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WATER QUALITY CRITERIA FOR THE PROTECTION OF AQUATIC LIFE  AND  ITS  USES
                                AMMONIA
                              Final Draft




                            20 January 1983
                              Prepared By








                 U.S.  Environmental Protection Agency




                  Office of Research and Development




                   Environmental Research Laboratory



                           Duluth,  Minnesota











                    US Frv'r^r>?v^nt?! Protection Agency

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                              TABLE OF CONTENTS
                                                                         Page
Introduction		     1

Acute Toxicity to Animals   	     7

Chronic Toxicity to Animals   	    42
Toxicity to Plants	    59
Bioaccumulat ion	    64

Other Data	    65
Unused Data	    76
Summary	    80

National Criteria  	    84
Examples of Site-Specific Criteria 	    87

Tables	    91

     1 .  Acute values for ammonia	    91

     2.  Chronic values for ammonia  	   110

     3.  Species and family mean acute values (corrected  for pH)
         for ammonia	.  .  .  .   114
     4.  Plant values for ammonia	117
     5.  Other data for ammonia	119
References	132
                                          U.S. Envlronrnentc!  Trctc^'cn Agenc
                                     li

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                                 INTRODUCTION*




      In aqueous ammonia  solutions,  un-ionized ammonia exists in equilibrium




with  the ammonium  ion and  the  hydroxide  ion.   The  equation expressing this




equilibrium can be written as:





      ^3(8) + nH2ฐU) *  NH3'nH2ฐ(aq) * m*+ + OH~  +  (n-OH2ฐU)*



As indicated in this equation,  the  dissolved  ammonia molecule exists in




hydrated form.  The dissolved  un-ionized ammonia is  represented for




convenience simply as NHj.  The  ionized form  is represented as NH^+.




The term total ammonia refers  to  the sum of these; i.e.,  NH3 + NH^  .




      The toxicity of aqueous ammonia solutions to  aquatic organisms is




primarily attributable to  the  NH-j Species,  with the  NH^"*"  species  being




relatively less toxic (Chipraan  1934; Wuhrmann et al.  1947;  Wuhrmann and Woker




1948; Tabata 1962; Armstrong et al. 1978; Thurston et  al.  1981c).   It is,




therefore, important to know the concentration of  NH-j  in  any aqueous




ammonia solution in order  to determine what concentrations  of  total ammonia




are toxic to aquatic life.




     The concentration of  NH^  is dependent  on a number  of  factors in




addition to total ammonia  concentration (Emerson et  al. 1975;  Thurston  et  al.




1979).  Most important among these  are pH and temperature;  the concentration




of NH-j increases with increasing pH and with  increasing temperature.  The




ionic strength is another  important influence on this equilibrium.   There
* The reader is referred to the Guidelines for Deriving Numerical National




  Water Quality Criteria for the Protection of Aquatic Life and  Its  Uses  (5




  July 1983) (U.S. Environmental Protection Agency  1983 ) in order to




  understand better the following text, tables, and calculations.



                                      1

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is a decrease in tht; percentage of un-Ionized ammonia as  the  limit.:  strength




increases in hard water or in dilute saline solution.  In most natural




freshwater systems the reduction of percent NHg attributable  to dissolved




solids is negligible.  In saline or very hard waters there will be  small but




measurable decreases in the percent NHo-




     Current analytical methods do not permit measurement of  NH3 and




NH^  separately.  A number of analytical methods are available for  direct




determination of total ammonia concentrations in aqueous solutions.   Once




total ammonia is measured, and the pH and temperature of the  solution




determined, the fraction of total ammonia present as NH3 can  be calculated




based on the ammonia-water equilibrium.  A review of analytical methods for




ammonia in aqueous solution has been prepared by Richards (1981).




     Emerson et al. (1975) carried out a critical evaluation  of the




literature data on the ammonia-water equilibrium system and published




calculations of values of pKa at different temperatures and of percent




NH-j in ammonia solutions of zero salinity as a function of pH and




temperature.  The following table, reproduced from Emerson et al. (1975),




provides values for percent NHg at one-degree temperature intervals from 0




to 30 C, and pH intervals of 0.5 pH unit from pH 6.0 to 10.0.  An expanded




version of this percent NH^ table is provided in Thurston et  al. (1979),




which provides tabulated values of the NH-j fraction, expressed as




percentage of total ammonia, at temperature intervals of 0.2  degree from 0.0




to 40.0 C, and pH intervals of 0.01 pH unit from pH 5.00 to 12.00.  For




salt water, reports by Whitfield (1974) and Skarheim (1973) provide




calculations of NHg as a function of pH, temperature, and salinity.

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           Percent NH-j in aqueous ammonia solutions  tor 0-30 C  and  pH  b-10.
Temp.
(C)
0
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30

6.0
.00827
.00899
.00977
.0106
.0115
.0125
.0136
.0147
.0159
.0172
.0186
.0201
.0218
.0235
.0254
.0274
.0295
.0318
.0343
.0369
.0397
.0427
.0459
.0493
.0530
.0569
.0610
.0654
.0701
.0752
.0805

6.5
.0261
.0284
.0309
.0336
.0364
.0395
.0429
.0464
.0503
.0544
.0589
.0637
.0688
.0743
.0802
.0865
.0933
.101
.108
.117
.125
.135
.145
.156
.167
.180
.193
.207
.221
.237
.254

7.0
.0826
.0898
.0977
.106
.115
.125
.135
.147
.159
.172
.186
.201
.217
.235
.253
.273
.294
.317
.342
.368
.396
.425
.457
.491
.527
.566
.607
.651
.697
.747
.799

7.5
.261
.284
.308
.335
.363
.394
.427
.462
.501
.542
.586
.633
.684
.738
.796
.859
.925
.996
1.07
1.15
1.24
1.33
1.43
1.54
1.65
1.77
1.89
2.03
2.17
2.32
2.48
pH
8.0
.820
.891
.968
1.05
1.14
1.23
1.34
1.45
1.57
1.69
1.83
1.97
2.13
2.30
2.48
2.67
2.87
3.08
3.31
3.56
3.82
4.10
4.39
4.70
5.03
5.38
5.75
6.15
6.56
7.00
7.46

8.5
2.55
2.77
3.00
3.25
3.52
3.80
4.11
4.44
4.79
5.16
5.56
5.99
6.44
6.92
7.43
7.97
8.54
9.14
9.78
10.5
11.2
11.9
12.7
13.5
14.4
15.3
16.2
17.2
18.2
19.2
20.3

9.0
7.64
8.25
8.90
9.60
10.3
11.1
11.9
12.8
13.7
14.7
15.7
16.8
17.9
19.0
20.2
21.5
22.8
24.1
25.5
27.0
28.4
29.9
31.5
33.0
34.6
36.3
37.9
39.6
41.2
42.9
44.6

9.5
20.7
22.1
23.6
25.1
26.7
28.3
30.0
31.7
33.5
35.3
37.1
38.9
40.8
42.6
44.5
46.4
48.3
50.2
52.0
53.9
55.7-
57.5
59.2
60.9
62.6
64.3
65.9
67.4
68.9
70.4
71.8

10.0
45.3
47.3
49.4
51.5
53.5
55.6
57.6
59.5
61.4
63.3
65.1
66.8
68.5
70.2
71.7
73.3
74.7
76.1
77.4
78.7
79.9
81.0
82.1
83.2
84.1
85.1
85.9
86.8
87.5
88.3
89.0
[From Emerson et al. 1975;  reproduced with permission from the Journal of the
Fisheries Research Board of Canada.]

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     Concentrations of ammonia have been  reported  in  the  .-U|u.-ic u~  toxu'ltv




literature as a variety of different forms,  such as NH3ป  NH^, N^-N,




       NH^Cl, and others.  The use in a literature article of  the  terras




NHg, NH-j-N, or ammonia-nitrogen may not necessarily mean  un-ionized




ammonia, but may be the author's way of expressing total  ammonia.  The use  of




the term NH-j in this document always means un-ionized ammonia, and NH^N




means un-ionized ammonia-nitrogen.




     Throughout the following, all quantitative ammonia data have been




expressed in terms of un-ionized ammonia, as mg/liter Nt^, for ease  in




discussion and comparison.  Authors' ammonia concentration values are given




as reported if authors provided data expressed as rag/liter NHg-  If  authors




reported only total ammonia values, or used concentration units other than




mg/liter, these were used with the reported pH and temperature values to




calculate mg/liter un-ionized NH3ซ  For calculations of NH3 in fresh




water the table of Thurston et al. (1979) was used.  For  calculations in salt




water the table of Skarheira (1973) was used.




     This Criterion Document was prepared in accordance with U.S.




Environmental Protection Agency guidelines for the derivation of national




water quality criteria (U.S. Environmental Protection Agency 1982a).  The




literature cited herein was obtained from a search of the literature through




October 1982; data from primary references only were used.




     Of the literature cited in this document, a significant number  of papers




provided insufficient pH and temperature data to enable calculation  of NH3




concentrations; such papers were relegated to the section on "Unused Data"




unless they provided useful qualitative or descriptive information.  In some




instances information missing in published papers on experimental  conditions

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,was  obtained  through  correspondence  with  authors;  data  obtained in this
manner  are  so  Indicated by  footnotes.
      Compounds used in the  ammonia toxicity  testing  summarized  here and their
formulas and  Chemical Abstracts  Services  (CAS)  Registry Numbers are given
below:
          Compound                      Formula                 CAS No.
      Ammonia                            NH3                     7664417
     Ammonium acetate                   ^^02^02               631618
     Ammonium bicarbonate               NH^Hct^                 1066337
     Ammonium carbonate                 (NH^^COg               506876
     Ammonium chloride                  NH^Cl                   12125029
     Ammonium hydrogen phosphate        (NH^HPO^              7783280
     Ammonium hydroxide                 NH4OH (NH3'H20)         1336216
     Ammonium sulfate                   (NHSQ                7783202
Papers stating use of other sources of ammonia were included  if  the  source
(e.g., excreted NH^ from fish) was deemed satisfactory.  Papers  using
complex chemicals (e.g., ammonium ferricyanide, dec ylt rime thylammonium
bromide) were not used.  Finally, papers on ammonium compounds (e.g., NH^F,
          having anions that either might be themselves toxic or  that
would preclude calculation of NH3 concentration from the aqueous  ammonia
equilibrium relationship were not used.
     A number of review articles or books dealing with ammonia as an aquatic
pollutant are available.  Water quality criteria for ammonia  have been
recommended in some of these.  Liebmann (1960), McKee and Wolf (1963), Epler
(1971), Becker and Thatcher (1973), Tsai (1975), Hampson (1976),  Steffens
(1976), Colt and Armstrong (1979), and Armstrong (1979) have  published
summaries of ammonia toxicity.   Literature reviews including  factors
affecting ammonia toxicity and physiological effects of ammonia  toxicity to
aquatic, organisms have been published by Lloyd (1961b), Lloyd and Herbert
                                      5

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(1962), Warren (1962), Visek (1968), Lloyd and Swift (1976), and Kinne




(1976).  Literature reviews of ammonia toxicity information relating to




criteria recommendations have been published by U.S. Federal Water Pollution




Control Administration (1968), European Inland Fisheries Advisory Commission




(1970), National Academy of Sciences and National Academy of Engineering




(1973), Willingham (1976), U.S. Environmental Protection Agency (1977, 1980),




National Research Council (1979), Willinghara et al.  (1979), and Alabaster and




Lloyd (1980).

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                           ACUTE TOXICITY TO ANIMALS
1  i

 Freshwater Invertebrates


      Acute toxicity of ammonia to freshwater invertebrate species has been

 much less studied than that to fishes.   The preponderance of available


 invertebrate data is comprised of studies with arthropods, primarily

 crustaceans and insects.   LC50 and EC50 data are  summarized in Table 1 for 12


 species representing nine families.

