycle tests have been conducted with the sheepshead
minnow and a mysid. For the sheepshead minnow, chronic values of 1.658
fig/L from the early life-stage test and 0.7141 ng/L from the life-cycle tox-
icity test are similar to the 96-hour LC50 of 1.1 ng/L. Killifish are more
chronically sensitive, with effects noted at 0.3 ng/L. In the life-cycle test
with the mysid, no adverse effects were observed at the highest concentra-
tion tested, which was only slightly below the 96-hour LC50, resulting in
an acute-chronic ratio of 1.132.
Toxaphene is bioconcentrated by an oyster, Crassostrea virginica, and
two fishes, C. variegatus and F. similis, to concentrations that range from
9,380 to 70,140 times that in the test solution.
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National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
not be affected unacceptably if the four-day average concentration of toxa-
phene does not exceed 0.0002 ng/L more than once every three years on
the average and if the one-hour average concentration does not exceed 0.73
Hg/L more than once every three years on the average. If the concentration
of toxaphene does exceed 0.0002 ng/L, the edible portions of consumed
species should be analyzed to determine whether the concentration of tox-
aphene exceeds the FDA action level of 5 mg/kg. If the channel catfish is as
acutely sensitive as some data indicate it might be, it will not be protected
by this criterion.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of toxaphene does not exceed 0.0002 (xg/L more
than once every three years on the average and if the one-hour average
concentration does not exceed 0.21 jxg/L more than once every three years
on the average. If the concentration of toxaphene does exceed 0.0002 ng/L,
the edible portions of consumed species should be analyzed to determine
whether the concentration of toxaphene exceeds the FDA action level of 5
mg/kg.
In the EPA's best scientific judgment, three years is the average amount
of time aquatic ecosystems should be provided between excursions. The
resiliences of ecosystems and their abilities to recover differ greatly, how-
ever, and site-specific allowed excursion frequencies may be established if
adequate justification is provided.
Criteria for developing water quality-based permit limits and for
designing waste treatment facilities must be applied to an appropriate
wasteload allocation model; dynamic models are preferred. Limited data
or other considerations might make their use impractical; in which case,
rely on a steady-state model.
Human Health Human health criteria were recalculated using Integrated Risk Information
System (IRIS) to reflect available data as of 12/92 (57 F.R. 60848). Recalcul-
ated IRIS values for toxaphene are 0.00073 ng/L for ingestion of
contaminated water and organisms and 0.00075 fig/L for ingestion of con-
taminated organisms only.
(51 F.R. 43665, December 3,1986) (57 F.R. 60848, December 22,1992)
See Appendix A for Aquatic Life Methodology.
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TRICHLOROETHYLENE
79-01-6
CRITERIA
Aquatic Life The available data for trichloroethylene indicate that acute toxicity to
freshwater aquatic life occurs at concentrations as low as 45,000 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
trichloroethylene to sensitive freshwater aquatic life, but the behavior of
one species is adversely affected at concentrations as low as 21,900 ^g/L.
The available data for trichloroethylene indicate that acute toxicity to
saltwater aquatic life occurs at concentrations as low as 2,000 ng/L and
would occur at lower concentrations among species that are more sensitive
than those tested. No data are available concerning the chronic toxicity of
trichloroethylene to sensitive saltwater aquatic life.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to trichloroethylene through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentration should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10"5,10"6, and 10"7. The cor-
responding recommended criteria are 27 ng/L, 2.7 ng/L, and 0.27 ng/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 807 fig/L, 80.7
Hg/L, and 8.07 ng/L, respectively.
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
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232
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VINYL CHLORIDE
75-01-4
CRITERIA
Aquatic Life No freshwater or saltwater organisms have been tested with vinyl chlo-
ride; therefore, no statement can be made concerning acute or chronic
toxicity.
Human Health For the maximum protection of human health from the potential carcino-
genic effects of exposure to vinyl chloride through ingestion of
contaminated water and contaminated aquatic organisms, the ambient
water concentrations should be zero, based on the nonthreshold assump-
tion for this chemical. However, zero level may not be attainable at the
present time. Therefore, the levels that may result in incremental increase
of cancer risk over the lifetime are estimated at 10~5,10"6, and 10"7. The cor-
responding recommended criteria are 20 ng/L, 2.0 ng/L, and 0.2 ng/L,
respectively. If these estimates are made for consumption of aquatic organ-
isms only, excluding consumption of water, the levels are 5,246 ng/L, 525
Hg/L, and 52.5 ng/L, respectively:
(45 F.R. 79318, November 28,1980)
See Appendix C for Human Health Methodology.
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234
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ZINC
7440-66-6
CRITERIA
Aquatic Life Saltwater — 4-day average of 86 ng/L
1-hour average of 95 fig/L
Freshwater criteria are hardness dependent. See text.
Summary Acute toxicity values available for 43 species of freshwater animals and
data for eight species indicate that acute toxicity decreases as hardness in-
creases. When adjusted to a hardness of 50 mg/L, sensitivities range from
50.70 ng/L for Ceriodaphnia reticulata to 88,960 ng/L for a damselfly. Addi-
tional data indicate that toxicity increases as temperature increases.
Chronic toxicity data are available for nine freshwater species. Chronic
values for two invertebrates ranged from 46.73 ng/L for Daphtiia magna to
5,243 ng/L for the caddisfly (Clistoronia magnificia). Chronic values for
seven fish species ranged from 36.41 ng/L for the flagfish (Jordanella
floridae) to 854.7 M-g/L for the brook trout (Salvelinus fontinalis). Acute-
chronic ratios ranged from 0.2614 to 41.20, but the ratios for the sensitive
species were all less than 7.3.
The sensitivity range of freshwater plants to zinc is greater than that
for animals. Growth of the alga (Selenastrum capriocornutum) was inhibited
by 30 ng/L. On the other hand, with several other species of green algae,
four-day EC50s exceeded 200,000 ng/L. Zinc was found to bioaccumulate
in freshwater animal tissues from 51 to 1,130 times the concentration pres-
ent in the water.
Acceptable acute toxicity values for zinc are available for 33 species of
saltwater animals including 26 invertebrates and 7 fish. LC50s range from
191.5 ng/L for cabezon (Scorpanichthys marmoratus) to 320,400 ng/L for
adults of another clam, Macoma balthica. Early life stages of saltwater inver-
tebrates and fishes are generally more sensitive to zinc than juveniles and
adults. Temperature has variable and inconsistent effects on the sensitivity
of saltwater invertebrates to zinc. The sensitivity of saltwater vertebrate
animals to zinc decreases with increasing salinity, but the magnitude of the
effect is species-specific.
A life-cycle test with the mysid (Mysidopsia bahia) found unacceptable
effects at 120 fig/L but not at 231 ng/L, and the acute-chronic ratio was
2.997. Seven species of saltwater plants were affected at concentrations
ranging from 19 to 10,100 ng/L. Bioaccumulation data for zinc are avail-
able for seven species of saltwater algae and five species of saltwater
animals. Steady-state zinc bioconcentration factors for the 12 species range
from 3.692 to 23,820.
National Criteria The procedures described in the "Guidelines for Deriving Numerical Na-
tional Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important spe-
cies is very sensitive, freshwater aquatic organisms and their uses should
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not be affected unacceptably if the four-day average concentration of zinc
(in fig/L) does not exceed the numerical value given by
g(0.8473[ln (hardness)]+0.76l4)
more than once every three years on the average and if the one-hour
average concentration (in ng/L) does not exceed the numerical value given
by
e(0.8473[ln (hardness)]+0.8604)
more than once every three years on the average. For example, at
hardnesses of 50, 100, and 200 mg/L as CaCO^ the four-day average
concentrations of zinc are 59, 110 and 190 ng/L, respectively, and the
one-hour average concentrations are 65, 120, and 210 ng/L. If the striped
bass is as sensitive as some data indicate, it will not be protected by this
criterion.
The procedures described in the guidelines indicate that, except possi-
bly where a locally important species is very sensitive, saltwater aquatic
organisms and their uses should not be affected unacceptably if the four-
day average concentration of zinc does not exceed 86 ng/L more than once
every three years on the average and if the one-hour average concentration
does not exceed 95 ng/L more than once every three years on the average.
Criteria for developing water quality-based permit limits and for
designing waste treatment facilities must be applied to an appropriate
wasteload allocation model; dynamic models are preferred. Limited data
or other considerations might make their use impractical, in which case
rely on a steady-state model.
(52 F.R. 6213, March 2,1987)
See Appendix A for Aquatic Life Methodology.
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APPENDIX A
Derivation of the 1985
Aquatic Life Critera
The following is a summary of the Guidelines for Derivation of Criteria for Aquatic Life. The complete text is found in "Guidelines for
Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses," available from National
Technical Information Service - PB85-227049.
Derivation of numerical national water quality criteria for the protection of aquatic organisms and
their uses is a complex process that uses information from many areas of aquatic toxicology. When a
national criterion is needed for a particular material, all available information concerning toxicity to
and bioaccumulation by aquatic organisms is collected, reviewed for acceptability, and sorted. If
enough acceptable data on acute toxicity to aquatic animals are available, they are used to estimate
the highest one-hour average concentration that should not result in unacceptable effects on aquatic
organisms and their uses. If justified, this concentration is made a function of water quality
characteristics such as pH, salinity, or hardness. Similarly, data on the chronic toxicity of the
material to aquatic animals are used to estimate the highest four-day average concentration that
should not cause unacceptable toxicity during a long-term exposure. If appropriate, this
concentration is also related to a water quality characteristic.
Data on toxicity to aquatic plants are examined to determine whether plants are likely to be
unacceptably affected by concentrations that should not cause unacceptable effects on animals.
Data on bioaccumulation by aquatic organisms are used to determine if residues might subject
edible species to restrictions by the U.S. Food and Drug Administration (FDA), or if such residues
might harm wildlife that consumes aquatic life. All other available data are examined for adverse
effects that might be biologically important.
If a thorough review of the pertinent information indicates that enough acceptable data exists,
numerical national water quality criteria are derived for fresh water or salt water or both to protect
aquatic organisms and their uses from unacceptable effects due to exposures to high concentrations
for short periods of time, lower concentrations for longer periods of time, and combinations of the
two.
I. Definition of Material of Concern
A. Each separate chemical that does not ionize substantially in most natural bodies of water
should usually be considered a separate material, except possibly for structurally similar
organic compounds that exist only in large quantities as commercial mixtures of the
various compounds and apparently have similar biological, chemical, physical, and toxi-
cological properties.
B. For chemicals that do ionize substantially in most natural waterbodies (e.g., some phenols
and organic acids, some salts of phenols and organic acids, and most inorganic salts and
coordination complexes of metals), all forms in chemical equilibrium should usually be
considered one material. Each different oxidation state of a metal and each different
non-ionizable covalently bonded organometallic compound should usually be
considered a separate material.
C. The definition of the material should include an operational analytical component.
Identification of a material simply, for example, as "sodium" obviously implies "total
sodium" but leaves room for doubt. If "total" is meant, it should be explicitly stated. Even
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"total" has different operational definitions, some of which do not necessarily measure
"all that is there" in all sample. Thus, it is also necessary to reference or describe one
analytical method that is intended. The operational analytical component should take into
account the analytical and environmental chemistry of the material, the desirability of
using the same analytical method on samples from laboratory tests, ambient water and
aqueous effluents, and various practical considerations such as labor and equipment
requirements and whether the method would require measurement in the field or would
allow measurement after samples are transported to a laboratory.
The primary requirements of the operational analytical component are that it be
appropriate for use on samples of receiving water, compatible with the available toxicity
and bioaccumulation data without making overly hypothetical extrapolations, and rarely
result in underprotection or overprotection of aquatic organisms and their uses. Because
an ideal analytical measurement will rarely be available, a compromise measurement will
usually be used. This compromise measurement must fit with the general approach: if an
ambient concentration is lower than the national criterion, unacceptable effects will
probably not occur (i.e., the compromise measurement must not err on the side of
underprotection when measurements are made on a surface water). Because the chemical
and physical properties of an effluent are usually quite different from those of the
receiving water, an analytical method acceptable for analyzing an effluent might not be
appropriate for analyzing a receiving water, and vice versa. If the ambient concentration
calculated from a measured concentration in an effluent is higher than the national
criterion, an additional option is to measure the concentration after dilution of the effluent
with receiving water to determine if the measured concentration is lowered by such
phenomena as complexation or sorption. A further option, of course, is to derive a
site-specific criterion. Thus, the criterion should be based on an appropriate analytical
measurement, but the criterion is not rendered useless if an ideal measurement either is
not available or is not feasible.
The analytical chemistry of the material might need to be considered when defining
the material or when judging the acceptability of some toxicity tests, but a criterion should
not be based on the sensitivity of an analytical method. When aquatic organisms are more
sensitive than routine analytical methods, the proper solution is to develop better
analytical methods, not to underprotect aquatic life.
II. Collection of Data
A. Collect all available data on the material concerning toxicity to, and bioaccumulation by,
aquatic animals and plants; FDA action levels and chronic feeding studies and long-term
field studies with wildlife species that regularly consume aquatic organisms.
B. All data that are used should be available in typed, dated, and signed hard copy
(publication, manuscript, letter, memorandum) with enough supporting information to
indicate that acceptable test procedures were used and that the results are probably
reliable. In some cases, additional written information from the investigator may be
needed. Information that is confidential, privileged, or otherwise not available for
distribution should not be used.
C. Questionable data, whether published or unpublished, should not be used. Examples
would be data from tests that did not contain a control treatment, tests in which too many
organisms in the control treatment died or showed signs of stress or disease, and tests in
which distilled or deionized water was used as the dilution water without addition of
appropriate salts.
D. Data on technical grade materials may be used, if appropriate; but data on formulated
mixtures and emulsifiable concentrates of the material may not be used.
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E. For some highly volatile, hydrolyzable, or degradable materials, only use data from
flow-through tests in which the concentrations of test material were measured often
enough with acceptable analytical methods.
F. Data should be rejected if obtained by using:
• Brine shrimp — because they usually occur naturally only in water with salinity
greater than 35 g / kg;
• Species that do not have reproducing wild populations in North America; or
• Organisms that were previously exposed to substantial concentrations of the test
material or other contaminants.
G. Questionable data, data on formulated mixtures and emulsifiable concentrates, and data
obtained with nonresident species or previously exposed organisms may be used to
provide auxiliary information but should not be used in the derivation of criteria.
Required Data
A. Certain data should be available to help ensure that each of the four major kinds of
possible adverse effects receives adequate consideration: results of acute and chronic
toxicity tests with representative species of aquatic animals are necessary to indicate the
sensitivities of appropriate untested species. However, since procedures for conducting
tests with aquatic plants and interpreting the results are not as well developed, fewer data
concerning toxicity are required. Finally, data concerning bioaccumulation by aquatic
organisms are required only when relevant information on the significance of residues in
aquatic organisms is available.
B. To derive a criterion for freshwater aquatic organisms and their uses, the following should
be available:
1. Results of acceptable acute tests (see section IV) with at least one species of freshwater
animal in at least eight different families including all of the following:
• The family Salmonidae in the class Osteichthyes.
• A second family in the class Osteichthyes, preferably a commercially or
recreationally important warm water species, such as bluegill or channel catfish.
• A third family in the phylum Chordata (may be in the class Osteichthyes or may
be an amphibian, etc.).
• A planktonic crustacean such as a cladoceran or copepod.
• A benthic crustacean (ostracod, isopod, amphipod, crayfish, etc.).
• An insect (mayfly, dragonfly, damselfly, stonefly, caddisfly, mosquito, midge, etc.).
• A family in a phylum other than Arthropoda or Chordata, such as Rotifera,
Annelida, Mollusca.
• A family in any order of insect or any phylum not already represented.
2. Acute-chronic ratios (see section VI) with species of aquatic animals in at least three
different families, provided that:
• At least one is a fish;
• At least one is an invertebrate; and
• At least one is an acutely sensitive freshwater species (the other two may be
saltwater species).
3. Results of at least one acceptable test with a freshwater alga or vascular plant (see
section VIII). If the plants are among the aquatic organisms that are most sensitive to
the material, test data on a plant in another phylum (division) should also be available.
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4. At least one acceptable bioconcentration factor determined with an appropriate
freshwater species, if a maximum permissible tissue concentration is available (see
section IX).
C. To derive a criterion for saltwater aquatic organisms and their uses, the following should
be available:
1. Results of acceptable acute tests (see section IV) with at least one species of saltwater
animal in at least eight different families, including all of the following:
• Two families in the phylum Chordata;
• A family in a phylum other than Arthropoda or Chordata;
• Either the Mysidae or Penaeidae family;
• Three other families not in the phylum Chordata (may include Mysidae or
Penaeidae, whichever was not used previously); and
• Any other family.
2. Acute-chronic ratios (see section VI) with species of aquatic animals in at least three
different families, provided that of the three species:
• At least one is a fish;
• At least one is an invertebrate; and
• At least one is an acutely sensitive saltwater species (the other may be an acutely
sensitive freshwater species).
3. Results of at least one acceptable test with a saltwater alga or vascular plant (see
section VIII). If plants are among the aquatic organisms most sensitive to the material,
results of a test with a plant in another phylum (division) should also be available.
4. At least one acceptable, bioconcentration factor determined with an appropriate
saltwater species, if a maximum permissible tissue concentration is available (see
section IX).
D. If all required data are available, a numerical criterion can usually be derived, except in
special cases. For example, derivation of a criterion might not be possible if the available
acute-chronic ratios vary by more than a factor of 10 with no apparent pattern. Also, if a
criterion is to be related to a water quality characteristic T (see sections V and VII), more
data will be necessary.
Similarly, if all required data are not available, a numerical criterion should not be
derived except in special cases. For example, even if not enough acute and chronic data are
available, it might be possible to derive a criterion if the available data clearly indicate that
the Final Residue Value should be much lower than either the Final Chronic Value or the
Final Plant Value.
E. Confidence in a criterion usually increases as the amount of available pertinent data
increases. Thus, additional data are usually desirable.
IV. Final Acute Value
A. Appropriate measures of the acute (short-term) toxicity of the material to a variety of
species of aquatic animals are used to calculate the Final Acute Value. The Final Acute
Value is an estimate of the concentration of the material, corresponding to a cumulative
probability of 0.05 in the acute toxicity values for genera used in acceptable acute tests
conducted on the material. However, in some cases, if the Species Mean Acute Value of a
commercially or recreationally important species is lower than the calculated Final Acute
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Value, then that Species Mean Acute Value replaces the calculated Final Acute Value to
protect that important species.
B. Acute toxicity tests should have been conducted using acceptable procedures such as
those described in
1. ASTM Standard E 729-80, Practice for Conducting Acute Toxicity Tests with Fishes,
Macroinvertebrates, and Amphibians. American Society for Testing and Materials,
1916 Race Street, Philadelphia, PA 19103.
2. ASTM Standard E 724-80, Practice for Conducting Static Acute Toxicity Tests with
Larvae of Four Species of Bivalve Molluscs. American Society for Testing and
Materials, 1916 Race Street, Philadelphia, PA 19103.
C. Except for tests with saltwater annelids and mysids, do not use results of acute tests
during which test organisms were fed, unless data indicate that the food did not affect the
toxicity of the test material.
D. Results of acute tests conducted in unusual dilution water (e.g., dilution water in which
total organic carbon or particulate matter exceeded 5 mg/L) should not be used unless a
relationship is developed between acute toxicity and organic carbon or particulate matter
or unless data show that the organic carbon or particulate matter does not affect toxicity.
E. Acute values should be based on endpoints that reflect the total severe acute adverse
impact of the test material on the organisms used in the test. Therefore, only the following
kinds of data on acute toxicity to aquatic animals should be used:
1. Tests with daphnids and other cladocerans should be started with organisms less than
24-hours old, and tests with midges should be stressed with second- or third-instar
larvae. The result should be the 48-hour EC50 based on percentage of organisms
immobilized plus percentage of organisms killed. If such an EC50 is not available from
a test, the 48-hour LC50 should be used in place of the desired 48-hour EC50. An EC50
or LC50 of longer than 48 hours can be used as long as the animals were not fed and
the control animals were acceptable at the end of the test.
2. The result of a test with embryos and larvae of barnacles, bivalve molluscs (clams,
mussels, oysters, and scallops), sea urchins, lobsters, crabs, shrimp, and abalones
should be the 96-hour EC50 based on the percentage of organisms with incompletely
developed shells plus the percentage of organisms killed. If such an EC50 is not
available from a test, the lower of the 96-hour EC50, based on the percentage of
organisms with incompletely developed shells and the 96-hour LC50 should be used
in place of the desired 96-hour EC50. If the duration of the test was between 48 and 96
hours, the EC50 or LC50 at the end of the test should be used.
3. The acute values from tests with all other freshwater and saltwater animal species and
older life stages of barnacles, bivalve molluscs, sea urchins, lobsters, crabs, shrimps,
and abalones should be the 96-hour EC50 based on the percentage of organisms
exhibiting loss of equilibrium, plus the percentage of organisms immobilized, plus the
percentage of organisms killed. If such an EC50 is not available from a test, the 96-hour
LC50 should be used in place of the desired 96-hour EC50.
4. Tests with single-celled organisms are not considered acute tests, even if the duration
was 96 hours or less.
5. If the tests were conducted properly, acute values reported as "greater than" values