      The acute toxicity of ammonia to Daphnia magna (Table 1)  has been


 studied by several investigators, with  reported 48-hour LC50 values  ranging

 from 0.53 to 4.94 mg/liter NH3 (Parkhurst et al.  1979,  1981;  Reinbold and


 Pescitelli 1982a;  Russo et al.,  in prep.).


      Exposures (48 hours) of  I),  magna to NlfyCl in dilution water  from two

 different sources were conducted by Russo et al.  (in prep.).   LC50 values


 (Table  1) ranged  from 2.4 to  2.8 mg/liter NH3 in  water  of pH  7.95 to 8.15


 and  hardness 192  to 202 rag/liter CaC03,  and  from  0.53 to 0.90  mg/liter

 NH3  in  water of pH 7.4 to 7.5 and hardness 42 to  48  mg/liter  CaC03-   On

 an acute (48-hour LC50) basis,  in dilution water  from the same source,


 Ceriodaphnia acanthina, Simocephalus  vetulus,  and IK magna all exhibited

 similar sensitivities (Table  1)  to ammonia (Mount 1982;  Russo  et  al.,  in

 prep.).  The 48-hour LC50 value  (Table  1)  of 1.16 mg/liter NH3 reported by

 DeGraeve et  al. (1980) for Daphnia pulicaria falls within the  range  of values

 reported for IK magna.  Anderson (1948)  reported  a  threshold  toxicity  value

 (Table  5)  for ฃ.  magna of 2.4 to 3.6  mg/liter NH3  in Lake Erie water.

 Threshold  concentration was taken to  mean  the  highest concentration  that

 would just  fail to  immobilize the test  animals under prolonged exposure


 (Anderson  1944).   A minimum lethal concentration  of 0.55  rag/liter  NH3  was

 reported  for _D. magna by  Malacca  (1966), and  a 24-hour I C50 value  of  1.50


                                      7

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rag/liter NH3 was reported by Gyore and Olah (1980) for Moina rectirostris




(Table 5).




     Buikema et al. (1974) reported an EC50 (Table 5) for NH3 toxicity  to a




bdelloid rotifer, Philodina acuticornis, to be 2.9-9.1 mg/liter NH3




(calculated using reported pH values of 7.4 to 7.9).  Tests of ammonia




toxicity to a flatworm, Dendrocoelum lacteum (Procotyla fluviatilis), and




tubificid worm, Tubifex tubifex, yielded LC50 values (Table 1) of 1.4 and 2.7




mg/liter NH3, respectively (Stammer 1953).




     Thurston et al. (in prep., a) conducted 26 flow-through toxicity tests




with three mayfly, two stonefly, one caddisfly, and one isopod species; all




tests were conducted with water of similar chemical composition.




Ninety-six-hour LC50 values ranged from 1.8 to 5.9 mg/liter NH3 (Table  1).




Results also indicated that a 96-hour test is not long enough to determine




the acutely lethal effects of ammonia to the species tested, inasmuch as an




asymptotic LC50 is not obtained within 96 hours.  Percent survival data




(Table 5) were reported for some mayfly, stonefly, and caddisfly tests  in




which LC50 values were not obtained;  60 to 100 percent survival occurred at




test concentrations ranging from 1.5 to 7.5 mg/liter NH3-  Gall (1980)




tested NH^Ci with Ephemerella sp. near excrucians.  Organisms were exposed




to ammonia for 24 hours, followed by 72 hours in ammonia-free water;




mortality observations were made at the end of the overall 96-hour period.




An EC50 value (Table 5) of 4.7 mg/liter NH3 was obtained.




     Ammonia toxicity tests conducted using dilution water from the Blue




River in Colorado resulted in no mortalities of either scud (Gammarus




lac us tr is) or D^. majgjia after 96 hours'  exposure to 0.08 ing/liter NH3ซ   In a




second test using river water buffered with sodium bicarbonate, 13 percent




mortal icy occurred with scud at several concentrations tested, including the




                                      8

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 highest and lowest of 0.~"* And O.I? mg/Htov \'H^; *ov*n ami 1'ป ;*ซM .^ u

 mortality  occurrtvi with r>- '-c-a^'.ia at  the s'ilie*  ct =>l.

 1981).

      Five  freshwater  mussel species  (Amblema _p.  plicata,  Anodonta imbecillis,

 Corbicula  manilensis,  Cyrtonaias t amp i c o en si s, and Toxqlasma texasensis) were

 exposed  for  165  hours  (Table 5)  to  a concentration of 0.32 rag/liter  NH^J T_.

 texasensis was most tolerant to  ammonia,  and A.  j>. plicata was  most  sensitive

 (Home  and Mc.Tntosh 1979).  During  the  ammonia tolerance  tests, the  more

 tolerant species generally had  their shells  tightly shut,  whereas the least

 tolerant species continued siphoning or had  their mantles  exposed.  Acute

 exposures  of  the freshwater crayfish (Orconectes nais)  to  NH^Ci gave LC50

 values  (Table 1) of 3.15  and 3.82 mg/liter NH3 (Evans 1979;  Hazel et al.

 1979).  An LC50 value  (Table 1)  of 8.00 mg/liter was  reported  for the beetle

 (Stenelmis sexlineata) by Hazel  et al.  (1979).


 Freshwater Fishes

     Acute toxicity tests with freshwater  fish species  have  been  conducted

with 23 different  species  from nine  families; 96-hour LC50 values are

 summarized in Table 1.

     The acute toxicity of  ammonia to rainbow trout (Salmo gairdneri)  has

been studied by many investigators,  with reported  96-hour  LC50  values  ranging

 from 0.16 to 1.1 mg/liter  NH3 (Table 1).

     Thurston and Russo (in  press) conducted  71  toxicity tests  with  rainbow

 tiout ranging in size  from  sac fry (<0.1 g)  to 4-year-old  adults  (2.6 kg),  in

wai er of uniform chemical  composition.  LC50  values (Table 1) ranged from

O.lo to 1.1 mg/liter NH-j  for 96-hour exposures.  Fish susceptibility to

NH-j decreased with increasing weight over  the range 0.06-2.0 g, but

gradually increased above  that weight range.  LC50 values  for 12- and  35-day
                                      9

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exposures (Table 5) were not greatly different from 96-hour values.  No




statistically significant differences in results were observed when different




ammonium salts [NH^Cl, NH^HCO-j, (NH^HPO^, (NH^SO^l were used as the




toxicants.  Grindley (1946) also reported observing no appreciable difference




in toxicity between toxicant solutions of NH^Cl and (NH^)9So^ with




rainbow trout tests (Table 5).




     LC50 values (Table 1) ranging from 0.16 to 1.0 mg/liter NH3 for




96-hour exposures of rainbow trout to ammonia were reported by Calamari et




al. (1977, 1981), Broderius and Smith (1979), Holt and Malcolm (1979),




DeGraeve et al. (1980), and Reinbold and Pescitelli (1982b).  Ball (1967)




reported an asymptotic (five-day) LC50 value of 0.50 mg/liter NH3•  Acute




exposures to ammonia of rainbow trout of life stages ranging from one to 345




days'  post-fertilization (325 days post-hatch) were conducted by Calamari et




al. (1977, 1981).  They reported a tenfold increase in the speed of




intoKication processes between the embryonic and free larval stages; embryos




and fingerlings (about one year old) were found to be less sensitive than the




other  life stages studied.




     LCSO values ranging from 0.49 to 0.70 mg/liter NH3 for 3-,  24-, and




48-hour exposures (Table 5) were reported by Herbert (1961, 1962), Herbert




and Shurben (1964,  1965), and Herbert and Vandyke (1964).  Rainbow trout (826




days old) subjected to 29.6 mg/liter NH-j reacted rapidly and strongly,




overturned within two to three hours, and died within four hours (Corti 1951)




(Table 5).  Rainbow trout embryos and alevins were reported (Rice and Stokes




1975)  to tolerate 3.58 mg/liter NH3 during 24-hour exposures; suscep-




tibility increased  during yolk absorption, with the 24-hour LCSO for




85-day-old fry being 0.068 mg/liter NH3 (Table 5).  Nehring (1962-63)




reported survival times of 1.3 and 3.0 hours at concenttations of 4.1 and 0.7




                                     10

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,mg/liter  NH^,  respectively  (Table  5).   Danecker  (1964)  reported survival




 times  of  S  to  60  minutes  at 0.4  to 4.0  rag/liter  NHy,  respectively,  with




 <0.2 m,f*/l.1ter  given  as  ,1  no-mortality concentration  (Table  5).   Allan et al.




 (1958)  reported a median  survival  time  of  1000 minutes  at 0.18  mg/liter NH3




 (Table  5).




     An acute  value  of  0.2  mg/liter NH3 attributed to Liebmann  (1960) has




been widely cited, in the EPA  "Red Book" (U.S. Environmental  Protection




Agency 1977) and  elsewhere,  as being the lowest  lethal  concentration  reported




for salmonids.  It is worthwhile to mention here a clarification  and




correction that was  published  in the American Fisheries Society's "Red Book




Review" (Willingham  et  al.  1979):   The  research  reported by Liebmann  (1960)




was that  of Wuhrmann and  Woker (1948);   recomputation of the Wuhrmann  and




Woker data, using more  accurate aqueous ammonia equilibrium tables, indicates




an effect level of approximately 0.32 mg/liter NH3, not 0.2 mg/liter  NH3




as cited by Liebmann.




     A 96-hour LC50 value of 0.44 mg/liter NH3 was reported for rainbow




trout In a test conducted using dilution water from the Blue  River  in




Colorado  (Miller  et  al. 1981).  Pitts (1980) conducted  toxicity tests using




ammonium chloride and river water.   Tests were conducted with rainbow trout,




and LC50 values ranged  from 0.2 to 0.9 mg/liter NH3 for 96-hour exposures




at temperatures of 10 and 15 C.




     Although acute toxicity studies with salmonids have been conducted




preponderantly with rainbow trout,  some data are also available for a  few




other salmonid species.  Thurston et al. (1978)  investigated  the  toxicity of




a.timoriia to cutthroat  trout (Salmo dark!),  and reported 96-hour LC50  values




of 0.52 to 0.80 rag/liter NH3 (Table 1).   Thurston and Russo (1981)  reported




a 96-hoi,r  LC50 value  of 0.76 rag/liter NH3 for golden trout (Salmo




                                     il

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aguabonlta) (Table 1).  Taylor (1973) subjected brown trout (Salmo trutta) to




0.15 rag/liter NH3 for 18 hours, resulting in 36 percent mortality (Table




5); when returned to ammonia-free water, the test fish recovered after nearly




24 hours.  No mortalities occurred during a 96-hour exposure at 0.090




mg/liter NH3, although fish would not feed.  Woker and Wuhrmann (1950)




reported 0.8 mg/liter NH-J was not acutely toxic to brown trout (Table 5).




A 96-hour LC50 value of 0.47 mg/liter NII3 was reported for brown trout




tested using dilution water from the Blue River in Colorado (Miller et al.




1981).  Phillips (1950) reported that brook trout (Salvelinus fontinalis)




evidenced distress within 1.75 hours at a concentration of 3.25 mg/liter




NH3 and within 2.5 hours at 5.5 mg/liter (Table 5).




     Toxicity tests (Tables 1, 5) on (NH^^SO^ with pink salmon




(Oncorhynchus gorbuscha) at different stages of early life stage development




(Rice and Bailey 1980) showed that late alevins near swim-up stage were the




most sensitive (96-hour LC50 = 0.083 mg/liter NHg), and eyed embryos were




the most tolerant, surviving 96 hours at >1.5 mg/liter NH3.  Buckley (1978)




reported a 96-hour LC50 value of 0.55 mg/liter NH3 for fingerling coho




salmon, Oncorhynchus kisutch (Table 1).  Herbert and Shurben (1965) reported




a 24-hour LC50 value (Table 5) of 0.28 mg/liter NH3 for Atlantic salmon




(Salmo s_alar) .  A comparison of relative susceptibilities of salmon smolts




and yearling rainbow trout to 24-hour exposures to NH^Cl showed that the




salmon were appreciably more susceptible than the trout in fresh water




(Ministry of Technology, U.K. 1963).