and those above the solubility of the test material should be used because rejection of
such acute values would unnecessarily lower the Final Acute Value by eliminating
acute values for resistant species.
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F. If the acute toxicity of the material to aquatic animals apparently has been shown to be
related to a water quality characteristic such as hardness or particulate matter for
freshwater animals or salinity or particulate matter for saltwater animals, a Final Acute
Equation should be derived based on that water quality characteristic. (Go to section V.)
G. If the available data indicate that one or more life stages are at least a factor of 2 more resistant
than one or more other life stages of the same species, the data for the more resistant life stages
should not be used in the calculation of the Species Mean Acute Value because a species can be
considered protected from acute toxicity only if all life stages are protected.
H. The agreement of the data within and between species should be considered. Acute values
that appear to be questionable in comparison with other acute and chronic data for the
same species and for other species in the same genus probably should not be used in
calculation of a Species Mean Acute Value. For example, if the acute values available for a
species or genus differ by more than a factor of 10, some or all of the values probably
should not be used in calculations.
I. For each species for which at least one acute value is available, the Species Mean Acute
Value should be calculated as the geometric mean of the results of all flow-through tests in
which the concentrations of test material were measured. For a species for which no such
result is available, the Species Mean Acute Value should be calculated as the geometric
mean of all available.acute values — i.e., results of flow-through tests in which the
concentrations were not measured and results of static and renewal tests based on initial
concentrations of test material. (Nominal concentrations are acceptable for most test
materials if measured concentrations are not available.)
NOTE: Data reported by original investigators should not be rounded off. Results of all
intermediate calculations should be rounded to four significant digits.
NOTE: The geometric mean of N numbers is the Nth root of the product of the N numbers.
Alternatively, the geometric mean can be calculated by adding the logarithms of the N
numbers, dividing the sum by N, and taking the antilog of the quotient. The geometric mean
of two numbers is the square root of the product of the two numbers, and the geometric mean
of one number is that number. Either natural (base e) or common (base 10) logarithms can be
used to calculate geometric means as long as they are used consistently within each set of data
(i.e., the antilog used must match the logarithm used).
NOTE: Geometric means rather than arithmetic means are used here because the distributions
of individual organisms' sensitivities in toxicity tests on most materials, and the distributions
of species' sensitivities within a genus, are more likely to be lognormal than normal. Similarly,
geometric means are used for acute-chronic ratios and bioconcentration factors because
quotients are likely to be closer to lognormal than normal distributions. In addition, division
of the geometric mean of a set of numerators by the geometric mean of the set of
corresponding denominators will result in the geometric mean of the set of corresponding
quotients.
J. The Genus Mean Acute Value should be calculated as the geometric mean of the Species
Mean Acute Values available for each genus.
K. Order the Genus Mean Acute Value from high to low.
L. Assign ranks, R, to the Genus Mean Acute Value from "1" for the lowest to "N" for the
highest. If two or more Genus Mean Acute Values are identical, arbitrarily assign them
successive ranks.
M. Calculate the cumulative probability, P, for each Genus Mean Acute Value as R/ (N+l).
N. Select the four Genus Mean Acute Values that have cumulative probabilities closest to
0.05. (If there are less than 59 Genus Mean Acute Values, these will always be the four
lowest Genus Mean Acute Values).
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O. Using the selected Genus Mean Acute Values and Ps, calculate:
C2 _ S((ln GMAV)2) - ((S(ln GMAV))2/4)
2(P) - ((Z(VF))2/4)
L = (2(ln GMAV) - S(I(VF)))/4
A = S(v^05j + L
FAV = eA
(See original document, referenced at beginning of this appendix, for development of the
calculation procedure and for an example calculation and computer program.)
NOTE: Natural logarithms (logarithms to base e, denoted as In) are used herein merely
because they are easier to use on some hand calculators and computers than common (base 10)
logarithms. Consistent use of either will produce the same result.
P. If for a commercially or recreationally important species the geometric mean of the acute
values from flow-through tests in which the concentrations of test material were
measured is lower than the calculated Final Acute Value, then that geometric mean should
be used as the Final Acute Value instead of the calculated Final Acute Value.
Q. Go to section VI.
Final Acute Equation
A. When enough data are available to show that acute toxicity to two or more species is similarly
related to a water quality characteristic, the relationship should be taken into account as
described in section IV, steps B through G, or using analysis of covariance. The two methods
are equivalent and produce identical results. The manual method described below provides
an understanding of this application of covariance analysis, but computerized versions of
covariance analysis are much more convenient for analyzing large data tests. If two or more
factors affect toxicity, multiple regression analysis should be used.
B. For each species for which comparable acute toxicity values are available at two or more
different values of the water quality characteristic, perform a least squares regression of
the acute toxicity values on the corresponding values of the water quality characteristic to
obtain the slope and its 95 percent confidence limits for each species.
NOTE: Because the best documented relationship is that between hardness and acute toxicity
of metals in freshwater and a log-log relationship fits these data, geometric means and natural
logarithms of both toxicity and water quality are used in the rest of this section. For
relationships based on other water quality characteristics such as pH, temperature, or salinity,
no transformation or a different transformation might fit the data better, and appropriate
changes will be necessary.
C. Decide whether the data for each species are useful, taking into account the range and
number of the tested values of the water quality characteristic and the degree of
agreement within and between species. For example, a slope based on six data points
might be of limited value if based only on data for a very narrow range of water quality
characteristic values. A slope based on only two data points, however, might be useful if
consistent with other information and if the two points cover a broad enough range of the
water quality characteristic.
In addition, acute values that appear to be questionable in comparison with other
acute and chronic data available for the same species and for other species in the same
genus probably should not be used. For example, if after adjustment for the water quality
characteristic the acute values available for a species or genus differ by more than a factor
of 10, probably some or all of the values should be rejected. If useful slopes are not
available for at least one fish and one invertebrate, or if the available slopes are too
dissimilar, or if too few data are available to adequately define the relationship between
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acute toxicity and the water quality characteristic, return to section IV.G, using the results
of tests conducted under conditions and in waters similar to those commonly used for
toxicity tests with the species.
D. Individually for each species, calculate the geometric mean of the available acute values
and then divide each of these acute values by the mean for the species. This normalizes the
values so that the geometric mean of the normalized values for each species, individually,
and for any combination of species is 1.0.
E. Similarly normalize the values of the water quality characteristic for each species,
individually.
F. Individually for each species, perform a least squares regression of the normalized acute
toxicity values on the corresponding normalized values of the water quality characteristic.
The resulting slopes and 95 percent confidence limits will be identical to those obtained in
step B. However, now, if the data are actually plotted, the line of best fit for each individual
species will go through the point 1,1 in the center of the graph.
G. Treat the normalized data as if they were all for the same species and perform a least
squares regression of all the normalized acute values on the corresponding normalized
values of the water quality characteristic to obtain the pooled acute slope, V, and its 95
percent confidence limits. If all the normalized data are actually plotted, the line of best fit
will go through the point 1,1 in the center of the graph.
H. For each species, calculate the geometric mean, W, of the acute toxicity values and the
geometric mean, X, of the values of the water quality characteristic. (These were calculated
in steps D and E.)
I. For each species, calculate the logarithm, Y, of the Species Mean Acute Value at a selected
value, Z, of the water quality characteristic using the equation:
Y= In W- V(ln X-ln Z).
J. For each species, calculate the SMAV at Z using the equation:
SMAV = e*
NOTE: Alternatively, the Species Mean Acute Values at Z can be obtained by skipping step H,
using the equations in steps I and J to adjust each acute value individually to Z, and then
calculating the geometric mean of the adjusted values for each species individually.
This alternative procedure allows an examination of the range of the adjusted acute values
for each species.
K. Obtain the Final Acute Value at Z by using the procedure described in section IV, steps J
through O.
L. If the Species Mean Acute Value at Z of a commercially or recreationally important species
is lower than the calculated Final Acute Value at Z, then that Species Mean Acute Value
should be used as the Final Acute Value at Z instead of the calculated Final Acute Value.
M. The Final Acute Equation is written as:
Final Acute Value = e(V[ln(*'ater
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Final Chronic Value
A. Depending on the data that are available concerning chronic toxicity to aquatic animals,
the Final Chronic Value might be calculated in the same manner as the Final Acute Value
or by dividing the Final Acute Value by the Final Acute-Chronic Ratio. In some cases, it
may not be possible to calculate a Final Chronic Value.
NOTE: As the name implies, the Acute-Chronic Ratio is a way of relating acute and chronic
toxicities. The Acute-Chronic Ratio is basically the inverse of the application factor, but this
new name is better because it is more descriptive and should help prevent confusion between
"application factors" and "safety factors." Acute-Chronic Ratios and application factors are
ways of relating the acute and chronic toxicities of a material to aquatic organisms. Safety
factors are used to provide an extra margin of safety beyond the known or estimated
sensitivities of aquatic organisms. Another advantage of the Acute-Chronic Ratio is that it will
usually be greater than 1; this should avoid the confusion as to whether a large application
factor is one that is close to unity or one that has a denominator that is much greater than the
numerator.
B. Chronic values should be based on results of flow- through chronic tests in which the
concentrations of test material in the test solutions were properly measured at appropriate
times during the test. (Exception: renewal, which is acceptable for daphnids.)
C. Results of chronic tests in which survival, growth, or reproduction in the control treatment
was unacceptably low should not be used. The limits of acceptability will depend on the
species.
D. Results of chronic tests conducted in unusual dilution water (e.g., dilution water in which
total organic carbon or particulate matter exceeded 5 mg/L) should not be used, unless a
relationship is developed between chronic toxicity and organic carbon or particulate
matter, or unless data show that organic carbon, particulate matter (and so forth) do not
affect toxicity.
E. Chronic values should be based on endpoints and lengths of exposure appropriate to the
species. Therefore, only results of the following kinds of chronic toxicity tests should be
used:
1. Life-cycle toxicity tests consisting of exposures of each of two or more groups of
individuals of a species to a different concentration of the test material throughout a
life cycle. To ensure that all life stages and life processes are exposed, tests with fish
should begin with embryos or newly hatched young less than 48-hours old, continue
through maturation and reproduction, and end not less than 24 days (90 days for
salmonids) after the hatching of the next generation. Tests with daphnids should
begin with young less than 24-hours old and last for not less than 21 days. Tests with
mysids should begin with young less than 24-hours old and continue until seven days
past the median time of first brood release in the controls.
For fish, data should be obtained and analyzed on survival and growth of adults
and young, maturation of males and females, eggs spawned per female, embryo
viability (salmonids only), and hatchability. For daphnids, data should be obtained
and analyzed on survival and young per female. For mysids, data should be
obtained and analyzed on survival, growth, and young per female.
2. Partial life-cycle toxicity tests consisting of exposures of each of two or more groups of
individuals in a fish species to a concentration of the test material through most
portions of a life cycle. Partial life-cycle tests are allowed with fish species that require
more than a year to reach sexual maturity so that all major life stages can be exposed to
the test material in less than 15 months.
Exposure to the test material should begin with immature juveniles at least two
months prior to active gonad development, continue through maturation and
reproduction, and end not less than 24 days (90 days for salmonids) after the
hatching of the next generation. Data should be obtained and analyzed on survival
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and growth of adults and young, maturation of males and females, eggs spawned
per female, embryo viability (salmonids only), and hatchability.
3. Early life-stage toxicity tests consisting of 28- to 32-day (60 days post hatch for
salmonids) exposures of the early life stages of a fish species from shortly after
fertilization through embryonic, larval, and early juvenile development. Data should
be obtained and analyzed on survival and growth.
NOTE: Results of an early life-stage test are used as predictions of results of life-cycle and
partial life-cycle tests with the same species. Therefore, when results of a total or partial
life-cycle test are available, results of an early life-stage test with the same species should
not be used. Also, results of early life-stage tests in which the incidence of mortalities or
abnormalities increased substantially near the end should not be used because these
results are possibly not good predictions of the results of comparable life-cycle or partial
life-cycle tests.
F. A chronic value can be obtained by calculating the geometric mean of the lower and upper
chronic limits from a chronic test or by analyzing chronic data using regression analysis. A
lower chronic limit is the highest tested concentration in an acceptable chronic test that did
not cause an unacceptable amount of adverse effect on any of the specified biological
measurements and below which no tested concentration caused an unacceptable effect. An
upper chronic limit is the lowest tested concentration in an acceptable chronic test that did
cause an unacceptable amount of adverse effect on one or more of the specified biological
measurements and above which all tested concentrations also caused such an effect.
NOTE: Because various authors have used a variety of terms and definitions to interpret and
report results of chronic tests, reported results should be reviewed carefully. The amount of
effect that is considered unacceptable is often based on a statistical hypothesis test but might
also be defined in terms of a specified percent reduction from the controls. A small percent
reduction (e.g., 3 percent) might be considered acceptable even if it is statistically significantly
different from the control, whereas a large percent reduction (e.g., 30 percent) might be
considered unacceptable even if it is not statistically significant.
G. If the chronic toxicity of the material to aquatic animals apparently has been shown to be
related to a water quality characteristic such as hardness or particulate matter for
freshwater animals or salinity or particulate matter for saltwater animals, a Final Chronic
Equation should be derived based on that water quality characteristic. Go to section VII.
H. If chronic values are available for species in eight families as described in sections III.B.l or
IH.C.l, a Species Mean Chronic Value should also be calculated for each species for which
at least one chronic value is available by calculating the geometric mean of all chronic
values available for the species; appropriate Genus Mean Chronic Values should also be
calculated. The Final Chronic Value should then be obtained using the procedure
described in section HI, steps J through O. Then go to section VI.M.
I. For each chronic value for which at least one corresponding appropriate acute value is
available, calculate an acute-chronic ratio using for the numerator the geometric mean of
the results of all acceptable flow-through acute tests in the same dilution water and in
which the concentrations were measured. (Exception: static is acceptable for daphnids.)
For fish, the acute test(s) should have been conducted with juveniles and should have
been part of the same study as the chronic test. If acute tests were not conducted as part of
the same study, acute tests conducted in the same laboratory and dilution water but in a
different study may be used. If no such acute tests are available, results of acute tests
conducted in the same dilution water in a different laboratory may be used. If no such
acute tests are available, an acute-chronic ratio should not be calculated.
J. For each species, calculate the species mean acute-chronic ratio as the geometric mean of
all acute-chronic ratios available for that species.
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K. For some materials, the acute-chronic ratio seems to be the same for all species, but for
other materials, the ratio seems to increase or decrease as the Species Mean Acute Value
increases. Thus the Final Acute-Chronic Ratio can be obtained in four ways, depending on
the data available:
1. If the Species Mean Acute-Chronic ratio seems to increase or decrease as the Species
Mean Acute Value increases, the Final Acute-Chronic Ratio should be calculated as the
geometric mean of the acute-chronic ratios for species whose Species Mean Acute
Values are close to the Final Acute Value.
2. If no major trend is apparent, and the acute-chronic ratios for a number of species are
within a factor of 10, the Final Acute-Chronic Ratio should be calculated as the
geometric mean of all the Species Mean Acute-Chronic Ratios available for both
freshwater and saltwater species.
3. For acute tests conducted on metals and possibly other substances with embryos and
larvae of barnacles, bivalve molluscs, sea urchins, lobsters, crabs, shrimp, and
abalones (see section IV.E.2), it is probably appropriate to assume that the
acute-chronic ratio is 2. Chronic tests are very difficult to conduct with most such
species, but the sensitivities of embryos and larvae would likely determine the results
of life-cycle tests. Thus, if the lowest available Species Mean Acute Values were
determined with embryos and larvae of such species, the Final Acute-Chronic Ratio
should probably be assumed to be 2, so that the Final Chronic Value is equal to the
Criterion Maximum Concentration (see section XI.B)
4. If the most appropriate Species Mean Acute-Chronic Ratios are less than 2.0, and
especially if they are less than 1.0, acclimation has probably occurred during the
chronic test. Because continuous exposure and acclimation cannot be assured to
provide adequate protection in field situations, the Final Acute-Chronic Ratio should
be assumed to be 2, so that the Final Chronic Value is equal to the Criterion Maximum
Concentration (see section XI.B).
If the available Species Mean Acute-Chronic Ratios do not fit one of these cases, a
Final Acute-Chronic Ratio probably cannot be obtained, and a Final Chronic Value
probably cannot be calculated.
L. Calculate the Final Chronic Value by dividing the Final Acute Value by the Final
Acute-Chronic Ratio. If there was a Final Acute Equation rather than a Final Acute Value,
see also section VII.A.
M. If the Species Mean Chronic Value of a commercially or recreationally important species is
lower than the calculated Final Chronic Value, then that Species Mean Chronic Value
should be used as the Final Chronic Value instead of the calculated Final Chronic Value.
N. Go to section VIII.
VII. Final Chronic Equation
A. A Final Chronic Equation can be derived in two ways. The procedure described here will
result in the chronic slope being the same as the acute slope. The procedure described in
steps B through N usually will result in the chronic slope being different from the acute
slope.
1. If acute-chronic ratios are available for enough species at enough values of the water
quality characteristic to indicate that the acute-chronic ratio is probably the same for
all species and is probably independent of the water quality characteristic, calculate
the Final Acute-Chronic Ratio as the geometric mean of the available Species Mean
Acute-Chronic Ratios.
2. Calculate the Final Chronic Value at the selected value Z of the water quality
characteristic by dividing the Final Acute Value at Z (see section V.M) by the Final
Acute-Chronic Ratio.
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3. Use V = pooled acute slope (see section V.M) as L = pooled chronic slope.
4. Go to section VII.M.
B. When enough data are available to show that chronic toxicity to at least one species is
related to a water quality characteristic, the relationship should be taken into account as
described in steps B through G or using analysis of covariance. The two methods are
equivalent and produce identical results. The manual method described in the following
paragraphs provides an understanding of this application of covariance analysis, but
computerized versions of covariance analysis are much more convenient for analyzing
large data sets. If two or more factors affect toxicity, multiple regression analysis should be
used.
C. For each species for which comparable chronic toxicity values are available at two or more
different values of the water quality characteristic, perform a least squares regression of
the chronic toxicity values on the corresponding values of the water quality characteristic
to obtain the slope and its 95 percent confidence limits for each species.
NOTE: Because the best-documented relationship is that between hardness and acute toxicity
of metals in fresh water and a log-log relationship fits these data, geometric means and natural
logarithms of both toxicity and water quality are used in the rest of this section. For
relationships based on other water quality characteristics such as pH, temperature, or salinity,
no transformation or a different transformation might fit the data better, and appropriate
changes will be necessary throughout this section. It is probably preferable, but not necessary,
to use the same transformation that was used with the acute values in section V.
D. Decide whether the data for each species are useful, taking into account the range and
number of the tested values of the water quality characteristic and the degree of
agreement within and between species. For example, a slope based on six data points
might be of limited value if based only on data for a very narrow range of values of the
water quality characteristic. A slope based on only two data points, however, might be
useful if it is consistent with other information and if the two points cover a broad enough
range of the water quality characteristic. In addition, chronic values that appear to be
questionable in comparison with other acute and chronic data available for the same
species and for other species in the same genus probably should not be used. For example,
if after adjustment for the water quality characteristic the chronic values available for a
species or genus differ by more than a factor of 10, probably some or all of the values
should be rejected.
If a useful chronic slope is not available for at least one species, or if the available
slopes are too dissimilar, or if too few data are available to adequately define the
relationship between chronic toxicity and the water quality characteristic, the chronic
slope is probably the same as the acute slope, which is equivalent to assuming that the
acute-chronic ratio is independent of the water quality characteristic. Alternatively, return
to section VI.H, using the results of tests conducted under conditions and in waters similar
to those commonly used for toxicity tests with the species.
E. Individually for each species, calculate the geometric mean of the available chronic values
and then divide each chronic value for a species by its mean. This normalizes the chronic
values so that the geometric mean of the normalized values for each species individually,
and for any combination of species, is 1.0.
F. Similarly normalize the values of the water quality characteristic for each species,
individually.
G. Individually for each species, perform a least squares regression of the normalized chronic
toxicity values on the corresponding normalized values of the water quality characteristic.
The resulting slopes and the 95 percent confidence limits will be identical to those
obtained in section B. Now, however, if the data are actually plotted, the line of best fit for