     Data are available on the acute toxicity of ammonia to a variety of




non-salmonid fish species.  Thurston et al. (in press, a) studied the




toxicity of ammonia to fathead minnows (Pimephales promelas) of sizes ranging




from 0.1 to 2.3 g;  LC50 values from 29 tests ranged from 0.75 to 3.4 mg/liter




NH3 (Taoie 1).  Toxictty was not dependent upon test tish size or source.



                                     12

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'•LC50 values ranging from 0.73 to 1.7 rag/liter NH^ (Tables 1,5) for fathead




 minnows were also reported  by Sparks (1975),  DeGraeve et al.  (1980),  and




 Reinbold and Pescitelli (1982b).  Toxicity tests with fathead minnows using




 ammonium chloride and  river water yielded 96-hour LC50 values ranging from




 0.6 to 2.4  tng/liter Nt^J  fathead minnows  exposed to 0.12 mg/liter NH3 in




 river water for 28 days incurred no  mortalities  (Pitts 1980).




      LC50 values (Table 1)  for white sucker (Catostomus commersoni) exposed




 to  ammonium chloride solutions for 96 hours (Reinbold and Pescitelli  1982c)




 were 1.40 and 1.35 mg/liter NH3•  Reported LC50  values (Table 1)  for




 96-hour exposures of bluegill (Lepomis macrochirus)  ranged from 0.26  to 4.60




 mg/liter NH3 (Emery and Welch 1969;  Lubinski  et  al.  1974;  Roseboom and




 Richey 1977;  Reinbold  and Pescitelli 1982b;  Smith and Roush,  in prep.).   LC50




 values (Table 1) of 0.7 to  1.8 mg/liter NH3 for  smallmouth bass




 (Micropterus dolomieui)  and 1.0 to 1.7 mg/liter  NH-j  for largemouth bass




 (Micropterus saltnoides) were reported by  Broderius  et al.  (in prep.)  and




 Roseboom and Richey (1977),  respectively,  for 96-hour exposures.   Sparks




 (1975)  reported 48-hour LC50 values  (in parentheses,  as rag/liter  ^3) for




 bluegill  (2.30)  and channel  catfish  (2.92), Dowden and Bennett  (1965)




 reported a  24-hour LC50 value for  goldfish (Carassius auratus)  (7.2), and




 Chipman (1934)  reported  lethal threshold  values  of 0.97 to 3.8  mg/liter  NH3




 for  goldfish (Table 5).  Turnbull  et  al.  (1954)  reported  a 48-hour LC50  value




 for  bluegill to be within the range  0.024  to 0.093 mg/liter NH3 (Table 5);




 during  the  exposure they observed  that the  fish  exhibited  a lack  of




 perception  to avoid objects.




      Reported 96-hour LC50 values  (Table  1) for  channel catfish (Ictalurus




 pane, tat us)  ranged  from  1.8  to  4.2 mg/liter  NH3 (Colt  and Tchobanoglous




 1976;  Koseboora  and  Rirhey 1977;  Reinbold and Pescitelli  L982d).   Vaughn  and




                                      13

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Simco (1977) reported a 48-hour LC50 for channel catfish of 1.24 to 1.96




mg/liter NH-j, and Knepp and Arkin (1973) reported one-week LC50 values of




0.97 to 2.0 mg/liter NH3 (Table 5).  From studies with bluegill, channel




catfish, and largemouth bass, Roseboom and Richey (1977) reported that




bluegill susceptibility was dependent upon fish weight, with 0.07-g fish




being slightly more sensitive than either 0.22- or 0.65-g fish; size had




little effect upon channel catfish or bass susceptibility.




     LC50 values (Table 1) were determined with two species of field-




collected fishes indigenous to Kansas streams, orangethroat darter




(Etheostoma spectabile ) and red shiner (Notrqpis lutrensis) (Hazel et al.




1979); 96-hour LC50 values were 0.90 and 1.07 mg/liter NH3 for darter and




2.83 for shiner.  Commercially obtained largeraouth bass, channel catfish, and




bluegill (18 fish of each species) were also exposed for 96 hours to a




concentration of 0.21 mg/liter NH3, resulting in zero mortality for




bluegill and channel catfish and one mortality (6 percent) among the




largeraouth bass tested.




     LC50 values (Table 1) ranging from 0.48 to 3.2 mg/liter NH3 for




NH4C2H3ฐ2' (NH4)2C03, NH4Cl, NH4OH, and (NH4)2S04 in 96-hour




exposures of mosquitofish (Gambusia affinis) in waters with suspended solids




ranging from <25 to 1400 mg/liter were reported by Wallen et al. (1957).




Susceptibility of mosquitofish to ammonia was studied by Hemens (1966) who




reported a 17-hour LC50 value of 1.3 mg/liter NH3 (Table 5); he also




observed that male fish were more susceptible than females.  Powers (1920)




reported the relative susceptibilities of three fish species to ammonium




chloride to be (most sensitive to least sensitive):  straw-colored minnow




(Notropis b 1 ennius) > bluntnose minnow (Pimephale s no ta tusj > goldfish.
                                     14

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      Rubin  and Elmaraghy  (1976,  1977)  tested  guppy  (Poecilla reticulata)  fry




 and  reported 96-hour LC50 values  (Table  1) averaging 1.50 mg/liter NH3;




 mature guppy males were more  tolerant, with 100 percent survival for 96 hours




 at concentrations of 0.17 to  1.58 mg/liter NH3ซ  LC50 values (Table 1) of




 0.15  and 0.20 mg/liter NH3 at pH  6.0,  and of  0.52 and 2.13 mg/liter NH3




 at pTI 8.0,  were  reported  by Stevenson  (1977)  for white perch (Morone




 americana).  LC50 values  (96  hours) of 1.20 and 1.62 mg/liter NH3 for




 spotfln shiner (Notr o pis  sp1lopterus), and of 1.20 mg/liter NHj for golden




 shiner (No t emi gonus cryso1eucas), were reported by Rosage et al. (1979) and




 Baird et al. (1979), respectively (Table 1).  Jude (1973), Reinbold and




 Pescitelli  (1982a), and McCormick et al. (in  prep.) reported 96-hour LC50




 values ranging from 0.6 to 2.1 mg/liter NH3 for green sunfish (Lepomis




 cyanellus)  (Table 1).  Pumpkinseed sunfish (Lepomi8 gibbosus) were tested by




 Jude  (1973) and Thurston  (1981), with  reported 96-hour LC50 values ranging




 from  0.14 to 0.86 mg/liter NH3 •  Mottled sculpin (Cottus bairdi) were




 tested by Thurston and Russo  (1981), yielding a 96-hour LC50 of 1.39 mg/liter




 NH3 (Table  1).  Ball (1967) determined an asymptotic (six-day) LC50 value




 (Table 5) of 0.44 mg/liter NH3 for rudd (Scardinius erythrophthalmus).  He




 compared the asymptotic LC50 values for this species against that obtained




 within two  days for rainbow trout.  Although  the trout had proven to be more




 sensitive to ammonia than had rudd during the first day of the tests, the




 asymptotic LC50 for both species showed little difference.




      Rao et al. (1975) reported a 96-hour LC50 for carp (Cyprinus carpio) of




 1.1 mg/liter NH3 (Table 1).  Carp exposed to 0.24 mg/liter NH3 exhibited




 no adverse effects in 18 hours (Vamos 1963).  Exposure to 0.67 mg/liter NH3




 caused gasping and equilibrium disturbance in 18 min,  frenetic swimming




 activity in 25 min,  then sinking to the tank bottom after 60 min;  after 75




min the fish were placed in ammonia-free water and all  revived.   Similar




                                     15

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effects were observed at a concentration of 0.52 rag/liter NHj (Table 5).




Pre-treating fish orally with 12.5 rag Suprastin (N-dimethyl-arainoethyl-N-p-




chlorobenzyl-o-aminopyridin hydrochlor) , a chemical which reduces cell




membrane permeability, somewhat reduced the toxic effect of ammonia.




     A lethal concentration (Table 5) for carp was reported to be 7.5




rag/liter NHg (Kempinska 1968).  Acute exposures (Table 5) to ammonium




sulfate of bitterling (Rhodeus sericeus) and carp were conducted by Malacca




(1966), who determined minimum lethal concentrations (i.e., after such




exposure, fish placed in ammonia-free water were unable to recover) of 0.76




mg/liter NH3 for bitterling and 1.4 mg/liter NH3 for carp.  Nehring




(1962-63) reported survival times of carp to be 2.4 and 6.0 hours at NH3




concentrations of 9.7 and 2.1 mg/liter N^, respectively (Table 5).




Danecker (1964) reported survival time for tench (Tinea tinea) to be 20 to 24




hours at 2.5 mg/liter NH-} (Table 5).  In a 24-hour exposure of creek chub




(Semotilus atromaculatus) to NlfyOH solution (Gillette et al. 1952), the
"critical range" below which all test fish lived and above which all died was




reported to be 0.26 to 1.2 mg/liter NH3 (Table 5).




     In static exposures lasting 9 to 24 hours, with gradual increases in




NH-j Content, lethal concentrations (Table 5) were determined for oscar




(Astronutus ocellatus) (Magalhaes Bastos 1954); mortalities occurred at 0.50




rag/liter MH-j (4 percent) to 1.8 mg/liter (100 percent).  Tests on oscar of




two different sizes (average weights 1.6 g for "small" fish and 22.5 g for




"medium" fish) showed no difference in susceptibility related to fish size.




A 72-hour LC50 value (Table 5) of 2.85 mg/liter NH^ was reported by Redner




and Stickney (1979) for blue tilapia (Tilapia aurea) .






Factors Affecting Acute Toxicity of Ammonia




     There are a number of factors that can affect the toxicity of ammonia to




aquatic organisms.  These factors include effects of dissolved oxygen




                                     16

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.concentration,  temperature,  pH,  previous  acclimation  to  ammonia,  fluctuating




or  intermittent  exposures, carbon dioxide  concentration,  salinity,  and




presence  of  other  toxicants.  Almost  all  studies  of factors  affecting ammonia




toxicity  have been carried out using  only  acute exposures.




(a)  Dissolved Oxygen




     A decrease  in dissolved oxygen concentration  in  the  water  can  increase




ammonia toxicity.  Vamos and Tasnadi  (1967) observed  mortalities  in  carp




ponds at  ammonia concentrations  lower than would normally be  lethal, and




attributed this  to periodic  low  concentrations of  oxygen.  Based  on  research




in  warmwater (20-22 C) fish  ponds, Sel^si  and Vamos (1976) projected a




"lethal line" relating acute ammonia  toxicity and  dissolved oxygen, below




which carp died.  The line ran between 0.2 mg/liter Nltj at 5 mg/liter




dissolved oxygen and 1.2 mg/liter NH3 at 10 mg/liter  dissolved  oxygen.




Thurston  et al.  (in press, a) compared the acute toxicity of ammonia to




fathead minnows at reduced and normal dissolved oxygen concentrations; seven




96-hour tests were conducted within the range 2.6  to  4.9 mg/liter dissolved




oxygen, and three between 8.7 and 8.9 mg/liter.  There was a slight positive




trend between 96-hour LC50 and dissolved oxygen, although it was not shown to




be  statistically significant.




     Alabaster et al. (1979) tested Atlantic salmon smolts in both fresh



water and 30 percent salt water at 9.6-9.5 and 3.5-3.1 mg/liter dissolved




oxygen.  The reported 24-hour LC50 values at the higher oxygen  concentrations




were about twice that at the lower.




     Several studies have been reported on rainbow trout.  Allan (1955)




reported that below 0.12 mg/liter NH3 and at about 30  percent oxygen




saturation,  the median survival  time was greater than  24 hours, but at the




same Concentration with oxygen saturation below 30 percent, the median




                                     17

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survival time was less than 24 hours.  Downing and Merkens (1955)  tested




fingerling rainbow trout at three different concentrations of NH-j  at  five




different levels of dissolved oxygen.  They reported, in tests lasting up  to




17 hours, that decreasing the oxygen from 8.5 to 1.5 mg/liter shortened the




periods of survival at all ammonia concentrations, and that a decrease in




survival time produced by a given decrease in oxygen was greatest  in  the




lowest concentration of NH^.  Merkens and Downing (1957), in tests which




lasted up to 13 days, also reported that the effect of low concentrations  of




dissolved oxygen on the survival of rainbow trout was more pronounced at low




concentrations of NH3'  Lloyd (1961a) found NHo to be up to 2.5 times




more toxic when dissolved oxygen concentration was reduced from 100 to about




40 percent saturation.  Danecker (1964) reported that the toxicity of ammonia




increased rapidly when the oxygen concentration decreased below two-thirds of




the saturation value.




     Thurston et al. (1981b) conducted 15 96-hour acute toxicity tests with




rainbow trout over the dissolved oxygen range 2.6 to 8.6 mg/liter.  They




reported a positive linear correlation between 96-hour LC50 and dissolved




oxygen over the entire range tested.




     Herbert (1956) reported on rainbow trout mortalities in a channel




receiving sewage discharge containing 0.05 to 0.06 mg/liter NH^.  They




found that at 25-35 percent dissolved oxygen saturation more than 50 percent




of the fish died within 24 hours, compared with 50 percent mortality of test




fish in the laboratory at 15 percent dissolved oxygen saturation.  The




difference was attributed to unfavorable water conditions below the sewage




outflow, including ammonia, which increased the sensitivity of the fish to




the lack of oxygen.
                                     18

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      There  is  a  reduction  in  fish  blood  oxygen-carrying capacity following




ammonia exposure  (Brockway 1950; Danecker 1964;  Reichenbach-Klinke 1967;




Korting I969a,b;  Waluga  and Flis 1971).   Hypoxia would further exacerbate




problems of oxygen delivery and could  lead to  the early demise of  the fish.




(b)   Temperature




      Information  in  the  literature  on  the effects of  temperature on ammonia




toxicity is varied.  The concentration of NH3  increases with increasing




temperature.   Several researchers have reported  an effect  of temperature  on




the toxicity of the  un-ionized ammonia species,  independent  of the  effect of




temperature on the aqueous ammonia  equilibrium.




      Hazel et  al. (1971) tested ammonia  with striped bass  (Morone  saxatilis)




and stickleback (Gas teros teus aculeatus)  and found  little  difference in




toxicity between  15  and 23 C in fresh water, although  both fish species were




slightly more resistant at the lower temperature;  the  influence of




temperature on toxicity was less for striped bass  than for sticklebacks.




McCay and Vars (1931) reported that it took three  times  as long for  brown




bullheads (Ictalurus nebulosus) to  succumb  to  the  toxicity of  ammonia in




water at 10-13 C  than at 26 C.  The pH of  the  tested water was  not  reported;




however, within the  probable range  tested  (pH  7-8), the  percent NH-j  at  the




higher test temperature is approximately  three times that  at  the mean lower




temperature.  Powers (1920) reported the  toxicity  of ammonium  chloride  to




goldfish,  bluntnose minnow, and straw-colored minnow to be greater  at high




temperatures than at low; however,   in that study also  no consideration  was




given to the increase in relative concentration of NH-j  as  temperature




increased.




     Thurston et al.  (in press, a)   reported that  the acute toxicity  of  NH^




to fathead  minnows decreased with a rise  in temperature over  the range  12 to




                                     19

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22 C.  Bluegill and fathead minnow were tested at low and high  temperatures




of 4.0 to 4.6 C and 23.9 to 25.2 C, respectively; rainbow trout were  tested




at 3.0 and 14.0 C (Reinbold and Pescitelli 1982b).  All three species were




more sensitive to un-ionized ammonia at the low temperatures, with  toxicity




being 1.5 to 5 times greater in the colder water; bluegill appeared to be the




most sensitive of the three species to the effect of low temperature  on




ammonia toxicity.




     Colt and Tchobanoglous (1976) reported that the toxicity of NH3  to




channel catfish decreased with increasing temperature over the range  22 to 30




C.  LC50 values for bluegill, channel catfish, and largemouth bass  at 28 to




30 C were approximately twice that at 22 C (Roseboom and Richey 1977).  LC50




values for channel catfish tested in Iowa River water were 0.49 mg/liter




NH3 at 2.5 C and 0.56 mg/liter at 5.1 C (Miller and UNLV-EPA 1982).  An




effluent containing ammonia as a principal toxic component showed a marked




decrease in toxicity to channel catfish over the temperature range 4.6 to




21.3 C (Gary 1976).




     Herbert (1962) has reported that experiments with rainbow trout  in his




laboratory suggest that the effect of temperature on their susceptibility to




NH-j toxicity is little if at all affected by temperature change; no details




were provided.  The Ministry of Technology, U.K. (1968), however, has




reported that the toxicity of NH^ to rainbow trout was much greater at 5 C




than at 18 C.  Brown (1968) reported that the 48-hour LC50 for rainbow trout




increased with an increase in temperature over the range 3 to 18 C; the




reported Increase in tolerance between -12 to ~18 C was considerably  less




than that between ~3 to -12 C.  Thurston and Russo (in press) reported a




relationship between temperature and 96-hour LC50 for rainbow trout over the




                                     20

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 temperature  range  12  to  19  C;  ammonia  toxicity decreased with increasing




 temperature.




     Lloyd and Orr  (1969) investigated  the  effect  of  temperature over the




 range  10-20  C on urine flow rates  of rainbow  trout  exposed  to 0.30 mg/liter




 NH^, and  found no apparent  temperature  effect  on the  total  diuretic




 response  of  the fish, although  the relative increase  in  urine production was




 less at higher temperatures.  From a study  of  the behavioral  response of




 bluegill  to  gradients of ammonia chloride it was hypothesized that  low




 temperatures increased the  sensitivity  of bluegill  and interfered  with their




 ability either to detect ammonia after  a certain period  of  exposure  or to




 compensate behaviorally  for  physiological stress caused  by  ammonia  gradients




 (Lubinski 1979; Lubinski et  al. 1980).




     The  European Inland Fisheries Advisory Commission (1970)  has  cautioned




 that at temperatures below  5 C  the toxic effects of un-ionized ammonia may be




 greater than above 5 C.  The basis for  such a  statement  is  not clearly




 documented in that report.  Nevertheless, there is some merit  to the  argument




 that a decrease in temperature may increase the susceptibility of  fish to




 un-ionized ammonia toxicity.  It is important  that this  relationship  be




 further studied.   The available evidence that  temperature,  independent of its




 role in the aqueous ammonia equilibrium, affects the  toxicity  of NH3  to




 fishes argues for further consideration of the temperature/ammonia toxicity




 relationship.




 (c)  PH




     Tha  roxicity to fishes of aqueous solutions of ammonia and ammonium




compounds  has been attributed to the  un-ionized (undissociated) ammonia




present in the solution.   Although there were observations  in  the early




literature that  ammonia  toxicity was  greater in alkaline solutions, the




                                     21

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 earliest  reported thorough study of the pH dependence of ammonia  toxicity was




 that of Chipman  (1934).  He concluded from experiments with goldfish,




 amphipods, and cladocerans that the toxicity was a function of pll and




 therefore of the concentration of undissociated ammonia in the solution.




     Wuhrmann et al. (1947) discussed the importance of differentiating




between NH^ and NH^  when considering ammonia toxicity.  They




 summarized some unpublished experimental data indicating a correlation




between solution pH and ammonia toxicity to fish (indicated by persistent




 loss of balance).  Wuhrmann and Woker (1948) reported on the experiments




 referred  to in Wuhrmann et al. (1947); these were conducted using ammonium




 sulfate solutions at different pH values on rainbow trout.  Either four or




six fish were tested at each of nine ammonium sulfate concentrations.  The




authors concluded from the experimental results that NH^ was much more




toxic than NH^+.  Downing and Merkens (1955) tested rainbow trout at




different concentrations of ammonia at both pH 7 and 8.  They reported a




consistency of results when ammonia concentration was measured as NH^-




     Tabata (1962) conducted 24-hour tests (Table 5) on ammonia toxicity to




Daphnia (species not specified) and guppy at different pH values and




calculated the relative toxicity of N^/NH^"*" to be 190 for guppy (i.e.,




NH3 190 times more toxic than NH^+) and 48 for Daphnia.  From tests of




the toxicity of ammonium chloride to juvenile coho salmon in flow-through




bioassays within the pH range 7.0 to 8.5, the reported 96-hour LC50 for NH3




was approximately 60 percent less at pH 7.0 than at 8.5 (Robinson-Wilson and




Seim 1975).




     Armstrong et al. (1973) tested the toxicity of ammonium chloride to




larvae of prawn (Macrobrachium rosenbergii) in static six-day tests within




 the pH range 6.8 to 8.3;  test solutions were renewed every 24 hours.  They




                                     22

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'reported a 96-hour LC50 for NH3 at pH 6.83 which was approximately 70




 percent  less  than that  for pH 8.34.   They concluded that the toxicity of




 ammonia  was not  due solely to the NH3 "olecule,  that in solutions of




 different  pH  and equal  NH3 concentrations survival was greatly reduced as




 NH^+  levels increased.   Tomasso et al.  (1980)  tested the toxicity of




 ammonia  at pH 7,  8, and 9  on channel  catfish and reported that 24-hour NH-j




 LC50  values were significantly higher at  pH 8  than at pH 7 or 9.




      Thurston et  al.  (1981c)  tested the toxicity of ammonia to rainbow trout




 and to fathead minnows  in  96-hour flow-through tests at different pH levels




 within the range  6.5  to 9.0.   Results showed that the toxicity of ammonia, in




 terras of NH3>  increased at lower  pH values.  They concluded that  NH^




 exerts some measure of  toxictty,  and/or that increased H* concentration




 increases  the  toxicity  of  NH3 •




      Acute (96-hour)  exposures  of green sunfish  and smallmouth bass  were




 conducted  by  McCorraick  et  al.  (in prep.)  and Broderius  et al.  (in prep.) at




 four  different pH  levels over  the range 6.5  to 8.7.   For both  species,  NH3




 toxicity increased  markedly with  a decrease  in pH,  with LC50 values  at  the




 lowest pH  tested  (6.6 for  sunfish, 6.5  for bass)  being  3.6  (sunfish)  and 2.6




 (bass) times  smaller  than  those at the  highest pH tested (8.7).   LC50 values




 found wltli  rainbow  trout for  the  ammoniacal  portion (diammonium phosphate) of




 a chemical  fire retardant  at  two  different pH  levels  indicated greater  NH3




 toxicity at lower  pH; the  LC50 at  pH  7.0 was 0.15  mg/liter  NH3 and at pH




 R.O was 0.48 rag/liter NH3  (Blahra  1978).




 (d)  Acclimation and Fluctuating  Exposures




     The qviestion of whether fish can acquire  an  increased  tolerance  to




 ammonia by acclimation  to  low ammonia concentrations  is  an  important  one.   If




 fish had an increased ammonia tolerance developed  due  to  acclimation  or




                                     23

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conditioning to low ammonia levels, they would perhaps be able to survive




what otherwise might be acutely lethal ammonia concentrations.