each individual species will go through the point 1,1 in the center of the graph.
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H. Treat all the normalized data as if they were all for the same species and perform a least
squares regression of all the normalized chronic values on the corresponding normalized
values of the water quality characteristic to obtain the pooled chronic slope, L, and its 95
percent confidence limits. If all the normalized data are actually plotted, the line of best fit
will go through the point 1,1 in the center of the graph.
I. For each species, calculate the geometric mean, M, of the toxicity values and the geometric
mean, P, of the values of the water quality characteristic. (These were calculated in steps E
and F.)
J. For each species, calculate the logarithm, Q, of the Species Mean Chronic Value at a
selected value, Z, of the water quality characteristic using the equation:
Q = In M - L(ln P - In Z).
NOTE: Although it is not necessary, it will usually be best to use the same value of the water
quality characteristic here as was used in section V.I.
K. For each species, calculate a Species Mean Chronic Value at Z using the equation:
SMCV = eQ.
NOTE: Alternatively, the Species Mean Chronic Values at Z can be obtained by skipping step J,
using the equations in steps J and K to adjust each acute value individually to Z, and then
calculating the geometric means of the adjusted values for each species individually. This
alternative procedure allows an examination of the range of the adjusted chronic values for
each species.
L. Obtain the Final Chronic Value at Z by using the procedure described in section IV, steps J
through O.
M. If the Species Mean Chronic Value at Z of a commercially or recreationally important
species is lower than the calculated Final Chronic Value at Z, then that Species Mean
Chronic Value should be used as the Final Chronic Value at Z instead of the calculated
Final Chronic Value.
N. The Final Chronic Equation is written as:
Final Chronic Value = e' characteristic)] ~ in s - L[in Z])
where
L = pooled chronic slope
S = Final Chronic Value at Z.
Because L, S, and Z are known, the Final Chronic Value can be calculated for any selected
value of the water quality characteristic.
VIII. Final Plant Value
A. Appropriate measures of the toxicity of the material to aquatic plants are used to compare the
relative sensitivities of aquatic plants and animals. Although procedures for conducting and
interpreting the results of toxicity tests with plants are not well developed, results of tests with
plants usually indicate that criteria which adequately protect aquatic animals and their uses
will probably also protect aquatic plants and their uses.
B. A plant value is the result of a 96-hour test conducted with an alga, or a chronic test
conducted with an aquatic vascular plant.
NOTE: A test of the toxicity of a metal to a plant usually should not be used if the medium
contained an excessive amount of a complexing agent, such as EDTA, that might affect the
toxicity of the metal. Concentrations of EDTA above about 200 ng/L should probably be
considered excessive.
C. The Final Plant Value should be obtained by selecting the lowest result from a test with an
important aquatic plant species in which the concentrations of test material were
measured, and the endpoint was biologically important.
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IX. Final Residue Value
A. The Final Residue Value is intended to prevent concentrations in commercially or
recreationally important aquatic species from affecting marketability because they exceed
applicable FDA action levels and to protect wildlife (including fishes and birds) that
consume aquatic organisms from demonstrated unacceptable effects. The Final Residue
Value is the lowest of the residue values that are obtained by dividing maximum
permissible tissue concentrations by appropriate bioconcentration or bioaccumulation
factors. A maximum permissible tissue concentration is either (a) an FDA action level
(Compliance Policy Guide, U.S. Food & Drug Admin. 1981) for fish oil or for the edible
portion of fish or shellfish, or a maximum acceptable dietary intake based on observations
on survival, growth, or reproduction in a chronic wildlife feeding study or a long-term
wildlife field study. If no maximum permissible tissue concentration is available, go to
section X because no Final Residue Value can be derived.
B. Bioconcentration Factors (BCFs) and bioaccumulation factors (BAFs) are quotients of the
concentration of a material in one or more tissues of an aquatic organism, divided by the
average concentration in the solution in which the organism had been living. A BCF is
intended to account only for net uptake directly from water and thus almost must be
measured in a laboratory test. Some uptake during the bioconcentration test might not be
directly from water if the food sorbs some of the test material before it is eaten by the test
organisms. A BAF is intended to account for net uptake from both food and water in a
real-world situation. A BAF almost must be measured in a field situation in which
predators accumulate the material directly from water and by consuming prey that could
have accumulated the material from both food and water.
The BCF and BAF are probably similar for a material with a low BCF, but the BAF is
probably higher than the BCF for materials with high BCFs. Although BCFs are not too
difficult to determine, very few BAFs have been measured acceptably because adequate
measurements must be made of the material's concentration in water to ascertain if it was
reasonably constant for a long enough time over the range of territory inhabited by the
organisms. Because so few acceptable BAFs are available, only BCFs will be discussed
further. However, if an acceptable BAF is available for a material, it should be used instead
of any available BCFs.
C. If a maximum permissible tissue concentration is available for a substance (e.g., parent
material, parent material plus metabolites, etc.), the tissue concentration used in the
calculation of the BCF should be for the same substance. Otherwise, the tissue
concentration used in the calculation of the BCF should derive from the material and its
metabolites that are structurally similar and are not much more soluble in water than the
parent material.
1. A BCF should be used only if the test was flow-through, the BCF was calculated based
on measured concentrations of the test material in tissue and in the test solution, and
the exposure continued at least until either apparent steady state or 28 days was
reached. Steady state is reached when the BCF does not change significantly over a
period of time, such as 2 days or 16 percent of the length of the exposure, whichever is
longer. The BCF used from a test should be the highest of the apparent steady-state
BCF, if apparent steady state was reached; the highest BCF obtained, if apparent
steady state was not reached; and the projected steady state BCF, if calculated.
2. Whenever a BCF is determined for a lipophilic material, the percent lipids should also
be determined in the tissue(s) for which the BCF was calculated.
3. A BCF obtained from an exposure that adversely affected the test organisms may be
used only if it is similar to a BCF obtained with unaffected organisms of the same
species at lower concentrations that did not cause adverse effects.
4. Because maximum permissible tissue concentrations are almost never based on dry
weights, a BCF calculated using dry tissue weights must be converted to a wet tissue
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weight basis. If no conversion factor is reported with the BCF, multiply the dry weight
BCF by 0.1 for plankton and by 0.2 for individual species of fishes and invertebrates.
5. If more than one acceptable BCF is available for a species, the geometric mean of the
available values should be used; however, if the BCFs are from different lengths of
exposure and the BCF increases with length of exposure, then the BCF for the longest
exposure should be used.
E. If enough pertinent data exist, several residue values can be calculated by dividing
maximum permissible tissue concentrations by appropriate BCFs:
1. For each available maximum acceptable dietary intake derived from a chronic feeding
study or a long-term field study with wildlife (including birds and aquatic organisms),
the appropriate BCF is based on the whole body of aquatic species that constitutes or
represents a major portion of the diet of the tested wildlife species.
2. For an FDA action level for fish or shellfish, the appropriate BCF is the highest
geometric mean species BCF for the edible portion (muscle for decapods, muscle with
or without skin for fishes, adductor muscle for scallops, and total soft tissue for other
bivalve molluscs) of a consumed species. The highest species BCF is used because FDA
action levels are applied on a species-by-species basis.
F. For lipophilic materials, calculating additional residue values is possible. Because the
steady-state BCF for a lipophilic material seems to be proportional to percent lipids from
one tissue to another and from one species to another, extrapolations can be made from
tested tissues, or species to untested tissues, or species on the basis of percent lipids.
1. For each BCF for which the percent lipids is known for the same tissue for which the
BCF was measured, normalize the BCF to a 1 percent lipid basis by dividing it by the
percent lipids. This adjustment to a 1 percent lipid basis is intended to make all the
measured BCFs for a material comparable regardless of the species or tissue with
which the BCF was measured.
2. Calculate the geometric mean normalized BCF. Data for both saltwater and
freshwater species should be used to determine the mean normalized BCF unless they
show that the normalized BCFs are probably not similar.
3. Calculate all possible residue values by dividing the available maximum permissible
tissue concentrations by the mean normalized BCF and by the percent lipids values
appropriate to the maximum permissible tissue concentrations, i.e.,
Residue value = (max/mum permissible tissue concentration)
~ (mean normalized BCF)(appropriate percent lipids)
• For an FDA action level for fish oil, the appropriate percent lipids value is 100.
• For an FDA action level for fish, the appropriate percent lipids value is 11 for
freshwater criteria and 10 for saltwater criteria because FDA action levels are
applied species-by-species to commonly consumed species. The highest lipid
contents in the edible portions of important consumed species are about 11
percent for both the freshwater chinook salmon and lake trout and about 10
percent for the saltwater Atlantic herring.
• For a maximum acceptable dietary intake derived from a chronic feeding study or
a long-term field study with wildlife, the appropriate percent lipids is that of an
aquatic species or group of aquatic species that constitute a major portion of the
diet of the wildlife species.
G. The Final Residue Value is obtained by selecting the lowest of the available residue values.
NOTE: In some cases, the Final Residue Value will not be low enough. For example, a residue
value calculated from a FDA action level will probably result in an average concentration in
the edible portion of a fatty species at the action level. Some individual organisms and
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possibly some species will have residue concentrations higher than the mean value, but no
mechanism has been devised to provide appropriate additional protection. Also, some
chronic feeding studies and long-term field studies with wildlife identify concentrations that
cause adverse effects but do not identify concentrations that do not cause adverse effects;
again, no mechanism has been devised to provide appropriate additional protection. These
are some of the species and uses that are not protected at all times in all places.
X. Other Data
Pertinent information that could not be used in earlier sections might be available concerning
adverse effects on aquatic organisms and their uses. The most important of these are data on
cumulative and delayed toxicity, flavor impairment, reduction in survival, growth, or
reproduction, or any other adverse effect shown to be biologically important. Especially
important are data for species for which no other data are available. Data from behavioral,
biochemical, physiological, microcosm, and field studies might also be available. Data might be
available from tests conducted in unusual dilution water (see IV.D and VI.D), from chronic tests
in which the concentrations were not measured (see VI.B), from tests with previously exposed
organisms (see II.F), and from tests on formulated mixtures or emulsifiable concentrates (see
II.D). Such data might affect a criterion if they were obtained with an important species, the test
concentrations were measured, and the endpoint was biologically important.
XI. Criterion
A. A criterion consists of two concentrations: the Criterion Maximum Concentration and the
Criterion Continuous Concentration.
B. The Criterion Maximum Concentration (CMC) is equal to one-half the Final Acute Value.
C. The Criterion Continuous Concentration (CCC) is equal to the lowest of the Final Chronic
Value, the Final Plant Value, and the Final Residue Value, unless other data (see section X)
show that a lower value should be used. If toxicity is related to a water quality characteristic,
the Criterion Continuous Concentration is obtained from the Final Chronic Equation, the
Final Plant Value, and the Final Residue Value by selecting the one, or the combination, that
results in the lowest concentrations in the usual range of the water quality characteristic,
unless other data (see section X) show that a lower value should be used.
D. Round both the Criterion Maximum Concentration and the Criterion Continuous
Concentration to two significant digits.
E. The criterion is stated as follows:
The procedures described in the "Guidelines for Deriving Numerical National Water
Quality Criteria for the Protection of Aquatic Organisms and Their Uses" indicate that,
except possibly where a locally important species is very sensitive, (1) aquatic organisms
and their uses should not be affected unacceptably if the four-day average concentration
of (2) does not exceed (3) fig/L more than once every three years on the average, and if the
one-hour average concentration does not exceed (4) ng/L more than once every three
years on the average.
where (1) = insert freshwater or saltwater
(2) = insert name of material
(3) = insert the Criterion Continuous Concentration
(4) = insert the Criterion Maximum Concentration.
XII. Final Review
A. The derivation of the criterion should be carefully reviewed by rechecking each step of the
guidelines. Items that should be especially checked are
1. If unpublished data are used, are they well documented?
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2. Are all required data available?
3. Is the range of acute values for any species greater than a factor of 10?
4. Is the range of Species Mean Acute Values for any genus greater than a factor of 10?
5. Is there more than a factor of 10 difference between the four lowest Genus Mean Acute
Values?
6. Are any of the four lowest Genus Mean Acute Values questionable?
7. Is the Final Acute Value reasonable in comparison with the Species Mean Acute Values
and Genus Mean Acute Values?
8. For any commercially or recreationally important species, is the geometric mean of the
acute values from flow-through tests in which the concentrations of test material were
measured lower than the Final Acute Value?
9. Are any of the chronic values questionable?
10. Are chronic values available for acutely sensitive species?
11. Is the range of acute-chronic ratios greater than a factor of 10?
12. Is the Final Chronic Value reasonable in comparison with the available acute and
chronic data?
13. Is the measured or predicted chronic value for any commercially or recreationally
important species below the Final Chronic Value?
14. Are any of the other data important?
15. Do any data look like they might be outliers?
16. Are there any deviations from the guidelines? Are they acceptable?
B. On the basis of all available pertinent laboratory and field information, determine if the
criterion is consistent with sound scientific evidence. If not, another criterion — either
higher or lower — should be derived using appropriate modifications of these guidelines.
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APPENDIX B
Derivation of the 1980
Aquatic Life Criteria
This version of the Guidelines provides clarifications, additional details, and technical and editorial
changes in the last version published in the Federal Register [44 FR 15970 (March 15, 1979)]. It
incorporates changes resulting from comments on previous versions and from experience gained
during EPA's use of the previous versions. Future versions of the Guidelines will incorporate new
ideas and data as their usefulness is demonstrated.
Criteria may be expressed in several forms. The numerical form is commonly used, but
descriptive and procedural forms can be used if numerical criteria are not possible or desirable. The
purpose of these Guidelines is to describe an objective, internally consistent and appropriate way of
deriving numerical water quality criteria for the protection of the uses and presence of aquatic
organisms.
A numerical criterion is an estimate of the highest concentration of a substance in water that
does not present a significant risk to the aquatic organisms in the water and their uses. Thus the
Guidelines are intended to derive criteria that will protect aquatic communities by protecting most
of the species and their uses most of the time, but not necessarily all of the species all of the time.
Aquatic communities can tolerate some stress and occasional adverse effects on a few species, and
so total protection of all species all of the time is not necessary. Rather, the Guidelines attempts to
provide a reasonable and adequate amount of protection with only a small possibility of
considerable overprotection or underprotection. Within these constraints, it seems appropriate to
err on the side of overprotection.
The numerical aquatic life criteria derived using the Guidelines are expressed as two numbers,
rather than the traditional one number, so that the criteria can more accurately reflect toxicological
and practical realities. The combination of both a maximum value and a 24-hour average value is
designed to provide adequate protection of aquatic life and its uses from acute and chronic toxicity
to animals, toxicity to plants, and bioconcentration by aquatic organisms without the restrictions of
a one-number criterion to provide the same amount of protection. The only way to assure the same
degree of protection with a one-number criterion would be to use the 24-hour average as a
concentration that is not to be exceeded at any time in any place.
The two-number criterion is intended to identify an average pollutant concentration that will
produce a water quality generally suited to the maintenance of aquatic life and its uses, while
restricting the extent and duration of excursions over the average so that the total exposure will not
cause unacceptable adverse effects. Merely specifying an average value over a time period is
insufficient, unless the period of time is rather short, because concentration higher than the average
value can kill or cause substantial damage in short periods. Furthermore, for some substances the
effect of intermittent high exposures is cumulative. Therefore, placing an upper limit on pollutant
concentrations to which aquatic organisms might be exposed is necessary, especially when the
maximum value is not much higher than the average value. For some substances, the maximum may
be so much higher than the 24-hour average that in any real-world situation the maximum will never
be reached if the 24-hour average is achieved. In such cases, the 24-hour average will be limiting and
the maximum will have no practical significance, except to indicate that elevated concentrations are
acceptable as long as the 24-hour average is achieved.
These Guidelines have been developed on the assumption that the results of laboratory tests are
generally useful for predicting what will happen in field situations. The resulting criteria are meant to
apply to most U.S. bodies of water, except for the Great Salt Lake. All aquatic organisms and their
common uses are meant to be considered, but not necessarily protected, if relevant data are available,
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with at least one specific exception. This exception is the accumulation of residues of organic
compounds in the siscowet subspecies of lake trout which occurs in Lake Superior and contains up to 67
percent fat in the fillets (Thurston, C.E., 1962, Physical Characteristics and Chemical Composition of
Two Subspecies of Lake Trout,}. Fish. Res. Bd. Canada 19:39-44). Neither siscowet nor organisms in the
Great Salt Lake are intentionally protected by these Guidelines because both may be too atypical.
With appropriate modifications, these Guidelines can be used to derive criteria for any specified
geographical area, body of water (such as the Great Salt Lake), or group of similar bodies of water.
Thus, with appropriate modifications, the Guidelines can be used to derive national, state, or local
criteria if adequate information is available concerning the effects of the substance of concern on
appropriate species and their uses. However, the basic concepts described in the Guidelines should
be modified only when sound scientific evidence indicates that a criterion produced using the
Guidelines would probably significantly overprotect or underprotect the presence or uses of
aquatic life.
Criteria produced by these Guidelines are not enforceable numbers. They may be used in
developing enforceable numbers, such as water quality standards and effluent standards. However,
the development of standards may take into account additional factors such as social, legal,
economic, and hydrological considerations, the environmental and analytical chemistry of the
substance, the extrapolation from laboratory data to field situations, and the relationship between
the species for which data are available and the species to be protected.
Because fresh water and salt water (including both estuarine and marine waters) have basically
different chemical compositions and because freshwater and saltwater species rarely inhabit the
same water simultaneously, separate criteria should be derived for these two kinds of waters.
However, for some substances sufficient data may not be available to allow derivation of one or
both of these criteria using the Guidelines.
These Guidelines are meant to be used after a decision is made that a criterion is needed for a
substance. The Guidelines do not address the rationale for making that decision. If the potential for
adverse effects on aquatic life and its uses are part of the basis for deciding whether or not a criterion
is needed for a substance, these Guidelines may be helpful in collecting and interpretating relevant
data.
I. Define the Substance for Which the Criterion is to be Derived
A. Each separate chemical that would not ionize significantly in most natural bodies of water
should usually be considered a separate substance, except possibly for structurally similar
organic compounds that only differ in the number and location of atoms of a specific
halogen, and only exist in large quantities as commercial mixtures of the various
compounds, and apparently have similar chemical, biological, and toxicological
properties.
B. For chemicals that would ionize significantly in most natural bodies of water, such as
inorganic salts, organic acids and phenols, all forms that would be in chemical equilibrium
should usually be considered one substance. For metals, each different valence and each
different covalently bonded organometallic compound should usually be considered a
separate substance.
C. The definition of the substance may also need to take into account the analytical chemistry
and fate of the substance.
II. Collect and Review Available Data
A. Collect all available data on the substance concerning
1. toxicity to, and bioaccummulation by, aquatic animals and plants
2. FDA action levels
3. chronic feeding studies with wildlife
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B. Discard all data that are not available in hard copy (publication, manuscript, letter,
memorandum, etc.) with enough supporting information to indicate that acceptable test
procedures were used and that the results are reliable. Do not assume that all published
data are acceptable.
C. Discard questionable data. For example, discard data from tests for which no control
treatment existed, in which too many organisms in the control treatment died or showed
signs of stress or disease, or in which distilled or deionized water was used as the dilution
water for aquatic organisms. Discard data on formulated mixtures and emulsifiable
concentrates of the substance of concern, but not necessarily data on technical grade
material.
D. Do not use data obtained using
1. Brine shrimp, because they usually only occur naturally in water with salinity greater
than 35g/kg.
2. Species that do not have reproducing wild populations resident in — but not
necessarily native to — North America. Resident North American species of fishes are
defined as those listed in "A List of Common and Scientific Names of Fishes from the
United States and Canada," 3rd ed., Special Publication No. 6, American Fisheries
Society, Washington, D.C., 1970. Data obtained with non-resident species can be used
to indicate relationships and possible problem areas, but cannot be used in the
derivation of criteria.
3. Organisms that were previously exposed to significant concentrations of the test
material or other pollutants.
Minimum Data Base
A. A minimum amount of data should be available to help ensure that each of the four major
kinds of possible adverse effects receives some consideration. Results of acute and chronic
toxicity tests with a reasonable number and variety of aquatic animals are necessary so
that data available for tested species can be considered a useful indication of the
sensitivities of the numerous untested species. The requirements concerning toxicity to
aquatic plants are less stringent because procedures for conducting tests with plants are
not as well developed and the interpretation of the results is more questionable. Data
concerning bioconcentration by aquatic organisms can only be used if other relevant data
are available.
B. To derive a criterion for freshwater aquatic life, the following should be available:
1. Acute tests (see section IV) with freshwater animals in at least eight different families
provided that of the eight species at least one
• Is a salmonid fish
• Is a non-salmonid fish
• Is a planktonic crustacean
• Is a benthic crustacean
• Is a benthic insect
• Of the benthic species is a detritivore
2. Acute-chronic ratios (see section VI) for at least three species of aquatic animals
provided that of the three species at least one is
• A fish
• An invertebrate
• A freshwater species (the other two may be saltwater species)
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3. At least one test with a freshwater alga or a chronic test with a freshwater vascular
plant (see section VIII). If plants are among the aquatic organisms that are most
sensitive to the substance, tests with more than one species should be available.
4. At least one acceptable bioconcentration factor determined with an aquatic animal
species, if a maximum permissible tissue concentration is available (see section IX).
C. To derive a criterion for saltwater aquatic life, the following should be available:
1. Acute tests (see section IV) with saltwater animals in at least eight different families
provided that of the eight species
• At least two different fish families are included
• At least five different invertebrate families are included
• Either the Mysidae or Penaeidae family or both are included
• At least one of the invertebrate families is in a phylum other than Arthropoda
2. Acute-chronic ratios (see section VI) for at least three species of aquatic animals
provided that of the three species at least one is
• A fish
• An invertebrate
• A saltwater species (the other two may be freshwater species)
3. At least one test with a saltwater vascular plant (see section VIII). If plants are among
the aquatic organisms most sensitive to the substance, tests with more than one
species should be available.
4. At least one acceptable bioconcentration factor determined with an aquatic animal
species, if a maximum permissible tissue concentration is available (see section IX).
D. If all the requirements of the minimum data base are met, a criterion can usually be
derived, except in special cases. For example, a criterion might not be possible if the
acute-chronic ratios vary greatly with no apparent pattern. Also, if a criterion is to be
related to a water quality characteristic (see sections V and VII), more data will be
necessary.
Similarly, if the minimum data requirements are not satisfied, generally a criterion
should not be derived, except in special cases. One such special case would be when less
than the minimum amount of acute and chronic data are available, but the available data
clearly indicate that the Final Residue Value would be substantially lower then either the
Final Chronic Value or the Final Plant Value.
IV. Final Acute Value
A. Appropriate measures of the acute (short-term) toxicity of the substance to various species
of aquatic animals are used to calculate the Final Acute Value. If acute values are available
for fewer than 20 species, the Final Acute Value probably should be lower than the lowest
value. On the other hand, if acute values are available for more than 20 species, the Final
Acute Value probably should be higher than the lowest value, unless the most sensitive
species is an important one. Although the procedure used to calculate the Final Acute
Value has some limitations, it apparently is the best of the procedures currently available.
B. Acute toxicity tests should be conducted using procedures such as those described in
1. ASTM Standard E 729-80, Practice for Conducting Acute Toxicity Tests with Fishes,
Macroinvertebrates, and Amphibians. American Society for Testing and Materials,
1916 Race Street, Philadelphia, PA 19103.
2. ASTM Standard E 724-80, Practice for Conducting Static Acute Toxicity Tests with
Larvae of Four Species of Bivalve Molluscs. American Society for Testing and
Materials, 1916 Race Street, Philadelphia, PA 19103.
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C. Results of acute tests in which food was added to the test solutions should not be used,
because this may unnecessarily affect the results of the test.
D. Results of acute tests conducted with embryos should not be used (but see section IV.E.2),
because this is often an insensitive life stage.
E. Acute values should be based on endpoints and lengths of exposure appropriate to the life
stage of the species tested. Therefore, only the following kinds of data on acute toxicity to
aquatic animals should be used:
1. 48-hour EC50 values based on immobilization and 48-hour LC50 values for first-instar
(less than 24 hours old) daphnids and other cladocerans, and second- or third-instar
midge larvae.
2. 48- to 96-hour EC50 values based on incomplete shell development and 48- to 96-hour
LC50 values for embryos and larvae of barnacles, bivalve molluscs (clams, mussels,
oysters, and scallops), sea urchins, lobsters, crabs, shrimps, and abalones.
3. 96-hour EC50 values based on decreased shell deposition for oysters.
4. 96-hour EC50 values on immobilization or loss of equilibrium or both and 96-hour
LC50 values for aquatic animals, except for cladocerans, midges, and animals whose
behavior or physiology allows them to avoid exposure to toxicant or for whom the
acute adverse effect of the exposure cannot be adequately measured. Such freshwater
and saltwater animals include air-breathing molluscs, unionid clams, operculate
snails, and bivalve molluscs, except for some species that cannot "close up" and thus
prevent exposure to toxicant, such as the bay scallop (Argopecten irradians).
F. For the use of LC50 or EC50 values for durations shorter and longer than those listed
above, see section X.
G. If the acute toxicity of the substance to aquatic animals has been shown to be related to a
water quality characteristic such as hardness for freshwater organisms or salinity for
saltwater organisms, a Final Acute Equation should be derived based on that water
quality characteristic. Go to section V.
H. If the acute toxicity of the substance has not been adequately shown to be related to a
water quality characteristic, for each species for which at least one acute value is available,
calculate the geometric mean of the results of all flow-through tests in which the toxicant
concentrations were measured. For a species for which no such result is available,
calculate the geometric mean of all available acute values — i.e., results of flow-through
tests in which the toxicant concentrations were not measured and results of static and
renewal tests based on initial total toxicant concentrations.
NOTE: The geometric mean of N numbers is obtained by taking the Nth root of the product of
N numbers. Alternatively, the geometric mean can be calculated by adding the logarithms of
the N numbers, dividing the sum by N, and taking the antilog of the quotient. The geometric
mean of two numbers can also be calculated as the square root of the product of the two
numbers. The geometric mean of one number is that number. Either natural (base e) or
common (base 10) logarithms can be used to calculate geometric means as long as they are
used consistently within each set of data — i.e., the antilog used must match the logarithm
used.
I. Count the number = N of species for which a species mean acute value is available.
J. Order the species mean acute values from low to high. Take the common logarithms of the
N values (log mean values).
K. The intervals (cell widths) for the lower cumulative proportion calculations are 0.11
common log units apart, starting from the lowest log value. The value of 0.11 is an estimate
of average precision and was calculated from replicate species acute values.
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L. Starting with the lowest log mean value, separate the N values into intervals (or cells)
calculated in step IV.K.
M. Calculate cumulative proportions for each non-empty interval by summing the number of
values in the present and all lower intervals and dividing by N. These calculations only
need to be done for the first three non-empty intervals (or cells).
N. Calculate the arithmetic mean of the log mean values for each of the three intervals.
O. Using the two interval mean acute values and cumulative proportions closest to 0.05,
linearly extrapolate or interpolate to the 0.05 log concentration. The Final Acute Value is
the antilog of the 0.05 concentration.
In other words, where
Prop (1) and cone (1) are the cumulative proportion and mean log value for the
lowest non-empty interval.
Prop (2) and cone (2) are the cumulative proportion and mean log value for the
second lowest non-empty interval.
A = Slope of the cumulative proportions
B = The 0.05 log value
then:
A = [0.05 - Prop (1)] / [Prop(2) - Prop (1)]
B = cone (1) + A [cone (2) - cone (1)]
Final Acute Value = 105
P. If for an important species, such as a recreationally or commercially important species, the
geometric mean of the acute values from flow-through tests in which the toxicant
concentrations were measured is lower than the Final Acute Value, then that geometric
mean should be used as the Final Acute Value.
Q. Go to section VI.
Final Acute Equation
A. When enough data are available to show that acute toxicity to two or more species is
similarly affected by a water quality characteristic, this effect can be taken into account as
described below. Pooled regression analysis should produce similar results, although data
available for individual species would be weighted differently.
B. For each species for which comparable acute toxicity values are available at two or more
different values of a water quality characteristic which apparently affects toxicity, perform
a least squares regression of the natural logarithms of the acute toxicity values on the
natural logarithms of the values of the water quality characteristic. Natural logarithms
(logarithms to the base e, denoted as In) are used herein merely because they are easier to
use on some hand calculators and computers than common logarithms (logarithms to the
base 10). Consistent use of either will produce the same result. No transformation or a
different transformation may be used if it fits the data better, but appropriate changes will
be necessary throughout this section.
C. Determine whether or not each acute slope is meaningful, taking into account the range
and number of values of the water characteristic tested. For example, a slope based on four
data points may be of limited value if it is based only on data for a narrow range of values
of the water quality characteristic. On the other hand, a slope based on only two data
points may be meaningful if it is consistent with other information and if the two points
cover a broad enough range of the water quality characteristic. If meaningful slopes are
not available for at least two species, or if the available slopes are not similar, return to
section IV.H., using the results of tests conducted under conditions and in water similar to
those commonly used for toxicity tests with the species.
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D. Calculate the mean acute slope (V) as the arithmetic average of all the meaningful acute
slopes for individual species.
E. For each species calculate the geometric mean (W) of the acute toxicity values and the
geometric mean (X) of the related values of the water quality characteristics.
F. For each species calculate the logarithmic intercept (Y) using the equation:
Y = in W - V(lnX).
G. For each species calculate the species mean acute intercept as the antilog of Y.
H. Obtain the Final Acute Intercept by using the procedure described in section IV.I-O, except
insert "Intercept" for "Value."
I. If for an important species, such as a recreationally or commercially important species, the
intercept calculated only from results of flow-through tests in which the toxicant
concentrations were measured is lower than the Final Acute Intercept, then that intercept
should be used as the Final Acute Intercept.
J. The Final Acute Equation is written as
g(v[ln(water quality characteristic)+ln z)
where
V = mean acute slope
Z = Final Acute Intercept.
Final Chronic Value
A. The Final Chronic Value can be calculated in the same manner as the Final Acute Value or
by dividing the Final Acute Value by the Final Acute Chronic Ratio, depending on the data
available. In some cases it will not be possible to calculate a Final Chronic Value.
B. Use only the results of flow-through (except renewal is acceptable for daphnids) chronic
tests in which the concentrations of toxicant in the test solutions were measured.
C. Do not use the results of any chronic test in which survival, growth, or reproduction
among the controls was unacceptably low.
D. Chronic values should be based on endpoints and lengths of exposure appropriate to the
species. Therefore, only the results of the following kinds of chronic toxicity tests should
be used:
1. Life-cycle toxicity tests consisting of exposures of each of several groups of
individuals of a species to a different concentration of the toxicant throughout a life
cycle. To ensure that all life stages and life processes are exposed, the test should begin
with embryos or newly hatched young less than 48 hours old (less than 24 hours old
for daphnids), continue through maturation and reproduction, and with fish should
end not less than 24 days (90 days for salmonids) after the hatching of the next
generation. For fish, data should be obtained and analyzed on survival and growth of
adults and young, maturation of males and females, embryos spawned per female,
embryo viability (salmonids only) and hatchability. For daphnids, data should be
obtained and analyzed on survival and young per female.
2. Partial life-cycle toxicity tests consisting of exposures of each of several groups of
individuals of a species of fish to a different concentration of the toxicant through most
portions of a life cycle. Partial life-cycle tests are conducted with fish species that
require more than a year to reach sexual maturity, so that the test can be completed in
less than 15 months, but still expose all major life stages to the toxicant. Exposure to
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the toxicant begins with immature juveniles at least 2 months prior to active gonad
development, continues through maturation and reproduction, and ends not less than
24 days (90 days for salmonids) after the hatching of the next generation. Data should
be obtained and analyzed on survival and growth of adults and young, maturation of
males and females, embryos spawned per female, embryo viability (salmonids only),
and hatchability.
3. Early-life-stage toxicity tests consisting of 28- to 32-days (60 days post-hatch for
salmonids) exposures of the early life stages of a species of fish from shortly after
fertilization through embryonic, larval, and early juvenile development. Data should
be obtained and analyzed on survival and growth.
E. Do not use the results of an early-life-stage test if results of a life-cycle or partial life-cycle
test with the same species are available.
F. A chronic value is obtained by calculating the geometric mean of the lower and upper
chronic limits from a chronic test. A lower chronic limit is the highest tested concentration
1. in an acceptable chronic test,
2. which did not cause the occurrence (which was statistically significantly different
from the control at p = 0.05) of a specified adverse effect, and
3. below which no tested concentration caused such an occurrence.
An upper chronic limit is the lowest tested concentration
1. in an acceptable chronic test,
2. which did cause the occurrence (which was statistically significantly different from
the control at p = 0.05) of a specified adverse effect, and
3. above which all tested concentrations caused such an occurrence.
NOTE: Various authors have used a variety of terms of definitions to interpret the results
of chronic tests, so reported results should be reviewed carefully.
G. If the chronic toxicity of the substance to aquatic animals has been adequately shown to be
related to a water quality characteristic such as hardness for freshwater organisms or
salinity for saltwater organisms, a Final Chronic Equation should be derived based on that
water quality characteristic. Go to section VII.
H. If chronic values are available for eight species as described in section III. B.l or C.l, a
species' mean chronic value should be calculated for each species for which at least one
chronic value is available by calculating the geometric mean of all the chronic values for
the species. The Final Chronic Value should then be obtained using the procedures
described in section IV.I-O. Then go to section VI.M.
I. For each chronic value for which at least one appropriate acute value is available, calculate
an acute-chronic ratio, using for the numerator the arithmetic average of the results of all
standard flow-through acute tests in which the concentrations were measured and which
are from the same study as the chronic test. If such an acute test is not available, use for the
numerator the results of a standard acute test performed at the same laboratory with the
same species, toxicant, and dilution water. If no such acute test is available, use the species
mean acute value for the numerator.
NOTE: If the acute toxicity or chronic toxicity or both of the substance have been
adequately shown to be related to a water quality characteristic, the numerator and the
denominator must be based on tests performed in the same water.
J. For each species, calculate the species mean acute-chronic ratio as the geometric mean of
all the acute-chronic ratios available for that species.
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K. For some substances the species mean acute-chronic ratio seems to be the same for all
species; but for other substances the ratio seems to increase as the species mean acute
value increases. Thus the Final Acute-Chronic Ratio can be obtained in two ways,
depending on the data available.
1. If no major trend is apparent and the acute-chronic ratios for a number of species are
within a factor of 10, the final Acute-Chronic Ratio should be calculated as the
geometric mean of all the species' mean acute-chronic ratios available for both
freshwater and saltwater species.
2. If the species' mean acute-chronic ratio seems to increase as the species' mean acute
value increases, the value of the acute-chronic ratio for species whose acute values are
close to the Final Acute Value should be chosen as the Final Acute-Chronic Ratio.
L. Calculate the Final Chronic Value by dividing the Final Acute Value by the Final
Acute-Chronic Ratio.
M. If the species mean chronic value of an important species, such as a commercially or
recreationally important species, is lower than the Final Chronic Value, then that species'
mean chronic value should be used as the Final Chronic Value.
N. Go to section VIII.
VII. Final Chronic Equation
A. For each species for which comparable chronic toxicity values are available at two or more
different values of a water quality characteristic that apparently affects chronic toxicity,
perform a least squares regression of the natural logarithms of the chronic toxicity values
on the natural logarithms of the water quality characteristic values. No transformation or
a different transformation may be used if it fits the data better, but appropriate changes
will be necessary throughout this section. It is probably preferable, but not necessary, to
use the same transformation that was used with the acute values in section V.
B. Determine whether or not each chronic slope is meaningful, taking into account the range
and number of values of the water quality characteristic tested. For example, a slope based
on four data points may be of limited value if it is based only on data for a narrow range of
values of the water quality characteristic. On the other hand, a slope based on only two
data points may be meaningful if it is consistent with other information and if the two
points cover a broad enough range of the water quality characteristic. If a meaningful
chronic slope is not available for at least one species, return to section VI.H.
C. Calculate the mean chronic slope (L) as the arithmetic average of all the meaningful
chronic slopes for individual species.
D. For each species calculate the geometric mean (M) of the toxicity values and the geometric
mean (P) of the related values of the water quality characteristic.
E. For each species calculate the logarithmic intercept (Q) using the equation:
Q = In M - L(ln P).
F. For each species calculate a species mean chronic intercept as the antilog of Q.
G. Obtain the Final Chronic Intercept by using the procedure described in section IV. I-O,
except insert "Intercept" for "Value."
H. If the species mean chronic intercept of an important species, such as a commercially or
recreationally important species, is lower than the Final Chronic Intercept, then that
species' mean chronic intercept should be used as the Final Chronic Intercept.
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I. The Final Chronic Equation is written as
e (L[ln (water quality characteristic)^ In R)
where
L = mean chronic slope
R = Final Chronic Intercept.
VIII. Final Plant Value
A. Appropriate measures of the toxicity of the substance to aquatic plants are used to
compare the relative sensitivities of aquatic plants and animals.
B. A value is a concentration that decreased growth (as measured by dry weight, chlorophyll,
etc.) in a 96-hour or longer test with an alga or in a chronic test with an aquatic vascular
plant.
C. Obtain the Final Plant Value by selecting the lowest plant value from a test in which the
toxicant concentrations were measured.
IX. Final Residue Value
A. The Final Residue Value is derived in order to
1. Prevent commercially or recreationally important aquatic organisms from exceeding
relevant FDA action levels, and
2. Protect wildlife, including fishes and birds, that eat aquatic organisms from
demonstrated adverse effects.
A residue value is calculated by dividing a maximum permissible tissue concentration
by an appropriate bioconcentration factor (BCF), where the BCF is the quotient of the
concentration of a substance in all or part of an aquatic organism divided by the
concentration in water to which the organism has been exposed. A maximum
permissible tissues concentration is either
1. An action level from the FDA Administrative Guidelines Manual for fish oil or for the
edible portion of fish or shellfish, or
2. A maximum acceptable dietary intake based on observations on survival, growth, or
reproduction in a chronic wildlife feeding study.
If no maximum permissible tissue concentration is available, go to section X because no
Final Residue Value can be derived.
B. Bioconcentration factor
1. A BCF determined in a laboratory test should be used only if it was calculated based
on measured concentrations of the substance in the test solution and was based on an
exposure that continued until either steady-state or 28-days was reached. Steady-state
is reached when the BCF does not change significantly over a period of time, such as
two days or 16 percent of the length of the exposure, whichever is longer. If a
steady-state BCF is not available for a species, the available BCF for the longest
exposure over 28 days should be used for that species.
2. A BCF from a field exposure should be used only when it is known that the
concentration of the substance was reasonably constant for a long enough period of
time over the range of territory inhabited by the organisms.
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3. If BCF values from field exposures are consistently lower or higher than those from
laboratory exposures, then only those values from field exposures should be used if
possible.
4. A BCF should be calculated based on the concentration of the substance and its
metabolites, which are structurally similar and are not much more soluble in water
than the parent compound, in appropriate tissue and should be corrected for the
concentration in the organisms at the beginning of the test.
5. A BCF value obtained from a laboratory or field exposure that caused an observable
adverse effect on the test organism may be used only if it is similar to that obtained
with unaffected organisms at lower concentrations in the same test.
6. Whenever a BCF is determined for a lipid-soluble substance, the percent lipids should
also be determined in the tissue for which the BCF was calculated.
C. A BCF calculated using dry tissue weights must be converted to a wet tissue weight basis
by multiplying the dry weight BCF value by 0.1 for plankton and by 0.2 for individual
species of fishes and invertebrates.
D. If enough pertinent data exist, several residue values can be calculated by dividing
maximum permissible tissue concentrations by appropriate BCF values.
1. For each available maximum acceptable dietary intake derived from a chronic feeding
study with wildlife, including birds and aquatic organisms, the appropriate BCF is
based on the whole body of aquatic species that constitute or represent a major portion
of the diet of the tested wildlife species.
2. For an FDA action level, the appropriate BCF is the highest geometric mean species
BCF for the edible portion (muscle for decapods, muscle with or without skin for
fishes, adductor muscle for scallops and total living tissue for other bivalve molluscs)
of a consumed species. The highest species BCF is used because FDA action levels are
applied on a species-by-species basis.
E. For lipid-soluble substances, it may be possible to calculate additional residue values.
Because steady-state BCF values for a lipid-soluble chemical seem to be proportional to
percent lipids from one tissue to another and from one species to another, extrapolations
can be made from tested tissues or species to untested tissues or species on the basis of
percent lipids.
1. For each BCF where the percent lipids is known for the same tissue which the BCF was
measured, the BCF should be normalized to a 1 percent lipid basis by dividing the BCF
by the percent lipids. This adjustment to a 1 percent lipid basis makes all the measured
BCF values comparable, regardless of the species or tissue for which the BCF is
measured.
2. Calculate the geometric mean normalized BCF. Data for both saltwater and
freshwater species can be used to determine the mean normalized BCF, because the