     Observations by McCay and Vars (1931) indicated that bullheads subjected




to several successive exposures to ammonia, alternated with recovery In fresh




water, acquired no immunity from the earlier exposures to the later ones.  A




greater number of researchers have reported that previous exposure of fishes




to low concentrations of ammonia increases their resistance to lethal




concentrations.  Vamos (1963) conducted a single experiment in which carp




which had been revived in fresh water for 12 hours after exposure to 0.67 or




0.52 mg/liter NHg for 75 min were placed in a solution containing 0.7




mg/liter NH3•  The previously exposed fish exhibited symptoms of ammonia




toxicity in 60-85 min, whereas control fish developed symptoms within 20 min.




Redner and Stickney (1979) reported that blue tilapia acclimated for 35 days




to 0.52 to 0.64 mg/liter NH-j Subsequently survived 48 hours at 4.1




mg/liter;  the 48-hour LC50 for unacclimated fish was 2.9 mg/liter.




     Malacea (1968) studied the effect of acclimation of bitterling to




ammonium sulfate solutions.  A group of ten fish was held in an acclimation




solution of 0.26 mg/liter NHg for 94 hours, after which the fish were




exposed to a 5.1 mg/liter NH3 solution for 240 min; a control group of ten




was treated identically, except their acclimation aquarium did not contain




added (MH^^SO^*  The ratio of the mean survival times of "adapted" vs.




"unadapted" fish was 1.13; mean survival times for the adapted and unadapted




fish were  78 and 88 minutes, respectively, indicating somewhat higher ammonia




tolerance  for adapted fish.




     Fromm (1970) measured urea excretion rates of rainbow trout initially




acclimated to either 5 or 0.5 mg/liter NH3ป then subjected to 3 mg/liter
                                     24

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 NH3'   Fish  previously  exposed  to  5 mg/liter  NH3  excreted slightly less




 urea  than those  exposed  to  the lower  concentration.   Lloyd  and Orr (1969)




 conducted acclimation  experiments with  rainbow  trout  and found that  the rate




 of  urine excretion  increased with a rise  in  the  concentration  of  un-ionized




 ammonia  to  which the fish were exposed.   They presented  some evidence for




 acclimation of rainbow trout to sublethal  levels of ammonia, although these




 levels may  be as  low as  12  percent of the  "lethal  threshold concentration".




 Acclimation was  retained for 24 hours, but was not retained after three days.




 They  also suggested that environmental  factors which  affect the water balance




 of  fish may also  influence  susceptibility  to ammonia  toxicity.  Fromm (1970)




 acclimated  goldfish to low  (0.5 mg/liter)  or high  (5.0 or 25.0 rag/liter)




 ambient NH^ for  periods  of  20  to  56 days and found that  urea excretion rate




 in  subsequent 24-hour  exposures to concentrations  ranging from 0.08  to 2.37




 rag/1iter was independent of the previous acclimation  concentration or




 duration.




      Schulze-Wiehenbrauck (1976)  subjected to lethal  ammonia concentrations




 two groups  of rainbow  trout (56 g and 110 g) which had been held  for  at  least




 three weeks at sublethal ammonia  concentrations.  In  the experiment with




 110-g fish,  the sublethal acclimation concentrations were 0.007 (control),




 0.131, and  0.167 mg/liter NH-jJ the fish from these three tanks were  then




 subjected to concentrations of 0.45, 0.42, and 0.47 mg/liter NHg,




 respectively, for 8.5  hours.  Fish from the two higher sublethal  concentra-




 tions had 100 percent  survival after 8.5 hours in the 0.42  and 0.47 mg/liter




NH3 solutions, whereas  fish ,from the 0.007 mg/liter NHj concentration  had




only 50 percent survival in 0.45 mg/liter NH^•  In the experiment  with  56-g




fish,  the acclimation concentrations were 0.004 mg/liter NH3 (control)  and




0.159  nirf/liter NH35  these fish were  placed in NH3 concentrations  of 0.515




                                     25

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and 0.523 mg/llter, respectively, for 10.25 hours.  There was  100 percent




survival of the acclimated fish, and 85 percent survival of  the control  fish.




The results of these experiments thus showed an increase in  resistance of




trout to high ammonia levels after prior exposure to sublethal ammonia




levels.




     Alabaster et al. (1979) determined 24-hour LC50 values  of NH3 for




Atlantic salmon smolts under reduced dissolved oxygen test conditions.   Fish




acclimated to ammonia before oxygen reduction evidenced LC50 values 38 and 79




percent higher than fish without prior ammonia acclimation.




     Brown et al. (1969) tested rainbow trout in static tests  in which fish




were moved back and forth between tanks in which the ammonia concentrations




were 0.5 and 1.5 times a previously determined 48-hour LC50.  If fish were




transferred on an hourly basis, the median period of survival for the




fluctuating exposure was reported to be the same as that for constant




exposure (>700 rain).  If the fish were transferred at two-hour intervals, the




median survival time for the fluctuating exposure was reported to be less




(370 min), indicating that the toxic effects from exposure to the fluctuating




concentrations of ammonia was greater than those from exposure to the




constant concentration.




     Thurston et al. (1981a) conducted acute toxicity tests  on rainbow trout




and cutthroat trout in which fish were exposed to short-term cyclic




fluctuations of ammonia.  Companion tests were also conducted in which test




fish were subjected to ammonia at constant concentrations.  Median lethal




concentration (LC50) values in terms of both average and peak concentrations




of ammonia for the fluctuating concentration tests were compared with LC50




values for the constant concentration tests.  Based on comparisons of total




dose exposure, results showed that fish were more tolerant of constant




concentrations of ammonia than of fluctuating concentrations.  Fish subjected




                                     26

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 to fluctuating concentrations  of  ammonia  at  levels  below those  iicuteiy toxic




 were  subsequently better  able  to  withstand exposure to  higher fluctuating




 concentrations than  fish  not previously so acclimated.




      In  static renewal exposures  to ammonium chloride using  river  water  as




 the dilution water,  fathead minnows were  reported (Pitts 1980)  to  survive for




 28 days  exposures fluctuating  from 0.1 mg/liter NH3 for four days  to  0.2 or




 0.3 mg/liter Nil-} for  three days.  Four-day excursions above  0.1 mg/liter to




 concentrations of 0.42, 0.48,  and 0.52 rag/liter resulted in  80  to  100 percent




 mortality in 28 days, as  did four-day excursions to 0.73 mg/liter.  No




 constant exposure tests were conducted simultaneously for comparative




 purposes; however, constant exposure tests conducted approximately  a  year




 earlier  yielded LC50 values ranging from 0.6 to 2.4 mg/liter NHg •




      In  summary, there is reasonable evidence that  fishes with a history of




 prior exposure to some sublethal concentration of ammonia are better  able to




 withstand an acutely lethal concentration, at least for  some period of hours




 and possibly days.  The relative concentration limits for both acclimation




 and subsequent acute response  need better definition and  a more complete




 explanation.   Limited data on  fluctuating exposures indicate that fish are




 more  susceptible to fluctuating than to constant exposure to equivalent  NH^




 dose concentrations.   Much more research is needed  to examine further  the




 effects of fluctuating and intermittent exposures under  exposure regimes




 simulating actual field situations.




 (e)  Carbon Dioxide




     An increase in carbon dioxide concentrations up to 30 rag/liter decreases




 total  ammonia  toKicity (Alabaster and Herbert 1954;  Allan et al. 1958).




C02 causes a decrease in pH,  thereby decreasing the proportion of




un-ionized ammonia  in solution.  Lloyd  and Herbert  (i960) found, however,




                                     27

-------
that although  total ammonia  toxicity was  reduced  at  elevated  CC>2  levels,




the inverse was  true when considering un-ionized  ammonia  alone; more  NH-j  is




required in low  C(>2ป high pH water to exert  the same toxic  effect as  seen




in fish in high  C02> low pH water.  The explanation  presented by  Lloyd  and




Herbert (1960) for the decreased toxicity of NH3  in  low CC>2 water was




that Cr>2 excretion across the gills would reduce  pH,  and  therefore NH3




concentration, in water flowing over the  gills.




     The basic flaw in Lloyd and Herbert's (1960) hypothesis  has  been




discussed in Broderius et al. (1977).  C02 will only form protons very




slowly in water  at the tested temperature.  The uncatalyzed CC>2 hydration




reaction has a half-time of seconds or even minutes  (e.g.,  at pH  8:   25




seconds at 25 C, 300 seconds at 0 C (Kern I960)), and water does  not  remain




in the opercular cavity for more than a few seconds, and  at the surface of a




gill lamella for about 0.5 to 1 second (Randall 1970; Cameron 1979).  Thus




the liberation of C02 will have little, if any, effect on water pH or,




therefore, NH^ levels while the water body is in  contact  with the gills.




Hence the liberation of C02 across the gills can  have little, if  any,




effect on the NH3 gradient across the gills between  water and blood.




Szumski et al. (1982) hypothesized that in the course of  its  excretion  C02




is converted in  the gill epithelium to H+ and HCO^"  which then pass




directly Into the gill chamber where they cause an instantaneous  pH




reduction.  Their interpretation of the published literature  on fish




respiratory physiology is questionable, and experimental  evidence in  support




of their evaluation is required before it can be  given serious




consideration.
                                     28

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. (f)  Salinity




      Herbert and Shurben (1965) reported that the resistance of yearling




 rainbow trout to ammonium chloride increases with salinity up to levels of




 30-40 percent seawater;  above that level,  resistance appears to decrease.




 Katz and Pierro (1967)  tested fingerling coho salmon at salinity levels of 20




 to 30 parts per thousand (57  to 86 percent saltwater)  and found that toxicity




 of an ammonia-ammonium  waste  increased as  salinity increased.  These findings




 are in agreement at the levels tested with those of Herbert and Shurben




 (1965).   Atlantic salmon were exposed to ammonium chloride solutions for 24




 hours under both freshwater and 30 percent saltwater conditions; LC50 values




 (Table 5) were 0.15 and 0.3 rag/liter NH3,  respectively, in the two




 different waters (Alabaster et al. 1979).




      As  was discussed in Willingham et al. (1979), decreased NH3 toxicity




 with increased salinity may be partially explained,  at least for low salinity




 levels,  by the fact that there is  a slight decrease in the NH^ fraction of




 total ammonia as ionic  strength increases  in dilute saline solutions




 (Thurston et al. 1979).   At higher salinity levels,  however, the toxicity to




 fishes of ammonia solutions must be attributable to some  mechanism or




 mechanisms other than the changes  in the NH^+/NH3 ratio.   Further work




 is needed to confirm results  already reported and to clarify the observed




 mitigating effect of total dissolved solids.




 (g)   Presence of Other  Chemicals




      The  presence of other chemicals may have an effect on ammonia toxicity,




 and  some  experimental work has  investigated this topic.   Herbert and Vandyke




 (1964),  testing rainbow  trout,  determined  the 48-hour  LC50 value for a




 solution  of  ammonium chloride  and  that  for a  solution  of  copper  sulfate.




 They reported that  a solution  containing a mixture of  one half of each of




                                      29

-------
these LC50 concentrations was also the 48-hour LC50 for the two toxicants




combined; i.e., the toxic response was simply additive.  This  information was




also reported by the Ministry of Technology, U.K. (1964);  it is not clear




whether this was a separate study or the same study.




     Shemchuk (1971) measured copper uptake in two-year old carp from




solutions of Cu(NH-j)^^ ; copper uptake in various fish tissues was




reported, but no information was provided about toxicity.  Vamos and Tasnadi




(1967) applied cupric sulfate to a "carp pond" to reduce the concentration  of




free ammonia and reported that this measure proved successful  to reduce  the




toxic effect of ammonia; few details were provided.