normalized BCF seems to be about the same for both kinds of organisms.
3. Residue values can then be calculated by dividing the maximum permissible tissue
concentrations by the mean normalized BCF and by a percent lipids value appropriate
to the maximum permissible tissue concentration.
Residue Value - (max/mum permissible tissue concentration)
° (mean normalized BCF) (appropriate percent lipids)
a. For an FDA action level for fish oil, the appropriate percent lipids value is 100.
b. For an FDA action level for fish, the appropriate percent lipids value is 15 for
freshwater criteria and 16 for saltwater criteria because FDA action levels are
applied species-by-species to commonly consumed species. The edible portion of
the freshwater lake trout averages about 15 percent lipids, and the edible portion
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of the saltwater Atlantic herring averages about 16 percent lipids (Sidwell, V.D., et
al. 1974 Composition of the Edible Portion of Raw (Fresh or Frozen) Crustaceans,
Finfish, and Mollusks. I. Protein, Fat, Moisture, Ash, Carbohydrate, Energy Value,
and Cholesterol, Marine Fisheries Review 36:21-35).
c. For a maximum acceptable dietary intake derived from a chronic feeding study
with wildlife, the appropriate percent lipids is the percent lipids of an aquatic
species or group of aquatic species that constitutes a major portion of the diet of
the wildlife species.
F. The Final Residue Value is obtained by selecting the lowest of the available residue values.
In many cases the Final Residue Value will not be low enough. For example, a residue
value calculated from an FDA action level would result in an average concentration in the
edible portion of a fatty species that is at the action level. On the average half, of the
individuals of the species would have concentrations above the FDA action level. Also, the
results of many chronic feeding studies are concentrations that cause adverse effects.
X. Other Data
Pertinent information that could not be used in earlier sections may be available
concerning adverse effects on aquatic organisms and their uses. The most important of
these are data on flavor impairment, reduction in survival, growth, or reproduction, or
any other adverse effect that has been shown to be biologically significant. Especially
important are data for species for which no other data are available. Data from behavioral,
microcosm, field, and physiological studies may also be available.
X. Criterion
A. The criterion consists of two concentrations, one that should not be exceeded on the
average in a 24-hour period and one that should not be exceeded at any time during the
24-hour period. This two-number criterion is intended to identify water quality
conditions that should protect aquatic life and its uses from acute and chronic adverse
effects of both cumulative and noncumulative substances without being as restrictive as a
one-number criterion would have to be to provide the same degree of protection.
B. The maximum concentration is the Final Acute Value or is obtained from the Final Acute
Equation.
C. The 24-hour average concentration is obtained from the Final Chronic Value, the Final
Plant Value, and the Final Residue Value by selecting the lowest available value, unless
other data (see section X) from tests in which the toxicant concentrations were measured
show that a lower value should be used. If toxicity is related to a water quality
characteristic, the 24-hour average concentration is obtained from the Final Chronic
Equation, the Final Plant Value, and the Final Residue Value by selecting the one that
results in the lowest concentrations in the normal range of the water quality characteristic,
unless other data (see section X) from tests in which the toxicant concentrations were
measured show that a lower value should be used.
D. The criterion is (the 24-hour average concentration) as a 24-hour average and the
concentration should not exceed (the maximum concentration) at any time.
XII. Review
A. On the basis of all available pertinent laboratory and field information, determine if the
criterion is consistent with sound scientific evidence. If it is not, another criterion, either
higher or lower, should be derived using appropriate modifications of the Guidelines.
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APPENDIX C
Derivation of the 1980
Human Health Criteria
(Excerptedfrom 45 F.R. 79347, November28,1980)
Guidelines and Methodology Used in the Preparation of Health Effect
Assessment Chapters of the Consent Decree Water Criteria Documents
I. Objective
The objective of the health effect assessment chapters of the ambient water criteria documents
is to estimate ambient water concentrations that do not represent a significant risk to the
public. These assessments should constitute a review of all relevant information on individual
chemicals or chemical classes in order to derive criteria that represent, in the case of suspect or
proven carcinogens, various levels of incremental cancer risk, or, in the case of other
pollutants, estimates of no-effect levels.
Ideally, ambient water quality criteria should represent levels for compounds in ambient
water that do not pose a hazard to the human population. However, in any realistic
assessment of human health hazard, a fundamental distinction must be made between
absolute safety and recognizing some risk. Criteria for absolute safety would have to be based
on detailed knowledge of dose-response relationships in humans, including all sources of
chemical exposure, the types of toxic effects elicited, the existence of thresholds for the toxic
effects, the significance of toxicant interactions, and the variances of sensitivities and exposure
levels within the human population. In practice, such absolute criteria cannot be established
because of deficiencies in both the available data and the means of interpreting this
information. Consequently, the individual human health effects chapters propose criteria that
minimize or specify the potential risk of adverse human effects due to substances in ambient
water. Potential social or economic costs and benefits are not considered in the formulation of
the criteria.
II. Types of Criteria
Ambient water quality criteria are based on three types of biological endpoints:
carcinogenicity, toxicity (i.e., all adverse effects other than cancer), and organoleptic effects.
For the purpose of deriving ambient water quality criteria, carcinogenicity is regarded as a
non-threshold phenomenon. Using this assumption, "safe" or "no effect" levels for
carcinogens cannot be established because even extremely small doses must be assumed to
elicit a finite increase in the incidence of the response. Consequently, water quality criteria for
carcinogens are presented as a range of pollutant concentrations associated with
corresponding incremental risks.
For compounds that do not manifest any apparent carcinogenic effect, the threshold
assumption is used in deriving a criterion. This assumption is based on the premise that a
physiological reserve capacity exists within the organism that is thought to be depleted before
clinical disease ensues. Alternatively, the rate of damage will likely be insignificant over the
life span of the organism. Thus, ambient water quality criteria are derived for
non-carcinogenic chemicals, and presumably result in no observable-adverse-affect levels
(NOAELs) in the exposed human population.
In some instances, criteria are based on organoleptic characteristics (i.e., thresholds for
taste or odor). Such criteria are established when insufficient information is available on
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toxicologic effects or when the estimate of the level of the pollutant in ambient water based on
organoleptic effects is lower than the level calculated from toxicologic data. Criteria based
solely on organoleptic effects do not necessarily represent approximations of acceptable risk
levels for human health.
Several ambient water quality criteria documents deal with classes of compounds that
include chemicals exhibiting varying degrees of structural similarity. Because prediction of
biological effects based solely on structural parameters is difficult, the derivation of
compound-specific criteria is preferable to a class criterion. A compound-specific criterion is
defined as a level derived from data on each individual subject compound that does not
represent a significant risk to the public. For some chemical classes, however, a
compound-specific criterion cannot be derived for each member of a class. In such instances, it
is sometimes justifiable to derive a class criterion in which available data on one member of a
class may be used to estimate criteria for other chemicals of the class because a sufficient data
base is not available for those compounds.
For some chemicals and chemical classes, the data base was judged to be insufficient for the
derivation of a criterion. In those cases, deficiencies in the available information are detailed.
Approach
The human health effects chapters attempt to summarize all information on the individual
chemicals or classes of chemicals that might be useful in the risk assessment process to develop
water quality criteria. Although primary emphasis is placed on identifying epidemiologic and
toxicologic data, these assessments typically contain discussions on four topics: existing levels
of human exposure, pharmacokinetics, toxic effects, and criterion formulation.
For all documents, an attempt is made to include the known relevant information. Review
articles and reports are often used in the process of data evaluation and synthesis. Scientific
judgment is exercised in the review and evaluation of the data in each document and in the
identification of the adverse effects against which protective criteria are sought. In addition,
each of these documents is reviewed by a peer committee of scientists familiar with the
specific compound(s). These work groups evaluate the quality of the available data, the
completeness of the data summary, and the validity of the derived criterion.
In the analysis and organization of the data, an attempt is made to be consistent in the
format and application of acceptable scientific principles. Evaluation procedures used in the
hazard assessment process follow the principles outlined by the National Academy of
Sciences in Drinking Water and Health (1977) and the guidelines of the Carcinogen Assessment
Group of the U.S. EPA.
A. Exposure
The exposure section of the health effects chapters reviews known information on
current levels of human exposure to the individual pollutant from all sources. Much of the
data was obtained from monitoring studies of air, water, food, soil, and human or animal
tissue residues. The major purpose of this section is to provide background information
on the contribution of water exposure relative to all other sources. Consequently, the
exposure section includes subsections reviewing different routes of exposure, including
water and food ingestion, inhalation, and dermal contact.
Information on exposure can be valuable in developing and assessing a water quality
criterion. In these documents, exposure from consumption of contaminated water and
contaminated fish and shellfish products is used in criterion formulation. Data for all
modes of exposure are useful in relating total intake to the expected contribution from
contaminated water, fish, and shellfish. In addition, information for all routes of exposure,
not limited to drinking water and fish and shellfish ingestion, can be used to justify or
assess the feasibility of the formulation of criteria for ambient water.
The use of fish consumption as an exposure factor requires the quantitation of
pollutant residues in the edible portions of the ingested species. Accordingly,
bioconcentration factors (BCFs) are used to relate pollutant residues in aquatic organisms
to the pollutant concentration in the ambient waters in which they reside.
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To estimate the average per capita intake of a pollutant due to consumption of
contaminated fish and shellfish, the results of a diet survey were analyzed to calculate the
average consumption of freshwater and estuarine fish and shellfish. A species is
considered to be a consumed freshwater or estuarine fish and shellfish species if at some
stage in its life cycle it is harvested from fresh or estuarine water for human consumption
in significant quantities.
Three different procedures are used to estimate the weighted average BCF, depending
upon the lipid solubility of the chemical and the availability of bioconcentration data.
For lipid-soluble compounds, the average BCF is calculated from the weighted
average percent lipids in the edible portions of consumed freshwater and estuarine fish
and shellfish, which was calculated from data on consumption of each species and its
corresponding percent lipids to be 3 percent. Because the steady-state BCFs for
lipid-soluble compounds are proportional to percent lipids, bioconcentration factors for
fish and shellfish can be adjusted to the average percent lipids for aquatic organisms
consumed by Americans. For many lipid-soluble pollutants, there exists at least one BCF
for which the percent lipid value was measured for the tissues for which the BCF is
determined.
With 3 percent as the weighted average percent lipids for freshwater and estuarine
fish and shellfish in the average diets, a BCF, and a corresponding percent lipid value, the
weighted average bioconcentration factor can be calculated.
Example:
Weighted average percent lipids for average diet = 3.0 percent
Measured BCF of 17 for trichloroethylene with bluegills at 4.8 percent lipids
Weighted average BCF for average diet equals
3.0% ...
X 4.8% =
As an estimate, 10.6 is used for the BCF.
In those cases where an appropriate bioconcentration factor is not available, the
equation "Log BCF = (0.85 Log P)- 0.70" can be used to estimate the BCF for aquatic
organisms containing about 7.6 percent lipids from the octanol / water partition coefficient
P. An adjustment for percent lipids in the average diet versus 7.6 percent is made in order
to derive the weighted average bioconcentration factor.
For non-lipid-soluble compounds, the available BCFs for the edible portion of
consumed freshwater and estuarine fish and shellfish are weighted according to
consumption factors to determine a weighted BCF representative of the average diet.
B. Pharmacokinetics
This section summarizes the available information on the absorption, distribution,
metabolism, and elimination of the compound(s) in humans and experimental mammals.
Conceptually, such information is useful in validation of inter- and intraspecies
extrapolations, and in characterizing the modes of toxic actions. Sufficient information on
the absorption and excretion in animals — together with a knowledge of ambient
concentrations in water, food, and air — could be useful in estimating body burdens of
chemicals in the human population. Distribution data that suggest target organs or tissues
are desirable for interspecies comparison techniques. For derivation of criteria,
pharmacokinetic data are essential to estimate equivalent oral doses based on data from
inhalation or other routes of exposure.
C. Effects
This section summarizes information on biological effects in both humans and
experimental mammals resulting in acute, subacute, and chronic toxicity; synergism
and/or antagonism; teratogenicity, mutagenicity, or carcinogenicity.
The major goal of this section is to survey the suitability of the data for use in
assessment of hazard and to determine which biological end-point (i.e., non-threshold,
threshold, or organoleptic) should be selected for use in criterion formulation.
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Because this section attempts to assess potential human health effects, data on
documented human effects are thoroughly evaluated. However, several factors inherent
in human epidemiological studies usually preclude the use of such data in generating
water quality criteria. These problems, as summarized by the National Academy of
Sciences, are as follows:
1. Epidemiology cannot tell what effects a material will have until after humans have
been exposed. One must not conduct what might be hazardous experiments on man.
2. If exposure has been ubiquitous, assessing the effects of a material may be impossible
because no control group is unexposed. Statistics of morbidity obtained before using a
new material can sometimes be useful; but when latent periods are variable and times
of introduction and removal of materials overlap, historical data on chronic effects are
usually unsatisfactory.
3. Determining doses in human exposure is usually difficult.
4. Usually, it is hard to identify small changes in common effects, which may nonetheless
be important if the population is large.
5. Interactions in a "nature-designed" experiment usually cannot be controlled.
Although these problems often prevent the use of epidemiological data in
quantitative risk assessments, qualitative similarities or differences between documented
effects in humans and observed effects in experimental mammals are extremely useful in
testing the validity of animal-to-man extrapolations. Consequently, in each case, an
attempt is made to identify and utilize both epidemiological and animal dose-response
data. Criteria derived from such a confirmed data base are considered to be reliable.
The decision to establish a criterion based on a non-threshold model is made after
evaluating all available information on carcinogenicity and supportive information on
mutagenicity. The approach and conditions for the qualitative decision of carcinogenicity
are outlined in the U.S. EPA Interim Cancer Guidelines (41FR 21402), in a report by Albert,
et al. (1977), and in the Interagency Regulatory Liaison Group (IRLG) guidelines on
carcinogenic risks (IRLG, 1979). A substance that induces a statistically significant
carcinogenic response in animals is assumed to have the capacity to cause cancer in
humans. A chemical that has not induced a significant cancer response in humans or
experimental animals is not identified as a carcinogen, even though its metabolites or
close structural analogues might induce a carcinogenic response or it was shown to be
mutagenic in an in vitro system.
Some potential human carcinogens may not be identified by the guidelines given
above. For example, compounds that have plausible but weak qualitative evidence of
carcinogenicity in experimental animal systems (such as data from mouse skin painting or
strain A mouse pulmonary adenoma) would be included in this category. The derivation
of a criterion for human consumption from these studies in not valid, regardless of the
qualitative outcome. In addition, certain compounds (e.g., nickel and beryllium) shown to
be carcinogenic in humans after inhalation exposure by chemical form have induced, thus
far, no response in animals or humans via ingesting their soluble salts.
Nevertheless, nonthreshold criterion is developed for beryllium because tumors have
been produced in animals at a site removed from the site of administration. In contrast, a
threshold criterion is recommended for nickel because no evidence exists of tumors at
distant sites resulting from administration of nickel solutions by either ingestion or
injection.
For those compounds not reported to induce carcinogenic effects or for which
carcinogenic data are lacking or insufficient, an attempt is made to estimate a no-effect
level. In many respects, the hazard evaluation from these studies is similar to that of
bioassays for carcinogenicity. In order to more closely approximate conditions of human
exposure, preference is given to chronic studies involving oral exposures in water or diet
over a significant portion of the animal life span. Greatest confidence is placed in those
studies that demonstrate dose-related adverse effects as well as no-effect levels.
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The biological endpoints used to define a no-effect level vary considerably. They may
range from gross effects — such as mortality — to more subtle biochemical, physiological,
or pathological changes. Teratogenicity, reproductive impairment, and behavioral effects
are significant toxic consequences of environmental contamination. In instances where
carcinogenic or other chronic effects occur at exposure levels below those causing
teratogenicity, reproductive impairment, or behavioral effects, the former are used in
deriving the criterion. For most of the compounds evaluated thus far, teratogenicity and
reproductive impairment occur at doses near maximum tolerated levels with dose
administration schedules well above estimated environmental exposure levels. Moreover,
information on behavioral effects, which could be of significance, is not available for most
of the compounds under study. Consequently, most NOAELs derived from chronic
studies are based either on gross toxic effects or on effects directly related to functional
impairment or defined pathological lesions.
For compounds on which adequate chronic toxicity studies are not available, studies
on acute and subacute toxicity assume greater significance. Acute toxicity studies usually
involve single exposures at lethal or near lethal doses. Subacute studies often involve
exposure exceeding 10 percent of the life span of the test organism, (e.g., 90 days for the rat
with an average life span of 30 months). Such studies are useful in establishing the nature
of the compound's toxic effects and other parameters of compound toxicity, such as target
organ effects, metabolic behavior, physiological/biochemical effects, and patterns of
retention and tissue distribution. The utility of acute and subacute studies in deriving
environmentally meaningful NOELs is uncertain, although McNamara (1976) has
developed application factors for such derivations.
In some cases where adequate data are not available from studies utilizing oral routes
of administration, no-effect levels for oral exposures may be estimated from dermal or
inhalation studies. Such estimates involve approximations of the total dose administered
based on assumptions about breathing rates and/or magnitude of absorption.
D. Criterion Rationale
This section reviews existing standards for the chemical(s), summarizes data on
current levels of human exposure, attempts to identify special groups at risk, and defines
the basis for the recommended criterion.
Information on existing standards is included primarily for comparison with the
proposed water quality criteria. Some of the present standards, such as those
recommended by the Occupational Safety and Health Administration (OSHA) or the
American Conference of Governmental Industrial Hygienists (ACGIH), are based on
toxicologic data but are intended as acceptable levels for occupational rather than
environmental exposure. Other levels, such as those recommended by the National
Academy of Sciences in Drinking Water and Health (1977) or in the U.S. EPA Interim
Primary Drinking Water Standards, are more closely related to proposed water quality
criteria. Emphasis is placed on detailing the basis for the existing standards wherever
possible.
Summaries of current levels of human exposure, presented in this section, specifically
address the suitability of the data to derive water quality criteria. The identification of
special groups at risk, either because of geographical or occupational differences in
exposure or biological differences in susceptibility to the compound(s), focuses on the
impact that these groups should have on the development of water quality criteria.
The basis for the recommended criteria section summarizes and qualifies all of the
data used in developing the criteria.
Guidelines for Criteria Derivation
The derivation of water quality criteria from laboratory animal toxicity data is essentially a
two-step procedure. First, a total daily intake for humans must be estimated that establishes
either a defined level of risk for nonthreshold effects or a noeffect level for threshold effects.
Secondly, assumptions must be made about the contribution of contaminated water and the
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consumption of fish/shellfish to the total daily intake of the chemical. These estimates are
then used to establish the tolerable daily intake and consequently the water quality criterion.
A. Nonthreshold Effects
After the decision has been made that a compound can potentially cause cancers in
humans and that data exist that permit the derivation of a criterion, the water
concentration estimated to cause a lifetime carcinogenic risk of 10"5 is determined. The
lifetime carcinogenicity risk is the probability that a person would get cancer sometime in
his or her life assuming continuous exposure to the compound. The water concentration is
calculated by using the low-dose extrapolation procedure proposed by Crump (1980).
This procedure is an improvement on the multistage low-dose extrapolation procedure by
Crump, et al. (1977).
The data used for quantitative estimates are of two types: (1) lifetime animal studies,
and (2) human studies where excess cancer risk has been associated with exposure to the
agent. In animal studies it is assumed, unless evidence exists to the contrary, that if a
carcinogenic response occurs at the dose levels used in the study, then proportionately
lower responses will also occur at all lower doses, with an incidence determined by the
extrapolation model discussed below.
1. Choice of Model.
No really solid scientific basis exists for any mathematical extrapolation model
that relates carcinogen exposure to cancer risks at the extremely low levels of
concentration present to deal with in evaluating the environmental hazards. For
practical reasons, such low levels of risk cannot be measured directly either using
animal experiments or epidemiologic studies. We must, therefore, depend on our
current understanding of the mechanisms of carcinogenesis for guidance as to
which risk model to use.
At the present time, the dominant view of the carcinogenic process involves the
concept that most agents which cause cancer also cause irreversible damage to
DNA. This position is reflected by the fact that a very large proportion of agents that
cause cancer are also mutagenic. The quantal type of biological response
characteristic of mutagenesis is likely associated with a linear nonthreshold
dose-response relationship. Indeed, substantial evidence from mutagenesis studies
with both ionizing radiation and a wide variety of chemicals suggests that this type