     Ministry of Technology, U.K. (1962, 1963) reported on the results of




tests on rainbow trout in which 48-hour LC50 concentrations were determined




for solutions of ammonium chloride, zinc sulfate and mixtures of these two




salts.  A fraction of each of those 48-hour LC50 concentrations, when




combined in such a way that those fractions equalled unity, provided a




mixture with a 48-hour LC50 concentration equal to that of either of the two




toxicants alone.  Results were similar for tests conducted in waters with




alkalinities of 240 and 50 mg/liter CaC03.




     Herbert (1962) studied the toxicity to rainbow trout of ammonia-phenol




mixtures.  The mixtures contained fractions of the 48-hour LC50




concentrations of phenol and of ammonia; the combined fractions equaled




unity.  The toxicity of the combined fractions approximated the toxicity of




either phenol or ammonia when tested separately but under test conditions of




similar water chemical characteristics.  The same information was reported by




Ministry of Technology, U.K. (1961);  it is not clear whether this was a




separate study or the same study.
                                     30

-------
      Brown  et  al.  (1969)  conducted  48-hour  tests  on rainbow trout in mixtures

of anraonia, sine,  and phenol;  the mixture coi\tainปd equal  portions,  by

48-hour LC50 concentration, of  the  three toxicants. They  reported that  each


chemical nominally contributed  equally to the  toxicity.  In a  second series

of three tests in  which the mixture was adjusted  to include approximately 75


percent of a 48-hour LC50 concentration of  one toxicant and the balance  split


equally between  the other two,  they reported that  the  principal toxicant

contributed about  three-fourths of the toxicity.

      Broderius and Smith (1979), in 96-hour flow-through tests with  rainbow

trout, reported  a  synergistic effect for NH3 and HCN except at extremely

low concentrations.  Rubin and  Elmaraghy (1976, 1977)  estimated the


individual and joint toxicities of ammonia  and nitrate to  guppy fry;  the


toxicities of the  two in mixture were additive, except at  very low

ammonia-to-nitrate ratios.  Tomasso et al.  (1980)  reported  that elevated


calcium levels increased the tolerance to ammonia  of channel catfish.



Derivation of Maximum Criterion for Fresh Water


(a) pH Dependence  of Acute Ammonia Toxicity

     The acute toxicity of total ammonia to aquatic  organisms has been found

to increase markedly with pH, with LC50s declining by  as much as an  order  of

magnitude per unit pll increase.  Much of this  variation can be accounted for

by expressing toxicity on the basis of un-ionized  ammonia  concentration

rather than total  ammonia concentration, un-ionized  ammonia being considered

the principal source of toxicity.  Although such a mode of  expression

substantially reduces the variation with pH, available data  indicate  that  a

residual dependence exists, with NH^ LCSOs  increasing with  increasing  pH

and the rate of increase declining as pH increases, apparently leveling  off


at high pH.
                                     31

-------
     The task at hand was to select a suitable mathematical  expression  to




describe the pH dependence of LC50s, on an un-ionized ammonia basis,  in the




pH 6.S-9.0 range (this range being dictated by the water  quality  criterion




for pH (U.S. Environmental Protection Agency 1977)) and to determine  whether




the available data support the use of a common (i.e., for all species)




expression that has only a single species-dependent parameter.  The data used




(Figure 1) were for Daphnia sp., rainbow trout, fathead minnow, and coho




salmon (Tabata 1962; Robinson-Wilson and Seim 1975; Thurston et al. 1981c).




This includes all available sets for which, for a given species,  the  same




investigators determined LC50s at a minimum of four pHs distributed over at




least 1.5 pH units in the pH 6.5-9.0 range.  Furthermore, a  minimum of  six




data points (either by testing at more distinct pHs or by replication at some




pHs) was required so that meaningful regression analysis was possible.




     Inspection of the data suggested that, for all data sets and for the  pH




range of concern, log(LCSO) versus pH has a slope that decreases  with




Increasing pH, nearing zero at the upper part of the range and perhaps




approaching a limiting value at the lower part of the range.  A suitable




mathematical expression for such behavior is:


                                  	XI

                           L.C50 =       v, / v~  H7                        (1)
                                  I + 10X3(X2-pH)                        v '





where XI is the limiting LC50 at high pH as the slope approaches  zero,




      X3 is the limiting slope of log(LCSO) vs. pH at low pH,




      X2 allows the intermediate portion between the limits  to occur  at  a




         variable pH.




This equation is not only consistent with the form of the data, but also with




suggestions that the p!l dependence of ammonia toxicity is due to  both
                                     32

-------
   5.0
I
    i.o
a
    0.5
            7.0
      OAPHNIA
    8.0

  PH
9.0
                                              0.5
                                            I

                                            IO
                             0.2
              o
              10

              So.,
                                             0-05L_i—^
          RAINBOW

           TROUT
7.0      8.0      9.0

       pH
    I.Or
^ 0.5
 10
 ^0.2
    O.I
O
  0.05
                              2.0r
            7.0
       COHO

      SALMON
    8.0
9.0
                   PH
                              i.o
                                            10
                           O>
             O
             in
             o
                0.2
                             O.I
          FATHEAD

          MINNOW
7.0
8.0
9.O
                                             PH
       Figure 1.
Acute NH3 toAicity  at different pH values (data from Taoata 1962,

Robinson-Wilson and Seim 1975,  Thurston et al. 1981c).   Dotted

lines ซ  regression  based on individual data set; solid lines =

regression based on pooled data sets.


                  33

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un-ionized and cationic ammonia being toxic, but with different potencies



(Armstrong et al. 1978; Thurston et al. 1981c).



     Estimates for XI, X2, and X3 were obtained (Erickson 1982) for each of
    *


the available data sets by least-squares, nonlinear regression, using



log(LC50) as the dependent variable and pH as the independent variable.  A



logarithm transform was employed because the errors in LC50 estimates are



largely proportional to LC50.



     The results of these analyses indicated X2 and X3 did not differ



significantly (at the 0.10 confidence level) between data sets and were in



fact quite similar (X2 ป 7.3-7.7, X3 = 0.7-1.2).  The data sets were



therefore pooled for further regression analysis in which X2 and X3 were



assumed to be the same for all sets.  Deviations of this pooled analysis from



the individual regressions (Figure 1) were not statistically significant and



were, in fact, quite minor, the maximum deviation being less than 0.08 log



units and the average being less than 0.03 log units.  The individual



regressions accounted for 89%, 96%, 99%, and 96% of the variance of log(LCSO)



in the Daphnia, rainbow trout, coho salmon, and fathead minnow data sets,



respectively.  Respective percentages for the pooled analysis were 87%, 92%,



99%, and 92%.



     It was therefore concluded that the common values for X2 (7.32) and X3



(1.03) from the pooled analysis could be justifiably applied to these data



sets.  Equation 1 can then be interpreted as having a data set-dependent



numerator, the limiting LC50 at high pH (XI), and a set-independent



denominator,  a multiplicative, pH correction factor (1 + 101>^3(7.32-pH)^.



     Four other d.-ita sets were available for which acute ammonia toxicity to



a species was tested by the same investigators at three or more distinct pHs



distributed over at Least 1.5 pH units in the pH 6.5-9.0 range.  These sets



                                     34

-------
• .were for Macrobrachium rosenbergii (a saltwater prawn), channel  catfish,


 green sunfish, and sraallmouth bass (Armstrong et al.  1978; Tomasso  et  al.


 1980; McCormick et al., in prep.; Broderius et al., in prep.).   These  data


 sets were not used in the previously-discussed analysis due  to inadequate


 size, but rather were used to evaluate the applicability of  the  relationship


 derived above (Erickson 1982) .  A line was fit to each set by taking the


 above common estimates for X2 and X3 and calculating  XI as the geometric


 average of LC50 x (1 -I- 101 -03(7 -32-pH)) for the individual data  (this


 minimizes the variance of the residuals between the observed log (LC50)s and


 the fitted line).  Two of the sets (prawn, green sunfish) showed a  marked pH


 dependence qualitatively identical with the previous  discussion  and had good

                                                       P
 quantitative agreement with the lines so estimated (r  = 99% and 83%,


 respectively).


      Of the other two sets,  one (smallmouth bass) showed a marked increase of


 LCSOs with pH, but with a noticeably different shape  in the  pH range of


 Interest;  the other set (channel catfish) showed no apparent pH  trend  at all.

                                                                        2
 Neither of the lines fitted  to these two sets produced a significant r ,


 but they also did not deviate enough from the data to reject statistically


 either  the general mathematical model or the common estimates for X2 and X3.


 Furthermore,  the residuals between the data of these two sets and the  fitted


 lines were at most 0.3 log units  (a factor of 2)  and had standard deviations


 of only 0.13  log units (a factor  of 1.5).  Additionally,  if the common


 estimates  of  X2  and X3 are used to correct an LC50 in these sets from  the pH


 it was  measured  at to a pH of another datum in the set,  the discrepancy


 between the two  LC50s are lar^e (>0.2 log units)  only if  the pH difference is


 large (>1  unit)  and  one  of the  pHs is <7.5.

-------
     Available acute toxiclty data therefore suggest that the proposed




relationship will perform well in most circumstances and will produce large




errors only under restricted circumstances.  Although the proposed




relationship cannot be considered universally applicable, especially if




residuals between fitted lines and data are not to exceed measurement error,




the alternatives of using no pH relationship or of basing criteria only on




species tested over a range of pHs are clearly less desirable.




     Deviations of some data sets from the proposed relationship are probably




due to a combination of experimental error, species effects, and temperature




effects.  The relative contributions of each source are uncertain.  It may be




possible in the future to develop parameter estimates that better account for




variations due to temperature and species, but the current data base will not




support such an exercise.  Of course, in site-specific applications, if




evidence exists for significantly different pH relationships for species of




importance to setting criteria, appropriate modifications should be




considered.




(b) Temperature Dependence of Acute Ammonia Toxicity




     Investigations of the temperature dependence of LCSOs on an un-ionized




ammonia basis have produced conflicting results, varying from constancy with




temperature to major increases with temperature (see earlier discussion).




The relative contributions to this variability of species differences,




experimental error, and the relationship of test temperatures to species




temperature tolerance are uncertain and not definable from available data.




     Lacking a broadly-applicable description of the temperature dependence




of ammonia toxicity, it will be assumed that, above 10 C, LC50s on an




un-ionized ammonia basis are constant with temperature.  Where LC50s do
                                     36

-------
 increase with  temperature,  this  assumption will  result  in overprotection at




 temperatures above  that  at  which data  most important  to the criterion were




 gathered and underprotection at  lower  temperatures.   Because data  for




 criteria were  mostly  collected in the  10-25 C  range,  large errors  probably




 will not accumulate over  this range, due  to limited extrapolation  from tested




 temperatures.  Extension  of such an assumption to  low temperatures  is




 questionable,  however, because the paucity of  the  data  in this  range  makes




 such a move a  substantial extrapolation and because the temperature effects




 data presented earlier, unlike at higher  temperature, consistently  indicated




 increased  sensitivity associated with  low temperatures.   At  temperatures  less




 than 10 C, it will be assumed that LC50s  on a  total ammonia  basis are




 constant.  Due to the effect of  temperature on




 the pK of  ammonia, from 10 C to  0 C this  is equivalent  to approximately  a




 two-fold lowering of  the  LC50s on an un-ionized ammonia  basis.  Such  a




 lowering is not inconsistent with temperature  effects discussed earlier and




 constitutes a very simple algorithm.




     Of course, as for pll, where data  for  a species of  importance to  the




 setting of a criterion value contradicts  the above assumptions  regarding




 temperature, appropriate  modifications should be made.