of dose-response model is the appropriate one to use. This is particularly true at the
lower end of the dose-response curve; at higher doses, there can be an upward
curvature, probably reflecting the effects of multistage processes on the mutagenic
response.
The linear nonthreshold dose-response relationship is also consistent with the
relatively few epidemiological studies of cancer responses to specific agents that
contain enough information to make the evaluation possible (e.g.,
radiation-induced leukenlia, breast and thyroid cancer, skin cancer induced by
arsenic in drinking water, and liver cancer induced by aflatoxin in the diet). Some
animal experiments are consistent with the linear nonthreshold hypothesis (e.g.,
liver tumors induced in mice by 2-acetylaminofluorene in the large scale ED01.
study at the National Center of Toxicological Research, and the initiation stage of
the two-stage carcinogenesis model in the rat liver and the mouse skin).
Because it has the best, albeit limited, scientific basis of any of the current
mathematical extrapolation models, the linear nonthreshold model has been
adopted as the primary basis for risk extrapolation to low levels of the
dose-response relationship. The risk assessments made with this model should be
regarded as conservative, representing the most plausible upper limit for the risk
(i.e., the true risk is not likely to be higher than the estimate, but it could be smaller).
The mathematical formulation chosen to describe the linear, nonthreshold
dose-response relationship at low doses is the improved multistage model
developed by Crump (1980). This model employs enough arbitrary constants to be
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able to fit almost any monotonically increasing dose-response data, and it
incorporates a procedure for estimating the largest possible linear slope (in the 95
percent confidence limit sense) at low extrapolated doses consistent with the data at
all dose levels of the experiment. For this reason, it may be called a "linearized"
multistage model.
2. Procedure of Low-Dose Extrapolation Based on Animal Carcinogenicity Data.
A. Description of the Extrapolation Model
Let P(d) represent the lifetime risk (probability) of cancer at dose d. The
multistage model has the form
P(d) = 1 - exp [- (q0+ q2d + q2d3+...+qkdk)]
where:
qi> O, and i = 0,1,2,.... k
Equivalently,
A(d) = 1 - exp [- (qjd + c^d2 +...+ q,^)]
where:
A(d) = p(o)°^ extra over background rate at dose d.
The point estimate of the coefficients qj i = 0,1,2,...,k, and consequently the extra
risk function A(d) at any given dose d, is calculated by maximizing the likelihood
function of the data.
The point estimate and the 95 percent upper confidence limit of the extra risk
A(d) are calculated by using the computer program GLOBAL 79 developed by
Crump and Watson (1979). Upper 95 percent confidence limits on the extra risk and
lower 95 percent confidence limits on the dose producing a given risk are
determined from a 95 percent upper confidence limit q2*, on parameter q^
Whenever qj * 0, at low doses extra risk A(d) has approximately the form A(d) = q2
x d. Therefore, q2 x d is a 95 percent upper confidence limit on the extra risk and
R/qj* is a 95 percent lower confidence limit on the dose producing an extra risk of R.
Let L0 be the maximum value of the log-likelihood function. The upper limit q^ is
calculated buy increasing qj to a value q:* such that when the log-likelihood is again
maximized subject to this fixed value q-j* for the linear coefficient, the resulting
maximum value of the log-likelihood L] satisfies the equation 2(L0 - Lj) = 2.70554
where 2.70554 is the cumulative 90 percent point of the chi-square distribution with
one degree of freedom, which corresponds to a 95 percent upper limit (one-sided).
This approach of computing the upper confidence limit for the extra risk A(d) is an
improvement on the Crump, et al. (1977) model. The upper confidence limit for the
extra risk calculated at low doses is always linear. This is conceptually consistent
with the linear nonthreshold concept discussed earlier. The slope q:* is taken as an
upper bound of the potency of the chemical in inducing cancer at low doses.
In fitting the dose-response model, the number of terms in the polynomial g is
chosen equal to (h-1), where h is the number of dose groups in the experiment,
including the control group.
Whenever the multistage model does not fit the data sufficiently, data at the
highest dose is deleted and the model is refitted to the rest of the data. This is
continued until an acceptable fit to the data is obtained. To determine whether or
not a fit is acceptable, the chi-square statistic:
, (Xj-NjP;)2
~i.i NiP|(l-P,)
is calculated, where N-, is the number of animals in the ith dose group, Xj is the
number of animals in the ith dose group with a tumor response, Pj is the probability
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of a response in the ith dose group estimated by fitting the multistage model to the
data, and h is the number of remaining groups.
The fit is determined to be unacceptable whenever chi-square (X2) is larger than
the cumulative 99 percent point of the chi-square distribution with f degrees of
freedom, where f equals the number of dose groups minus the number of non-zero
multistage coefficients.
Selection and Form of Data used to Estimate Parameters in the Extrapolation Model.
For some chemicals, several studies in different animal species, strains, and
sexes each conducted at several doses and different routes of exposure are available.
A choice must be made as to which of the data sets from several studies are to be
used in the model. It is also necessary to correct for metabolism differences between
species and for differences in absorption via different routes of administration. The
procedures used in evaluating these data, listed below, are consistent with the
estimate of a maximum-likely-risk.
a. The tumor incidence data are separated according to organ sites or tumor types.
The set data (i.e., dose and tumor incidence) used in the model is set where the
incidence is statistically significantly higher than the control for at least one test
dose level and/or where the tumor incidence rate shows a statistically significant
trend with respect to dose level. The data set that gives the highest estimate of
lifetime carcinogenic risk qj* is selected in most cases. However, efforts are made
to exclude data sets that produce spuriously high risk estimates because of a small
number of animals. That is, if two sets of data show a similar dose-response
relationship and one has a very small sample size, the set of data that has the larger
sample size is selected for calculating the carcinogenic potency.
b. If there are two more data sets of comparable size that are identical with respect to
species, strain, sex, and tumor sites, the geometric mean of q^, estimated from
each of these data sets, is used for risk assessment. The geometric mean of
numbers Aj, A2,..., Am is defined as (Aj x A2 x... Am): /m
c. If sufficient data exist for two or more significant tumor sites in the same study, the
number of animals with at least one of the specific tumor sites under consideration
is used as incidence data in the model.
d. Following the suggestion of Mantel and Schneiderman (1975), we assume that
mg/surface area/day is an equivalent dose between species. Since to a close
approximation the surface area is proportional to the %rds power of the weight as
would be the case for a perfect sphere, the exposure in mg/%rds power of the
body weight/day is similarly considered to be an equivalent exposure. In an
animal experiment, this equivalent dose is computed in the following manner:
Let:
Le = duration of experiment
le= duration of exposure
m = average dose per day in mg during administration of the agent
(i.e., during le)
W = average weight of the experimental animal.
Then, the lifetime average exposure is d =
Often exposures are not given in units of mg/ day, and it becomes necessary to
convert the given exposures to mg/day. For example, in most feeding studies,
exposure is expressed as ppm in the diet. In this case the exposure (mg/day) is
derived by m = ppm iFir, where ppm is parts per million of the carcinogenic
agent in the diet, F is the weight of the food consumed per day in kgms, and r is the
absorption fraction.
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In the absence of any data to the contrary, r is assumed to be one. For a uniform
diet, the weight of the food consumed is proportioned to the calories required,
which, in turn, is proportional to the surface area or the ^Srds power of the weight,
so that mappm x x r or
m a VPm
r
As a result, ppm in the diet is often assumed to be an equivalent exposure
between species. However, this is not justified since the calories/kg of food is
significantly different in the diet of man vs. laboratory animals, primarily due to
moisture content differences. Instead, an empirically derived food factor, f=F/W,
is the fraction of a species body weight that is consumed per day as food. The rates
are
Species . . . .
W
f
Man
70
0.028
Rat
0.35
0.06
Mice
0.03
0.13
Thus, when the exposure is given as a certain dietary concentration in ppm, the
exposure in mg/W ^is
m ppm x F
rxW^ =
rpmxficW _ fxW»m
w*
When exposure is given in terms of mg/kg/day = m/Wr = s the conversion is
simply:
¦ s x
rW*
When exposure is via inhalation, the calculation of dose can be considered for
two cases where (1) the carcinogenic agent is either a completely water-soluble gas
or an aerosol and is absorbed proportionally to the amount of air breathed in, and
(2) where the carcinogen is a poorly water-soluble gas that reaches an equilibrium
between the air breathed and the body compartments. After equilibrium is
reached, the rate of absorption of these agents is expected to be proportional to
metabolic rate, which in turn is proportional to the rate of oxygen consumption,
which in turn is a function of surface area.
B. Threshold Effects
1. Use of Animal Toxicity Data (oral).
In developing guidelines for deriving criteria based on noncarcinogenic
responses, five types of response levels are considered:
NOEL—No Observed-Effect-Level
NO A EL—N o Observed-A d verse-Effect-Level
LOEL—Lowest-Observed-Effect-Level
LOAEL—Lowest-Observed-Adverse-Effect-Level
FEL—Frank-Effect-Level
Adverse effects are defined as any effects that result in functional impairment
and/or pathological lesions which may affect the performance of the whole
organism or reduce an organism's ability to respond to an additional challenge.
One of the major problems encountered in considering these concepts regards
the reporting of "observed effect levels" as contrasted to "observed adverse effect
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levels." The terms "adverse" versus "not adverse" are at times satisfactorily
defined. But due to increasingly sophisticated testing protocols, more subtle
responses are being identified, resulting in a need for judgment regarding the exact
definition of adversity.
The concepts listed above (NOEL, NOAEL, LOEL, LOAEL) have received much
attention because they represent landmarks that help define the threshold region in
specific experiments. Thus, if a single experiment yields a NOEL, a NOAEL, a
LOAEL, and a clearly defined FEL in closely spaced doses, the threshold region has
been well defined; such data are very useful for the purpose of deriving a criterion.
On the other hand, a clearly defined FEL has little utility in establishing criteria
when it stands alone, because such a level gives no indication how far removed the
data point is from the threshold region. Similarly, a free-standing NOEL has little
utility without an indication of its proximity to the LOEL, since a free-standing
NOEL may be many orders of magnitude below the threshold region.
Based on the above dose-response classification system, the following
guidelines for deriving criteria have been adopted:
a. A free-standing FEL is unsuitable for the derivation of criteria.
b. A free-standing NOEL is unsuitable for the derivation of criteria. If multiple
NOELs are available without additional data on LOELs, NOAELs, or LO AELs, the
highest NOEL should be used to derive a criterion.
c. A NOAEL, LOEL, OR LOAEL can be suitable for criteria derivation. A
well-defined NOAEL from a chronic (at least 90-day study) may be used directly,
applying the appropriate uncertainty factor. For a LOEL, a judgment needs to be
made as to whether it actually corresponds to a NOAEL or a LOAEL. In the case of
a LOAEL, an additional uncertainty factor is applied; the magnitude of the
additional uncertainty factor is judgmental and should lie in the range of 1 to 10.
Caution must be exercised not to substitute "Frank-Effect-Levels" for
"Lowest-Observable-Adverse-Effect-Levels."
d. If for reasonably closely spaced doses only a NOEL and a LOAEL of equal quality
are available, then the appropriate uncertainty factor is applied to the NOEL.
In using this approach, the selection and justification of uncertainty factors are
critical. The basis definition and guidelines for using uncertainty factors has been
given by the National Academy of Sciences (1977). "Safety Factor" or "Uncertainty
Factor" is defined as a number that reflects the degree or amount of uncertainty that
must be considered when experimental data in animals are extrapolated to man.
When the quality and quantity of experimental data are satisfactory, a low
uncertainty factor is used: when data is judged to be inadequate or equivocal, a larger
uncertainty factor is used. The following general guidelines have been adopted in
establishing the uncertainty factors:
a. Valid experiment results from studies on prolonged ingestion by man, with no
indication of carcinogenicity. Uncertainty Factor = 10
b. Experimental results of studies of human ingestion not available or scanty (e.g.,
acute exposure only) with valid results of long-term feeding studies on
experimental animals, or in the absence of human studies, valid animal studies on
one or more species. No indication of carcinogenicity. Uncertainty Factor = 100
c. No long-term or acute human data. Scanty results on experimental animals with
no indication of carcinogenicity. Uncertainty Factor = 1,000
Considerable judgment must be used in selecting the appropriate safety factors
for deriving a criterion. In those cases where the data do not completely fulfill the
conditions for one category and appear to be intermediate between two categories,
an intermediate uncertainty factor is used. Such as intermediate uncertainty factor
may be developed based on a logarithmic scale (e.g., 33 being halfway between 10
and 100 on a logarithmic scale).
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In determining the appropriate use of the uncertainty factors, the phrase "no
indication of carcinogenicity" is interpreted as the absence of carcinogenicity data
from animal experimental studies or human epidemiology. Available short-term
carcinogenicity screening tests are reported in the criteria documents, but they are
used neither for derivation of numerical criteria nor to rule out the uncertainty
factor approach.
Because of the high degree of judgment involved in the selection of a safety
factor, the criterion derivation section of each document should provide a detailed
discussion and justification for both the selection of the safety factor and the data to
which it is applied. This discussion should reflect a critical review of the available
data base. Factors to be considered include number of animals, species, and
parameters tested; quality of controls; dose levels; route; and dosing schedules. An
effort should be made to differentiate between results that constitute a
toxicologically sufficient data base and data which maybe spurious in nature.
2. Use of Acceptable Daily Intake (ADI).
For carcinogens, the assumption of low-dose linearity precludes the necessity
for defining total exposure in the estimation of increased incremental risk. For
non-carcinogens, ADIs and criteria derived therefrom are calculated from total
exposure data that include contributions from the diet and air. The equation used to
derive the criterion (C) is:
C = ADI - (DT+IN) / [21 + (0.0065 kg x R)]
where
21 = assumed daily water consumption
0.00065 kg = assumed daily fish consumption
R = bioconcentration factor in units of 1 / kg
DT = estimated non-fish dietary intake
IN = estimated daily intake by inhalation.
If estimates of IN and DT cannot be provided from experimental data, an
assumption must be made concerning total exposure. The inability to estimate DT
and IN due to either lack of data or the wide variability in DT and IN in different
states may add an additional element of uncertainty to the criterion formulation
process. To achieve scientific validity, the accurate estimate of the Acceptable Daily
Intake is the major factor in satisfactory derivation of water quality criteria.
3. Use of Threshold Limit Values or Animal Inhalation Studies.
Threshold Limit Values (TLVs) are established by the American Conference of
Governmental and Industrial Hygienists (ACGIH) and represent 8-hour
time-weighted average concentrations in air that are intended to protect workers
from various adverse health effects over a normal working lifetime. Similar values
are set by NIOSH (criteria) and OSHA (standards) for 10- and 8-hour exposures,
respectively. To the extent that these values are based on sound toxicologic
assessments and have been protective in the work environment, they provide useful
information for deriving or evaluating water quality criteria. However, each TLV
must be carefully examined to determine if the basis of the TLV contains data that
can be used directly to derive a water quality criterion using the uncertainty factor
approach. In addition, the history of each TLV must be examined to assess the extent
to which it has assured worker safety. In each case, the types of effects against which
TLVs are designed to protect are examined relative to exposure from water. It must
be demonstrated that the chemical is not a localized irritant and that the effect at the
site of entry is not significant irrespective of the routes of exposure (i.e., oral or
inhalation).
If the TLV or similar value is recommended as the basis of the criterion,
consideration of the previous points is explicitly stated in the document's criterion
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derivation section. Particular emphasis is placed on the quality of the TLV relative to
the available toxicity data that normally is given priority over TLVs or similar
established values. If the TLV can be justified as the basis for the criterion, then the
problems associated with estimating acceptable oral doses from inhalation data
must be addressed.
Estimating equivalencies of dose-response relationships from one route of
exposure to another introduces an additional element of uncertainty in the criteria
derivation. Consequently, whenever possible, ambient water quality criteria should
be based on data involving oral exposures. If oral data are insufficient, data from
other routes of exposure may be useful in the criterion derivation process.
Inhalation data, including TLVs or similar values, are the most common
alternatives to oral data. Estimates of equivalent doses can be based upon (1)
available pharmacokinetic data for oral and inhalation routes, (2) measurements of
absorption efficiency from ingested or inhaled chemicals, or (3) comparative
excretion data when the associated metabolic pathways are equivalent to those
following oral ingestion or inhalation. Given that sufficient pharmacokinetic data
are available, the use of accepted pharmacokinetic models provides the most
satisfactory approach for dose conversions. However, if available pharmacokinetic
data are marginal or of questionable quality, pharmacokinetic modeling is
inappropriate.
The Stokinger and Woodward (1958) approach, or similar models based on
assumptions of breathing rate and absorption efficiency, represents possible
alternatives when data are not sufficient to justify pharmacokinetic modeling. Such
alternative approaches, however, provide less satisfactory approximations because
they are not based on pharmacokinetic data. Consequently, in using the Stokinger
and Woodward or related models, the basis and uncertainties inherent in each
assumption must be clearly stated in the derivation of the criterion.
The use of data pertaining to other routes of exposure to derive water quality
criteria may also be considered. As with inhalation data, an attempt is made to use
accepted toxicologic and pharmacokinetic principles to estimate equivalent oral
doses. If simplifying assumptions are used, their bases and limitations must be
clearly specified.
Because of the uncertainties involved in extrapolating from one route of
exposure to another and the consequent limitations that this may place on the
derived criterion, the decision to disallow such extrapolation and recommend no
criterion is highly judgmental and must be made on a case-by-case basis. A decision
for or against criteria derivation must balance the quantity and quality of the
available data against a perceived risk to the human population.
If the Stokinger and Woodward (1958) approach is used to calculate an ADI
from a TLV, the general equation is
ADI = TLVxBRxDExdx Aa / (A0xSF)
where:
ADI = Acceptable daily intake in mg
TLV = Concentration in air in mg/m3
DE = Duration of exposure in hours per day
d = 5 days/7 days
Aa = Efficiency of absorption from air
A0 = Efficiency of absorption from oral exposure
SF = Safety factor following guidelines given above
BR = Amount of air breathed per day; assume 10 m
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For deriving an ADI from animal toxicity data, the equation is
ADI = CA x De x d x Aa x BR x 70 kg/ (BWA xAcx SF)
where:
ADI = Acceptable daily intake in mg
CA = Concentration in air ir» mg/m
De = Duration of exposure in hours per day
d = Number of days exposed / number of days observed
Aa = Efficiency of absorption from air
BR = Volume of air breathed per day in m
70 kg = Assumed human body weight
BWa = Body weight of experimental animals in kg
A0 = Efficiency of absorption from oral exposure
SF = Safety factor following guidelines given above
More formal pharmacokinetic models must be developed on a compound-
by-compound basis.
Safety factors used in the above formulae are intended to account for species
variability. Consequently, the mg surface area / day conversion factor is not used in
the derivation of toxicity based criterion.
C. Organoleptic Criteria
Organoleptic criteria define concentrations of materials that impart undesirable taste
and/or odor to water. In developing and utilizing such criteria two factors must be
appreciated: the limitations of most organoleptic data and the human health significance
of organoleptic properties.
The publications that report taste and odor thresholds are, with few exceptions,
cryptic in their descriptions of test methodologies, number of subjects tested,
concentration/response relationships, and sensory characteristics at specific
concentrations above threshold. Thus, the quality of organoleptic data is often
significantly less than that of toxicologic data used in establishing other criteria.
Consequently, a critical evaluation of the available organoleptic data must be made and
selection of the most appropriate data base for the criterion based on sound scientific
judgment.
Organoleptic criteria are not based on toxicologic information and have no direct
relationship to potential adverse human health effects. Sufficiently intense organoleptic
characteristics could result in depressed fluid intake that, in turn, might aggravate a
variety of functional disease states (i.e., kidney and circulatory diseases). However, such
effects are not used in the derivation process or organoleptic criteria unless available data
would indicate an indirect human health effect via decreased fluid consumption; criteria
derived solely from organoleptic data are based upon aesthetic qualities only.
Since organoleptic and human health effects criteria are based on different endpoints,
a distinction must be made between these two sets of information. In criteria summaries
involving both types of data, the following format is used:
For comparison purposes, two approaches were used to derive criterion levels
for . Based on available toxicity data, for the protection of public
health the derived level is . Using available organoleptic data, for
controlling undesirable taste and odor quality of ambient water the estimated
level is . It should be recognized that organoleptic data as a basis for
establishing a water quality criteria have no demonstrated relationship to
potential adverse human health effects.
In those instances where a level to limit toxicity cannot be derived, the
following statement is to be appropriately inserted:
Sufficiently data are not available for to derive a level that would
protect against the potential toxicity of this compound.
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D. Criteria for Chemical Classes
A chemical class is broadly defined as any group of chemical compounds that are
reviewed in a single risk assessment document. In criterion derivation, isomers should be
regarded as a part of a chemical class rather than as a single compound. A class criterion is
an estimate of risk/ safety that applies to more than one member of a class. It uses available
data on one or more chemicals of a class to derive criteria for other compounds of the same
class in the event that insufficient data are available to derive compound-specific criteria.
A class criterion usually applies to each member of a class rather than to the sum of the
compounds within the class. While the potential hazards of multiple toxicant exposure are
not to be minimized, a criterion, by definition, most often applies to an individual
compound. Exceptions may be made for complex mixtures that are produced, released,
and toxicologically tested as mixtures (e.g., toxaphene and PCBs). For such exceptions,
some attempt is made to assess the effects of environmental partitioning (i.e., different
patterns of environmental transport and degradation) on the criterion's validity. If these
effects cannot be assessed, an appropriate statement of uncertainty should accompany the
criterion.
Since relatively minor structural changes within a class of compounds can have
pronounced effects on their biological activities, reliance on class criteria should be
minimized. Whenever sufficient toxicologic data are available on a chemical within a
class, a compound-specific criterion should be derived. Nonetheless, for some chemical
classes, scientific judgment may suggest a sufficient degree of similarity among chemicals
within a class to justify a class criterion applicable to some of all members of a class.
The development of a class criterion takes into consideration the following:
1. A detailed review of the chemical and physical properties of chemicals within the
group should be made. A close relationship within the class with respect to chemical
activity would suggest a similar potential to reach common biological sites within
tissues. Likewise, similar lipid solubilities would suggest the possibility of
comparable absorption and tissue distribution.
2. Qualitative and quantitative data for chemicals within the group are examined.
Adequate toxicologic data on a number of compounds within a group provides a
more reasonable basis for extrapolation to other chemicals of the same class than
minimal data on one chemical or a few chemicals within the group.
3. Similarities in the nature of the toxicologic response to chemicals in the class provides
additional support for the prediction that the response to other members of the class
may be similar. In contrast, where the biological response has been shown to differ
markedly on a qualitative and quantitative basis for chemicals within a class, the
extrapolation of a criterion to other members of that class is not appropriate.
4. Additional support for the validity of extrapolation of a criterion to other members of
a class could be provided by evidence of similar metabolic and pharmacokinetic data
for some members of the class.
Based on the above considerations, in some cases a chemical class maybe divided into
various subclasses. Such divisions could be based on biological endpoints (e.g.,
carcinogens/non- carcinogens), potency, and/or sufficiency of data (e.g., a criterion for
some members of a class but no criterion for others). While no a priori limits can be placed
on the extent of subclassification, each subclassification must be explicitly justified by the
available data.
Class criteria, if properly derived and supported, can constitute valid scientific
assessments of potential risk/ safety. Conversely, the development of a class criterion from
an insufficient data base can lead to serious errors in underestimating or overestimating
risk/safety and should be rigorously avoided. Although scientific judgment has a proper
role in the development of class criteria, such criteria are useful and defensible only if
based on adequate data and scientific reasoning. The definition of sufficient data on
similarities in physical, chemical, pharmacokinetic, or toxicologic properties to justify a
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class criterion may vary markedly, depending on the degree of structural similarity and
the gravity of the perceived risk.
Consequently, the criterion derivation section of each document in which a class
criterion is recommended must explicitly address each of the key issues discussed above,
and define, as clearly as possible, the limitations of the proposed criterion as well as the
type of data needed to generate a compound-specific criterion.
A class criterion should be abandoned when sufficient data is available to derive a
compound-specific criterion that protects against the biological effect of primary concern;
e.g., the availability of a good subchronic study would not necessarily result in the
abandonment of a class criterion based on potential carcinogenicity.
The inability to derive a valid class criterion does not, and should not, preclude
regulation of a compound or group of compounds based on concern for potential human
health effects. The failure to recommend a criterion is simply a statement that the degree of
concern cannot be quantified based on the available data and risk assessment
methodology.
E. Essential Elements
Some chemicals, particularly certain metals, are essential to biological organisms at
low levels but may be toxic and/or carcinogenic at high levels. Because of potential toxic
effects, criteria should be established for such essential elements. However, criteria must
consider essentiality and cannot be established at levels that would result in deficiency of
the element in the human population.
Elements are accepted as essential if listed by NAS Food and Nutrition Board or a
comparably qualified panel. Elements not yet determined to be essential but for which
supportive data on essentiality exists need to be further reviewed by such a panel.
To modify the toxicity and carcinogenicity based criteria, essentiality must be
quantified either as a "recommended daily allowance" (RDA) or "minimum daily
requirement" (MDR). These levels are then compared to estimated daily doses associated
with the adverse effect of primary concern. The difference between the RDA or MDR and
the daily doses causing a specified risk level for carcinogens or ADIs for non-carcinogens
defines the spread of daily doses from which the criterion may be derived. Because errors
are inherent in defining both essential and maximum tolerable levels, the criterion is
derived from dose levels near the center of such a dose range. The decision to use either
the MDR or RDA is guided by the spread of the doses and the quality of the essentiality
and toxicity estimates.
The modification of criteria by consideration of essentiality must take into account all
routes of exposure. If water is a significant source of the MDR or RDA, the criterion must
allow for attainment of essential intake. Conversely, even when essentiality may be
attained from nonwater sources, standard criteria derivation methods may be adjusted if
the derived criterion represents a small fraction of the ADI or MDR. On a case-by-case
basis, the modification in the use of the guidelines may include the use of different safety
factors for non-carcinogens or other modifications which can be explicitly justified.
F. Use of Existing Standards
For some chemicals for which criteria are to be established, drinking water standards
already exist. These standards represent not only a critical assessment of literature, but
also a body of human experience since their promulgation. Therefore, accepting the
existing standard is valid unless compelling evidence exists to the contrary. This decision
should be made after considering the existing standards versus new scientific evidence
that has accumulated since the standards have been established. In several instances, the
peer review process recommended usage of the present drinking water standards.
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APPENDIX D
Derivation of1976 Philosophy of
Aquatic Life Criteria
Water quality criteria specify concentrations of water constituents that, if not exceeded, are
expected to support an organic ecosystem suitable for the higher uses of water. Such criteria are
derived from scientific facts obtained from experimental or in situ observations that depict organic
responses to a defined stimulus or material under identifiable or regulated environmental
conditions for a specified time period.
Water quality criteria are not intended to offer the same degree of strategy for survival and
propagation at all times to all organisms within a given ecosystem. They are intended not only to
protect essential and significant life in water and the direct users of water but also to protect life that
is dependent on water for its existence or may consume intentionally or unintentionally any edible
portion of such life.
The criteria levels for domestic water supply incorporate available data for human health
protection. Such values are different from the criteria levels necessary for protection of aquatic life.
The Agency's interim primary drinking water regulations (40 Federal Register 59566, December 24,
1975), as required by the Safe Drinking Water Act (42 U.S.C. 300f, et seq.), incorporate applicable
domestic water supply criteria. Where pollutants are identified in both the quality criteria for
domestic water supply and the Drinking Water Standards, the concentration levels are identical.
Water treatment may not significantly affect the removal of certain pollutants.
What is essential and significant life in water? Do Daphnia or stonefly nymphs qualify as such
life? Why does 1 / 100th of a concentration that is lethal to 50 percent of the test organisms (LC50)
constitute a criterion in some instances, whereas 1/2 or l/10th of some effect levels constitutes a
criterion in other instances? These are questions often asked of those who undertake the task of
criteria formulation.
The universe of organisms composing life in water is great in both kinds and numbers. As in the
human population, physiological variability exists among individuals of the same species in
response to a given stimulus. A much greater response variation exists among species of aquatic
organisms. Thus, aquatic organisms do not exhibit the same degree of harm, individually or by
species, from a given concentration of a toxicant or potential toxicant within the environment.
In establishing a level or concentration of a quality constituent as a criterion, it is necessary to
ensure a reasonable degree of safety for those more sensitive species that are important to the
functioning of the aquatic ecosystem, even though data on such species' response to the quality
constituent under consideration may not be available. The aquatic food web is an intricate
relationship of predator and prey organisms. A water constituent that may in some way destroy or
eliminate an important segment of that food web would, in all likelihood, destroy or seriously
impair other organisms associated with it.
Although experimentation relating to the effects of particular substances under controlled
conditions began in the early 1900s, the effects of any substance on more than a few of the vast
number of aquatic organisms have not been investigated. Certain test animals have been selected by
investigators for intensive investigation because of their importance to man, their availability to the
researcher, and their physiological responses to the laboratory environment. As general indicators
of organism responses, such test organisms are representative of the expected results for other
associated organisms. In this context, Daphnia or stoneflies or other associated organisms indicate
the general levels of toxicity to be expected among untested species. In addition, test organisms are
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themselves vital links within the food web that result in the fish population in a particular
waterway.
The ideal data base for criteria development would consist of information on a large percentage
of aquatic species and would show the community response to a range of concentrations for a tested
constituent during a long time period. This data is not available, but investigators are beginning to
derive such information for a few water constituents. Where only 96-hour bioassay data are
available, judgmental prudence dictates that a substantial safety factor be employed to protect all
life stages of the test organism in waters of varying quality, as well as associated organisms within
the aquatic environment that have not been tested and that may be more sensitive to the test
constituent.
Application factors have been used to provide the degree of protection required. Safe levels for
certain chlorinated hydrocarbons and heavy metals were estimated by applying an 0.01 application
factor to the 96-hour LC50 value for sensitive aquatic organisms. Flow-through bioassays have been
conducted for some test indicator organisms over a substantial period of their life history. In a few
other cases, information is available for the organism's natural life or for more than one generation
of the species. Such data may indicate a minimal effect level, as well as a no-effect level.
The word "criterion" should not be used interchangeably with or as a synonym for the word
"standard." The word "criterion" represents a constituent concentration or level associated with a
degree of environmental effect upon which scientific judgment may be based. As it is currently
associated with the water environment, it has come to mean a designated concentration of a
constituent that, when not exceeded, will protect an organism, an organism community, or a
prescribed water use or quality with an adequate degree of safety. A criterion, in some cases, may be
a narrative statement instead of a constituent concentration. On the other hand, a standard connotes
a legal entity for a particular reach of waterway or for an effluent. A water quality standard may use
a water quality criterion as a basis for regulation or enforcement, but the standard may differ from a
criterion because of prevailing local natural conditions — such as naturally occurring organic acids
— or because of the importance of a particular waterway, economic considerations, or the degree of
safety to a particular ecosystem that may be desired.
Toxicity to aquatic life generally is expressed in terms of acute (short-term) or chronic
(long-term) effects. Acute toxicity refers to effects occurring in a short time period: often death is the
end point. Acute toxicity can be expressed as the lethal concentration for a stated percentage of
organisms tested, or the reciprocal, which is the tolerance limit of a percentage of surviving
organisms. Acute toxicity for aquatic organisms generally has been expressed for 24- to 96-hour
exposures.
Chronic toxicity refers to effects through an extended time period. Chronic toxicity may be
expressed in terms of an observation period equal to the lifetime of an organism or to the time span
of more than one generation. Some chronic effects may be reversible, but most are not.
Chronic effects often occur in the species population rather than in the individual. If eggs fail to
develop or the sperm does not remain viable, the species would be eliminated from an ecosystem
because of reproductive failure. Physiological stress may make a species less competitive with
others and may result in a gradual population decline or absence from an area.
The elimination of a microcrustacean that serves as a vital food during the larval period of a
fish's life could result ultimately in the elimination of the fish from an area. The phenomenon of
bioaccumulation of certain materials may result in chronic toxicity to the ultimate consumer in a
food chain. Thus, fish may mobilize lethal toxicants from their fatty tissues during periods of
physiological stress; egg shells of predatory birds may be weakened to a point of destruction in the
nest; and bird chick embryos may have increased mortality rates. There may be a hazard to the
health of man if aquatic organisms with toxic residues are consumed.
The fact that living systems (i.e., individuals, populations, species, and ecosystems) can take up,
accumulate, and bioconcentrate human-made and natural toxicants is well documented. In aquatic
systems, biota are exposed directly to pollutant toxicants through submersion in a relatively
efficient solvent (water) and are exposed indirectly through food webs and other biological,
chemical, and physical interactions. Initial toxicant levels, if not immediately toxic and damaging,
may accumulate in the biota or sediment over time and increase to levels that are lethal or
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sublethally damaging to aquatic organisms or to consumers of these organisms. Water quality
criteria reflect a knowledge of the capacity for environmental accumulation, persistence, and effects
of specific toxicants in specific aquatic systems.
Ions of toxic materials frequently cause adverse effects because they pass through the
semipermeable membranes of an organism. Molecular diffusion through membranes may occur for
some compounds such as pesticides, polychlorinated biphenyls, and other toxicants. Some
materials may not pass through membranes in their natural or waste-discharged state, but in water
they may be converted to states that have increased ability to affect organisms. For example, certain
microorganisms can methylate mercury, thus producing a material that more readily enters
physiological systems. Some materials may have multiple effects: for example, an iron salt may not
be toxic; an iron floe or gel may be an irritant or clog fish gills to effect asphyxiation; iron at low
concentrations can be a trace nutrient but at high concentrations it can be a toxicant.
Materials also can affect organisms if their metabolic byproducts cannot be excreted. Unless
otherwise stated, criteria are based on the total concentration of the substance because an ecosystem
can produce chemical, physical, and biological changes that may be detrimental to organisms living
in or using the water.
In prescribing water quality criteria, certain fundamental principles dominate the reasoning
process. Establishing a level or concentration as a criterion for a given constituent assumed that
other factors within the aquatic environment are acceptable to maintain the integrity of the water.
Interrelationships and interactions among organisms and their environment, as well as the
interrelationships of sediments and the constituents they contain to the water above, are recognized
as fact.
Antagonistic and synergistic reactions among many quality constituents in water also are
recognized as fact. The precise definition of such reactions and their relative effects on particular
segments of aquatic life have not been identified with scientific precision. Historically, much of the
data to support criteria development was of an ambient concentration-organism response nature.
Recently, data are becoming available on long-term chronic effects on particular species. Studies
now determine carcinogenic, teratogenic, and other insidious effects of toxic materials.
Some unpolluted waters in the Nation may exceed designated criteria for particular
constituents. The natural quality of water is variable and certain organisms become adapted to that
quality, which may be considered extreme in other areas. Likewise, a single criterion cannot identify
minimal quality for the protection of the integrity of water for every aquatic ecosystem in the
Nation. Providing an adequate degree of safety to protect against long-term effects may result in a
criterion that cannot be detected with present analytical tools. In some cases, a mass balance
calculation can assure that the integrity of the waterway is not being degraded.
Water quality criteria do not have direct regulatory impact, but they form the basis for judgment
in several U.S. Environmental Protection Agency programs that are derived from water quality
considerations. For example, water quality standards developed by the States under section 303 of
the Act and approved by EPA are to be based on the water quality criteria, appropriately modified to
take account of local conditions. The local conditions to be considered include actual and projected
uses of the water, natural background levels of particular constituents, the presence or absence of
sensitive important species, characteristics of the local biological community, temperature and
weather, flow characteristics, and synergistic or antagonistic effects of combinations of pollutants.
Similarly, by providing a judgment on desirable levels of ambient water quality, water quality
criteria are the starting point in deriving toxic pollutant effluent standards pursuant to section
307(a) of the Act. Other EPA programs that use water quality criteria involve drinking water
standards, the ocean dumping program, designation of hazardous substances, dredge spoil criteria
development, removal of in-place toxic materials, thermal pollution, and pesticide registration.
To provide the water resource protection for which they are designed, quality criteria should
apply to virtually all of the Nation's navigable waters with modifications for local conditions as
needed. To violate quality criteria for any substantial length of time or in any substantial portion of a
waterway may result in an adverse affect on aquatic life and perhaps a hazard to humans or other
consumers of aquatic life.
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Quality criteria have been designed to provide long-term protection. Thus, they may provide a
basis for effluent standards, but criteria values need not become effluent standards. Certain
substances can be applied to the aquatic environment with the concurrence of a governmental
agency for the precise purpose of controlling or managing a portion of the aquatic ecosystem;
aquatic herbicides and pesticides are examples of such substances. For such occurrences, criteria
obviously do not apply.
In addition, pesticides applied according to official label instructions to agricultural lands and
forestlands may be washed to a receiving waterway by a torrential rainstorm. Under such
conditions, such diffuse source inflows should receive consideration similar to that of a discrete
effluent discharge, and in such instances the criteria should be applied to the principal portion of the
waterway rather than to that peripheral portion receiving the diffuse inflow.
The format for presenting water quality criteria includes a concise statement of the dominant
criterion or criteria for a particular constituent followed by a narrative introduction and a listing of
the references cited within the rationale. An effort has been made to restrict supporting data to those
that either have been published or are in press awaiting publication. A particular constituent may
have more than one criterion to ensure more than one water use or condition, i.e., hard or soft water
where applicable, suitability as a drinking water supply source, protection of human health when
edible portions of selected biota are consumed, provision for recreational bathing or waterskiing,
and permitting an appropriate factor of safety to ensure protection for essential warm- or coldwater
associated biota.
Criteria are presented for those substances that may occur in water where data indicate the
potential for harm to aquatic life, or to water users, or to the consumers of the water or aquatic life.
Presented criteria do not represent an all-inclusive list of constituent contaminants. Omissions from
criteria should not be construed to mean that an omitted quality constituent is either unimportant
or nonhazardous.
Excerpted from "Quality Criteria for Water," 1976, available from National Technical Information Service, #PB-263-943.
286
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Appendix E
Bioconcentration Factors
Used in the Calculation of Human Health Criteria
CHEMICAL NAME
BCF (3% lipid;
Acenaphthene
242
Acrolein
215
Acrylonitrile
30
Aldrin
4,670
Anthracene
30
Antimony
1
Arsenic
44
Benzene
5.2
Benzidine
87.5
Benzofluoranthene, 3,4-
30
Benzo(A) Anthracene
30
Benzo(A) Pyrene
30
Benzo(K) Fluoranthene
30
Bromoform
3.75
Butylbenzyl Phthalate
414
Cadmium
64
Carbon Tetrachloride
18.75
Chlordane
14,100
Chlorobenzene
10.3
Chlorodibromomethane
3.75
Chloroform
3.75
Chloronaphthalene, 2-
202
Chlorophenol, 2-
134
Chrysene
30
Copper
36
Cyanide
1
DDT
53,600
DDD
53,600
DDE
53,600
Dibenzo(A,H) Anthracene
30
Di-N-Butyl Phthalate
89
Dichlorobenzene, 1,2-
55.6
Dichlorobenzene, 1,3-
55.6
Dichlorobenzene, 1,4-
55.6
Dichlorobenzidine, 3,3-
312
Dichlorobromethane
3.75
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CHEMICAL NAME
BCF (3% lipid; 1/kg)
Dichloroethane, 1,1-
Dichloroethane, 1,2-
Dichloroethylene, 1,1-
Dichloroethylene, trans, 1,2-
Dichlorophenol, 2,4-
Dichloropropane, 1,2-
Dichloropropylene, 1,3-
Dieldrin
Diethyl Phthalate
Dimethylphenol, 2,4-
Dimethyl Phthalate
Dinitrophenol, 2,4-
2-Methyl-4,6-Dinitrophenol
Dinitrotoluene, 2,4-
Dioxin (2,3,7,8-TCDD)
Diphenylhydrazine, 1,2-
Di-2-Ethylhexyl Phthalate
Endosulfan-Alpha
Endosulfan-Beta
Endosulfan Sulfate
Endrin
Endrin Aldehyde
Bis(2-Chloroethyl) Ether
Bis(2-Chloroisopropyl) Ether
Bis(Chloromthyl) Ether
Ethylbenzene
Fluoranthene
Fluorene
Heptachlor
Heptachlor Epoxide
Hexachlorobenzene
Hexachlorobutadiene
Hexachlorocyclohexane-Alpha (A-BHC)
Hexachlorocyclohexane-Beta (B-BHC)
Hexachlorocyclohexane-Gamma (G-BHC)
Hexachlorocyclohexane-Delta (D-BHC)
Hexachlorocyclopentadiene
Hexachloroethane
Indeno (1,2,3-CD) Pyrene
Isophorone
Mercury
Methyl Bromide
Methylene Chloride
Naphthalene
288
1.2
5.6
1.58
40.7
4.11
1.91
4,670
73
93.8
36
1.5
5.5
3.8
5,000
24.9
130
270
270
270
3,970
3,970
6.9
2.47
0.63
37.5
1,150
30
11,200
11,200
8,690
2.78
130
130
130
130
4.34
86.9
30
4.38
5,500 Freshwater
3,760 Esturine
9,000 Marine
3.75
0.9
10.5
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CHEMICAL NAME
BCF (3% lipid; 1/kg)
Nickel
47
Nitronenzene
2.89
N-Nitrosodimethylamine
0.026
N-Nitrosodiphenylamine
136
N-Nitrosodi-N-Propylamine
1.13
PCB-1016
31,200
PCB-1221
31,200
PCB-1232
31,200
PCB-1242
31,200
PCB-1248
.31,200
PCB-1254
31,200
PCB-1260
31,200
Pentachlorophenol
11
Phenanthrene
30
Phenol
1.4
Pyrene
30
Selenium
6
Silver
0.5
Tetrachlorobenzene, 1,2,4,5-
1,125
Tetrachlorethane, 1,1,2,2-
5
Tetrachloroethylene
30.6
Thallium
116
Toluene
10.7
Toxaphene
13,100
Trichloroethane, 1,1,1-
5.6
Trichloroethane, 1,1,2-
10.6
Trichlorophenol, 2,4,6-
150
Vinyl Chloride
1.17
Zinc 47
NOTE: The bioconcentration factors (BCFs) listed here were used in the calculation of the
human health criteria in this document. EPA is considering updating these BCFs or using
Bioaccumulation Factors (BAFs). Until EPA formally updates these BCFs, these numbers
shuld be used to calculate the human health criteria.
289
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290
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APPENDIX F
Water Quality Criteria Documents
The U.S. Environmental Protection Agency has published water quality criteria for toxic
pollutant(s) categories. Copies of water quality criteria documents are available from the National
Technical Information Service (NTIS), 5285 Front Royal Road, Springfield, VA 22161, (703) 487-4650.
Prices of individual documents may be obtained by contacting NTIS. Order numbers are listed
below. Where indicated, documents may be obtained from the Water Resource Center, 401 M St.,
S.W. RC-4100, Washington, DC 20460, (202) 260-7786.
Chemical NTIS Order No. EPA Document No.
Acenaphthene
PB 81-117269
EPA 440/5-80-015
Acrolein
PB 81-117277
EPA 440/5-80-016
Acrylonitrile
PB 81-117285
EPA 440/5-80-017
Aesthetics
PB 263943
EPA 440/9-76-023
Aldrin/Dieldrin
PB 81-117301
EPA 440/5-80-019
Alkalinity
PB 263943
EPA 440/9-76-023
Aluminum
PB 88-245998
EPA 440/5-86-008
Ammonia
PB 85-227114
EPA 440/5-85-001
Ammonia (saltwater)
PB 89-195242
EPA 440/5-88-004
Antimony
PB 81-117319
EPA 440/5-80-020
Antimony (III) — aquatic