 (c) Application of pH and Temperature Relationships of  Acute Ammonia  Toxicity




    to Determination of Final Acute Values




     A Species Mean Acute Value  (SMAV) Is  the geometric  average of  the acute




values (AVs), usually LC50s, available for  a given species.  A  Family Mean




Acute Value (FMAV) is the geometric average of the SMAVs  available  for a




given family.  A Final Acute Value (FAV) for a material  is an estimate of  the

-------
FMAV at  the 0.05 cumulative proportion in  the cumulative  distribution  of

FHAVs for all families tested for that material.  These computations (see

Guidelines) are not a subject of this discussion, but  their application to  pH

and temperature dependent data is.

     The existence of pH and temperature dependence in AVs requires that they

be corrected to a common pH and temperature basis before  computing a FAV.

After a FAV at this common pH and temperature is computed, it can be applied

to other pHs and temperatures using the same equations used to correct the

AVs.

     Since AVs on an un-ionized ammonia basis are assumed constant with

temperature above 10 C, the "common" temperature can be the whole temperature

range XIO C.  Only AVs measured at <10 C must be temperature corrected, to  10 C,

based on the assumption that AVs on a total ammonia basis are constant

with temperature below 10 C.  The temperature correction  equation for AVs

from <10 C is therefore:
                                 (1 + ioPKT-pH)
                            10   (1 + 10PK10-PH)    T


where AV-p is an AV measured at a temperature T <10 C, AV^Q is the

estimated AV at temperatures >10 C, pH is the pH at which the AV was

measured, pK-p is the pK of ammonia at the measurement temperature (=0.902 +

         ) , and pK^Q is the pK at 10 C (=9.73) (pK/temperature relationship
,- • ?7'-- _y

from Emerson et al. 1975).

    After any necessary temperature corrections are made, Equation 1, with

the common estimates for X2 and X3, can be applied for correction to a common

pH.  Equation 1 has a single species-dependent parameter, XI, the limiting AV

at high pH, hereafter called LIMAV.  This limiting value will be considered

                                     38

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 the  common pH  value.   All available  AV^gS  will  therefore be converted to

 LIMAVs  by  the  relationship:

                    LIMAV - AV1Q  x (1 +  io1.03(7.32-pH))

      The limiting  SMAV (LIMSMAV)  for a species  then  can be  computed as the

 geometric  average  of  the  LIMAVs for  that species,  and  the limiting  FMAV

 (1IMFMAV)  can  be computed as  the  geometric average of  the LIMSMAVs  for that

 family.  The limiting  FAV (LIMFAV) for un-ionized  ammonia then  can  be

 computed from  the  LIMFMAVs available, by the  same  procedures  used for

 computing  FAVs from FMAVs for any material.   A  FAV on  an un-ionized ammonia

 basis at a  particular  pH  for  temperatures XLO C can be  computed  as:
                                  _ LIMFAV _
                          FAV10 =  l + 101.03(7.32-PH)                      <4>

 A FAV on an un-ionized ammonia basis at a temperature  T < 10  C  can  be

 computed from  the  >10  C FAV at the same pH as follows:
                                     (I + l09.73-pH)
                                                                          (5)
     Application of these techniques to the data proceeded as follows.  AVs

from Table 1 were corrected for temperature (where necessary) and  pH and

averaged to obtain the LIMSMAVs and LIMFMAVs reported in Table 3.  The  fifth

percentile was estimated, by the Guidelines method, to be 0.74 mg/liter

NH3 •  This number, however, exceeds the LIMSMAVs of important species

(walleye and rainbow trout) and thus, by Guidelines procedures, should be

decreased to the lower of these two LIMSMAVs.  Additionally, the rainbow

trout data in Table 1 indicate that sexually mature fish (M kg) are

significantly more sensitive than the average of the tested fish.  Since a

species is not protected if such an important life stage is not protected,

the LIMFAV was lowered to 0.30, the geometric average of the LIMAVs of

rainbow trout in this size range.   Thus, for temperatures >10 C:

                                     39

-------
                                        0.30
                            3    (1 +  101.03(7.32-pII))

 For  temperatures  <10 C:
                       	0.30	    (1 +  109-73-pH)
               FAVl "  (1 +  K)l-03(7.32-pH)) x  (1 4
     Available Information indicates that  the acute response  to ammonia  is

rapid.  Therefore, a criterion based on the FAV cannot be  treated  as  an

average over any appreciable span of time, since such averaging implicitly

allows significant excursions over the criterion value for an appreciable

fraction of the averaging period and thus  allows the occurrence of a  time

sequence of concentrations which would violate the intent of  the criterion.

Consequently, the criterion based on the FAV is a concentration that  at  no

time should be exceeded.


Saltwater Invertebrates

     Data on acute toxicity of ammonia to  saltwater invertebrate species are

very limited.  LC50 values are summarized  in Table 1 for five species

representing five families.  A 96-hour LC50 value (Table 1) of 1.5 mg/liter

NH3 was reported (Linden et al. 1979) for  the copepod, Nitocra spinipes.

Lethal effects of NH^Ci on the quahog clam (Mercenaria mercenaria) and

eastern oyster (Crassostrea virginica) were studied by Epifanio and Srna

(1975) (Table 1).  There was no observed difference in susceptibilities

between juveniles and adults of the two species.  Armstrong et al. (1978)

conducted acute toxicity tests (6 days) on ammonium chloride using prawn

larvae (Macrobrachium rosenbergii).  LC50  results (Tables 1, 5) were highly

pH-dependent.  Acute toxicity of NH^Cl to  penaeid shrimp was reported as a

48-hour composite LC50 value of 1.6 mg/liter NH^ for seven species pooled,

including the resident species Penaeus setiferus (Wickins 1976).  The acute

                                     40

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' toxicity of NH^Cl to the carldean prawn, >1. rosenbergii, was reported




 (Wicklns 1976) as LT50 values of 1700-560 minutes at concentrations of 1.74




 to 3.41 mg/liter NH3 (Table 5).  Hall et al. (1978) measured the acute




 toxicity of NH^Cl to grass shrimp (Palaemonetes pugio) (Table 5).  Catedral




 and coworkers (1977a,b) investigated the effect of NH^Cl on survival and




 growth of Penaeus monodon;  larvae had lower tolerance to ammonia compared




 with postlarvae.  Brown (1974) reported a time to 50 percent mortality of 106




 min for neraertine worm (Cerebratulus fuscus) at 2.3 mg/liter NH3 (Table 5).




      Effects of NH^Cl solutions on American lobster (Homarus americanus)




 were studied by Delistraty et al. (1977).  Their tests were performed on




 fourth stage larvae which they believed to be the most sensitive life stage,




 or nearly so.   They reported a 96-hour LC50 value (Table 1) of 2.2 mg/liter




 NH3 and an incipient LC50 (Table 5)  of 1.7 mg/liter NH3.  A "safe"




 concentration  of 0.17 mg/liter NH3 was tentatively recommended.







 Saltwater Fishes




      Very few  acute toxictty data are available for saltwater fish species.




 Holland et  al.  (1960) reported the critical level for chinook salmon




 (Oncorhynchus  tshawytscha)  to be between 0.04 and 0.11  mg/liter  NH3 and for




 coho salmon to  be 0.134 rag/liter NH3ซ  A static test with  coho  salmon




 provided a  48-hour LC50 value (Table 5)  of 0.50 mg/liter NH3 (Katz and




 Pierro 1967).   Atlantic salmon smolts and yearling rainbow trout tested for




 24 hours in 50  and 75 percent saltwater  solutions exhibited similar




 sensitivities to  ammonia (Ministry of Technology,  U.K.  1963).
                                     41

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                         CHRONIC TOXICITY TO ANIMALS




     The following discussion of chronic and partial chronic ammonia toxicity




includes both data used in the derivation of the numerical 30-day average




criterion (Table 2 data) and data that were not included in the criterion




derivation, but that are important for an understanding of long-terra lethal




and sublethal effects of ammonia on aquatic organisms (Table 5 data).







Freshwater Invertebrates




     Few studies have been conducted on long-term exposure of freshwater




invertebrates to ammonia, and life-cycle tests were conducted only for




cladocerans.




     The lowest concentrations affecting reproduction in two life-cycle tests




(Table 2) with ID. magna were 0.74 and 0.76 mg/liter NH-j (Russo et al., in




prep.);  a 28-day LC50 of 1.53 mg/liter NH^ was reported.  In a chronic test




(Table 2) conducted by Reinbold and Pescitelli (1982a), reproduction and




growth of I), magna were affected at a concentration of 1.6 mg/liter NH^-  A




life-cycle test (Table 2) with ฃ. acanthina (Mount 1982) showed effects on




reproduction at a concentration of 0.463 mg/liter NH3•




     Two tests lasting 42 days were conducted by Anderson et al. (1978) on




Nl^Cl with the fingernail clam, Musculium transversum (Table 5).




Significant mortalities (67 and 72 percent) occurred in both tests at a




concentration of 0.7 mg/liter NH^.  In one of the experiments, significant




reduction in growth was observed after 14 days of exposure to 0.41 mg/liter




NH3•  Sparks and Sandusky (1981) reported that fingernail clams exposed to




0.23 and 0.63 mg/liter NH3 incurred 36 and 23 percent mortality,




respectively, In four weeks;  after six weeks, 47 percent mortality occurred




at 0.073 rig/liter NH^, and 83 percent mortality occurred at 0.23 and 0.63






                                     42

-------
mg/liter NH-^-  No growth  at  All  occurred  In  all  test  chambers  (conoontra-




tions of 0.036 mg/liter NH3  and  higher) other  than  the  control after  six




weeks (Table  5).




     Two partial chronic  tests,  of 24- and 30-days' duration,  were  conducted




by Thurston et al.  (in prep.,  a) with the stonefly  Pteronarcella badia  (Table




5).  Adult stonefly emergence  was delayed with increasing  ammonia concentra-




tion, and little or no emergence occurred at concentrations exceeding 3.4




mg/liter NH-j.  There was  no  significant relationship  between food




consumption rates of nymphs  and  concentrations up to  6.9 mg/liter NH-j •




LC50 values for 24- and 30-day exposures were 1.45  and  4.57 mg/liter NH3,




respectively.







Freshwater Fishes




     A number of researchers have conducted  long-term ammonia  exposures  to




fishes, including complete life-cycle tests  on rainbow  trout and fathead




minnows.  Several  '<1nds of endpoints have been studied, including effects on




spawning and egg incubation, growth, survival, and  tissues.




     The effects of prolonged  exposure (up to 61 days)  to  ammonia of pink




salmon early life stages was studied by Rice and Bailey (1930).  Three series




of exposures were carried out, beginning at  selected  times after hatching:




for 21 days prior to completion of yolk absorption, for 40 days up  to 21 days




before yolk absorption, and  for 61 days up to yolk  absorption.  All test fish




were sampled for size when the controls had  completed yolk absorption.  NHg




concentrations ranged from 0 (control) up to 0.004 mg/liter.   For fry at the




highest concentration of 0.004 mg/liter NH3  (Table 2),  significant




decreases in weight were observed for all three exposure groups.  At a




concentration of 0.0024 mg/liter NH3 (Table 2) the group of fry exposed for




40 and 61  dayu were significantly smaller, whereas a concentration of 0.0012




                                     43

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mg/liter had no significant effect on growth.  Effects  were  consistently more




adverse for the 61-day-exposed fish.




     Thurston et al. (In press, b) tested rainbow  trout in a laboratory study




In which adult fish exposed for five months to concentrations of  ammonia from




0.01 to 0.07 mg/llter NH^ Spawned of their own volition; baskets  containing




crushed rock served as the spawning substrate.  There was no correlation




between ammonia concentration and numbers of egg lots spawned,  total  numbers




of eggs produced, or numbers of eggs subsequently  hatched.   Parental  fish




were exposed for 11 months, the first filial generation (F^) for  four




years, and the second filial generation (F2) for five months.   Pathologic




lesions were observed in both parental and F^ fish when ammonia concentra-




tions reached and exceeded 0.04 mg/liter NH^ (Table 2).  Measurements of




blood ammonia concentrations in four-year-old F^ fish showed an increase




when test water conditions reached or exceeded 0.04 mg/liter NH-j •  Trout




exposed for 52 months from day of hatching showed  no dose/growth  relationship




,it 10, 15, 21, and 5?. months.