(draft)
resource center

Arsenic —1980
PB 81-117327
EPA 440/5-80-021
— 1984
PB 85-227445
EPA 440/5-84-033
Asbestos
PB 81-117335
EPA 440/5-80-022
Bacteria —1976
PB 263943
EPA 440/9-76-023
— 1984
PB 86-158045
EPA 440/5-84-002
Barium
PB 263943
EPA 440/9-76-023
Benzene
PB 81-117293
EPA 440/5-80-018
Benzidine
PB 81-117343
EPA 440/5-80-023
Beryllium
PB 81-117350
EPA 440/5-80-024
Boron
PB 263943
EPA 440/9-76-023
Cadmium —1980
PB 81-117368
EPA 440/5-80-025
— 1984
PB 85-224031
EPA 440/5-84-032
Carbon Tetrachloride
PB 81-117376
EPA 440/5-80-026
Chlordane
PB 81-117384
EPA 440/5-80-027
Chloride
PB 88-175047
EPA 440/5-88-001
Chlorinated Benzenes
PB 81-117392
EPA 440/5-80-028
Chlorinated Ethanes
PB 81-117400
EPA 440/5-80-029
Chlorinated Naphthalene
PB 81-117426
EPA 440/5-80-031
Chlorinated Phenols
PB 81-117434
EPA 440/5-80-032
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Chemical
NTIS Order No. EPA Document No.
Chlorine
PB 85-227429
EPA 440/5-84-030
Chloroalkyl Ethers
PB 81-117418
EPA 440/5-80-030
Chloroform
PB 81-117442
EPA 440/5-80-033
2-Chlorophenol
PB 81-117459
EPA 440/5-80-034
Chlorophenoxy Herbicides
PB 263943
EPA 440/9-76-023
Chlorpyrifos
PB 87-105359
EPA 440/5-86-005
Chromium —1980
PB 81-117467
EPA 440/5-80-035
—1984
PB 85-227478
EPA 440/5-84-029
Color
PB 263943
EPA 440/9-76-023
Copper —1980
PB 81-117475
EPA 440/5-80-036
— 1984
PB 85-227023
EPA 440/5-84-031
Cyanide
PB 85-227460
EPA 440/5-84-028
Cyanides
PB 81-117483
EPA 440/5-80-037
DDT and Metabolites
PB 81-117491
EPA 440/5-80-038
Demeton
PB 263943
EPA 440/9-76-023
Dichlorobenzenes
PB 81-117509
EPA 440/5-80-039
Dichlorobenzidine
PB 81-117517
EPA 440/5-80-040
Dichloroethylenes
PB 81-117525
EPA 440/5-80-041
2,4-Dichlorophenol
PB 81-117533
EPA 440/5-80-042
Dichloropropane /

Dichloropropene
PB 81-117541
EPA 440/5-80-043
2,4-Dimethylphenol
PB 81-117558
EPA 440/5-80-044
Dinitrotoluene
PB 81-117566
EPA 440/5-80-045
Diphenylhydrazine
PB 81-117731
EPA 440/5-80-062
Di-2-Ethylhexyl Phthalate-
aquatic (draft)
Dissolved Oxygen
Endosulfan
Endrin
Ethylbenzene
Fluoranthene
Gasses, Total Dissolved
Guidelines for Deriving
Numerical National
Water Quality Criteria
for the Protection of
Aquatic Organisms and
Their Uses
Guthion
Haloethers
Halomethanes
Hardness
Heptachlor
Hexachlorobenzene—
aquatic (draft)
Hexachlorobutadiene
Hexachlorocyclohexane
resource center
PB 86-208253
PB 81-117574
PB 81-117582
PB 81-117590
PB 81-117608
PB 263943
EPA 440/5-86-003
EPA 440/5-80-046
EPA 440/5-80-047
EPA 440/5-80-048
EPA 440/5-80-049
EPA 440/9-76-023
PB 85-227049
PB 263943
PB 81-117616
PB 81-117624
PB 263943
PB 81-117632
resource center
PB 81-117640
PB 81-117657
EPA 440/9-76-023
EPA 440/ 5-80-050
EPA 440/5-80-051
EPA 440/9-76-023
EPA 440/5-80-052
EPA 440/5-80-053
EPA 440/5-80-054
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Chemical
NTIS Order No. EPA Document No.
Hexachlorocyclopentadiene PB 81-117665
EPA 440/5-80-055
Iron

PB 263943
EPA 440/9-76-023
Isophorone

PB 81-117673
EPA 440/5-80-056
Lead —1980

PB 81-117681
EPA 440/5-80-057
— 1984

PB 85-227437
EPA 440/5-84-027
Malathion

PB 263943
EPA 440/9-76-023
Manganese

PB 263943
EPA 440/9-76-023
Mercury —1980

PB 81-117699
EPA 440/5-80-058
— 1984

PB 85-227452
EPA 440/5-84-026
Methoxychlor

PB 263943
EPA 440/9-76-023
Mirex

PB 263943
EPA 440/9-76-023
Naphthalene

PB 81-117707
EPA 440/5-80-059
Nickel —1980

PB 81-117715
EPA 440/5-80-060
— 1986

PB 87-105359
EPA 440/5-86-004
Nitrates/Nitrites

PB 263943
EPA 440/9-76-023
Nitrobenzene

PB 81-117723
EPA 440/5-80-061
Nitrophenols

PB 81-117749
EPA 440/5-80-063
Nitrosamines

PB 81-117756
EPA 440/5-80-064
Oil and Grease

PB 263943
EPA 440/9-76-023
Parathion

PB 87-105383
EPA 440/5-86-007
Pentachlorophenol
— 1980
PB 81-117764
EPA 440/5-80-065

— 1986
PB 87-105391
EPA 440/5-85-009
PH

PB 263943
EPA 440/9-76-023
Phenanthrene — aquatic

(draft)

resource center

Phenol

PB 81-117772
EPA 440/5-80-066
Phosphorus

PB 263943 .
EPA 440/9-76-023
Phthalate Esters

PB 81-117780
EPA 440/5-80-067
Polychlorinated Biphenyls
Polynuclear Aromatic
Hydrocarbons
Selenium —1980
— 1987
Silver
Silver — aquatic (draft)
Solids (dissolved) and
Salinity
Solids (suspended) and
Turbidity
Sulfides/Hydrogen Sulfide
Tainting Substances
Temperature
2,3,7,8-Tetrachlorodibenzo-
P-Dioxin
Tetrachloroethylene
Thallium
Toluene
PB 81-117798
PB 81-117806
PB 81-117814
PB 88-142239
PB 81-117822
resource center
PB 263943
PB 263943
PB 263943
PB 263943
PB 263943
PB 89-169825
PB 81-117830
PB 81-117848
PB 81-117863
EPA 440/5-80-068
EPA 440/5-80-069
EPA 440/5-80-070
EPA 440/5-87-008
EPA 440/5-80-071
EPA 440/9-76-023
EPA 440/9-76-023
EPA 440/9-76-023
EPA 440/9-76-023
EPA 440/9-76-023
EPA 440/5-84-007
EPA 440/5-80-073
EPA 440/5-80-074
EPA 440/5-80-075
293
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Chemical
NTIS Order No. EPA Document No.
Toxaphene —1980
— 1986
Tributyltin — aquatic
(draft)
Trichloroethylene
2,4,5-Trichlorophenol —
aquatic (draft)
Vinyl Chloride
Zinc —1980
— 1987
PB 81-117863
PB 87-105375
resource center
PB 81-117871
resource center
PB 81-117889
PB 81-117897
PB 87-143581
EPA 440/5-80-076
EPA 440/5-86-006
EPA 440/5-80-077
EPA 440/5-80-078
£PA 440/ 5-80-079
EPA 440/5-87-003

EPA Library Region 4
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