     Burkhalter and Kaya (1977) tested ammonia at  concentrations  from 0.06  to




0.45 mg/liter NH-j On fertilised eggs 
-------
. after hatching, hypertrophy of secondary gill lamellae epithelium occurred at




 0.23 rag/liter NH3, and karyolysis and karyorrhexis in the secondary gill




 lamellae were observed after 28 days at 0.34 mg/liter NH3 and higher.




      CalaraarL et al. (1977, 1981) exposed rainbow trout to ammonium chloride




 solutions for 72 days, beginning one day after fertilization and ending when




 fry were fed for 30 days.  A 72-day LC50 of 0.056 mg/liter NH3 was




 calculated (Table 5); 23 percent mortality occurred at a concentration of




 0.025 rag/liter NH3 (Table 2).  Examination of 986 rainbow trout embryos at




 hatching stage after exposure to NH3 Concentrations of 0.010 to 0.193




 mg/liter for 24 days showed an increase in macroscopic malformations with




 increasing ammonia concentration.  Kinds of deformities observed were varying




 degree of curvature from median body axis,  which in extreme cases produced a




 complete spiral shape, and various kinds of malformations in the head region




 with a number of cases of double heads.  At the highest concentration tested,




 0.193 mg/liter NH-j > 60 percent of the observed fish were malformed.




 Microscopic  examination at hatching of 128  larvae from the same exposure




 showed abnormalities on the epidermis and pronephros  that correlated with




 ammonia concentrations.  The epidermis was  thickened  with an irregular




 arrangement  of the various layers of cells  and an increase in the number and




 dimensions of mucous cells.  The pronephros showed widespread vacuolization




 of the tubule cells, together with a thickening of the wall.  Increasing




 abnormalitLes were observed after exposure  to concentrations over 0.025




 mg/liter NH3 for epidermis and 0.063 mg/liter NH3 for pronephros.




      Broderius and Smith (1979) tested four-week-old  rainbow trout fry for 30




 days at concentrations of ammonia (reported grahically)  ranging from -0.06 to




 0.32 rag/liter NH3  (Table 5).   Growth rate at -0.06 mg/liter  NH3 was
                                      45

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comparable to that of controls; above -0.10 mg/liter NH-j growth  rate




decreased, correlated with Increased NH3 concentration.  The survival  at




0.32 mg/liter NHg was reduced to 70 percent that of the controls.  Schulze-




Wiehenbrauck (1976) tested juvenile rainbow trout, approximately




one-half-year-old but of different sizes, for periods of time from two  to




seven weeks, and at ammonia concentrations from 0.012 to 0.17 mg/liter  NH3 •




He concluded that 0.05 mg/liter NHj caused a slight decrease in growth




during the first 14-day interval on nonacclimatized fish, but that decrease




was completely compensated in the next growth interval; exposure to 0.13




mg/liter NH-j (apparently for 3 or A weeks) did not affect growth, food




consumption, or food conversion.




     Smith (1972) and Smith and Piper (1975) reared young rainbow trout at




three concentrations of ammonia (averaging 0.006, 0.012, and 0.017 mg/liter




NH^) for a period of one year.  There was no significant difference in  fish




growth reported among the three concentrations at four months.  There was,




however, a difference reported at 11 months;  the fish at 0.012 and 0.017




mg/liter NH3 weighed 9 and 38 percent less than the fish at 0.006 mg/liter.




Microscopic examination of tissues from fish exposed to the highest




concentration, examined at 6, 9, and 12 months, showed severe pathologic




changes in gill and liver tissues.  Gills showed extensive proliferation of




epithelium which resulted in severe fusion of gill lamellae which prevented




normal respiration.  Livers showed reduced glycogen storage and scattered




areas of dead cells;  these were more extensive as exposure titae increased.




     Ministry of Technology,  U.K. (1968) reported on tests in which rainbow




trout wore exposed for three  months to concentrations of 0.069, 0.14, and




0.23 mg/liter NH3•  The cumulative mortality of a control group (0.005
                                     46

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 tog/liter NTH^)  was  -2  percent.   Cumulative .nortallty ซt 0.069 and O.lt




 mg/liter NH-j was ~5 percent,  and  that  at  0.28  mg/llter was -15 percent.




 Reichenbach-Klinke (1967)  performed  a  series of  one-week ammonia tests on 240




 fishes  of nine  species  (including  rainbow trout,  goldfish, northern pike




 (Esox lucius),  carp,  and tench) at concentrations  of  0.1 to 0.4 mg/liter




 NH3•  He observed  swelling of  and  diminishing  of  the  number of red blood




 cells,  inflammations, and  hyperplasia.  Irreversible  blood damage occurred  in




 rainbow trout  fry  in  ammonia  concentrations above  0.27 mg/liter NH3 -   He




 also noted  that low NH^ concentrations  inhibited  the  growth of young  trout




 and lessened their resistance  to disease.




     Smart  (1976)  exposed  rainbow  trout to 0.30 to 0.36 mg/liter NH3  (Table




 5); 81  percent mortality occurred over  the 36-day  duration of  the test,  with




 most deaths occurring between  days 14 and 21.  Microscopic examination of the




 gills of  exposed rainbow trout revealed some thickening of the lamellar




 epithelium and an  increased mucous production.  The most characteristic




 feature was a large proportion of swollen, rounded  secondary lamellae;  in




 these the pillar system was broken down and the epithelium enclosed a




 disorganized mass  of pillar cells and erythrocytes.   Gill  hyperplasia was not




 a characteristic observation.




     Fronrn (1970)  exposed  rainbow trout to <0.0005 and  0.005 mg/liter NH3




 for eight weeks.   Subsequent examination of the gill  lamellae  of  fish from




 the trace concentrations showed them to be long and slender with  no




 significant pathology.  Fish exposed to 0.005 mg/liter  NH-j had  shorter and




 thicker  gill lamellae  with bulbous ends; some consolidation of  lamellae was




noticed.  Photomicrographs revealed that many filaments  showed  limited




hyperplasia accompanied by the appearance of cells containing  large vacuoles




whose contents  stained positive for protein.   Other lamellae showed a




                                     47

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definite hyperpLasi-i ot the epithelial layer, evidenced by  an  Increase  Li  Uie '




number of cell nuclei.




     Thurston et al. (1978) studied the toxicity of ammonia to cutthroat




trout fry in flow-through tests which lasted up to 36 days  (Table  5).




Results of duplicate tests on 1.0-g fish both showed 29- and 36-day  LC50




values of 0.56 mg/liter NT!-} •  Duplicate tests on 3.3-g fish provided 29-day




LC50 values of 0.37 and 0.34 mg/liter, slightly less than those of the  1.0-g




fish.  Tissues from heart, gastrointestinal tract, and thymus  of cutthroat




trout fry exposed to 0.34 mg/liter NH^ for 29 days were comparable to those




of control fish.  However, gills and kidneys of exposed fish showed




degenerative changes.  Gills showed hypertrophy of epithelium, some  necrosis




of epithelial cells, and separation of epithelium due to edema; kidneys




showed mild hydropic degeneration and accumulation of hyaline  droplets  in




renal tubule epithelium; reduced vacuolation was observed in livers-




     Samylin (1969) studied the effects of ammonium carbonate  on the early




stages of development of Atlantic salmon.  The first set oฃ  experiments




(temperature = 13 C) was conducted within the range 0.001 to >6.6 mg/liter




NH-j beginning with the "formed embryo" stage; the experiment lasted  53




days.  Accelerated hatching was observed with increasing (NH^^^G^




concentrations, but concentrations X).16 rag/liter NH^ were  lethal  in 12-36




hours to emerging larvae.  Because (NH^^CO-j was used as the toxicant,




the pll in the test aquaria Increased from 6.7 to 7.6 with increasing NH|j




concentration.  Growth inhibition was observed at 0.07 mg/liter NH-j  (Table




?).  Tissue disorders wore observed in eyes, brains, fins,  and blood of




MlanM^* salmon embryos and larvae exposed to concentrations from 0.16  to




>6.6 mg/liter NH^, with increased degree of symptom at increased ammonia




concentrations.  Effects observed included erosion of membranes of the  eyes




                                     48

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and shedding of  the crystalline  lens, dilation of blood  vessels  in  liver  and




brain, accumulation of blood  in  the occipital region and  in intestines.




Reaction  to light and mechanical  stimulation gradually disappeared  with




increased ammonia concentration,  and the pulseheat slowed.  Morphological




differences in development between experimental and control larvae  were




observed  from the tenth day of exposure, including a lag  in yolk resorption,




decrease  in growth of the skin fold, and contraction of  skin pigment cells




causing the skin color to become  paler than it was after  hatching.  At




concentrations up to 0.07 rag/liter NHj no significant morphological




differences were observed.




     A second series of experiments (temperature = 16.5 C) was carried out




in the 0.001 to 0.32 rag/liter NHj concentration range, and began with




larval salmon (Samylin 1969).  Concentrations of 0.21 rag/liter NH3  and




higher were lethal and caused weight loss in fry;  0.001 to 0.09 rag/liter




NH^ caused a decrease in weight gain, although no differences in feeding




activity, behavior, or development were observed in these concentrations




compared  to controls.  Dissolved  oxygen concentrations in this second series




of experiments dropped as low as  3.5 mg/liter.




     Burrows (1964) tested fingerling chinook salmon for  six weeks  in outdoor




raceways  into which ammonium hydroxide was Introduced.  Two experiments were




conducted, one at 6.1 C and the other at 13.9 C,  both at  pH 7.8.  In both




cases fish were subsequently maintained in fresh water for an additional




three weeks.  A recalculation of Burrows'  reported un-ionized ammonia




concentrations,  based on more recent aqueous ammonia equilibrium tables,




indicates that the concentrations at 6.1 C were 0.003 to 0.006 mg/liter




NH3>  and at 13.9 C were 0.005 to 0.011 mg/liter 1^3.  At both




temperatures some fish at all ammonia concentrations showed excessive




proliferation and clubbing of the gill filaments;  the degree of proliferation




                                     49

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was  progressive  for  the  first  four weeks,  after which  no  measurable  increase




was  discernible.  Examination  of a sample  of  the  fish  tested  at  6.1  C  after




three weeks  in fresh water  indicated no recovery  had taken  place from  the




extensive proliferation.  In the experiment with  larger fish  at  13.9 C a




marked recovery  from hyperplasia was noted after  the three-week  fresh  water




exposure period.  In the first experiment  the proliferated  areas had




consolidated; in the second they had not.  Burrows postulated  that continuous




ammonia exposure is a precursor of bacterial gill disease.




     Buckley et  al. (1979)  exposed duplicate groups (90 fish  each) of




hatchery-reared  coho salmon for 91 days to "river-water"  solutions of  NH^Cl




at concentrations of 0.019  to 0.33 mg/liter Nl^;  these were compared with




control groups reared at 0.002 mg/liter NI^-  Hemoglobin  content and




hematocrit readings were reduced slightly, but significantly,  at the highest




concentration tested, and there was also a greater percentage  of immature




erythrocytes at  the highest concentration.  Blood ammonia and  urea




concentrations were not significantly different after 91 days, regardless of




concentration of ammonia to which the fish were exposed.  Rankin (1979)




conducted ammonia tests with embryos of sockeye salmon (Oncorhynchus nerka)




from fertilization to hatching.  Total embryo mortality occurred at




concentrations of 0.49 to 4.9 mg/liter NH-j; times to 50 percent  mortality




at these concentrations were 40 to 26 days.  Mortality of the  embryos  exposed




to 0.12 mg/liter NTl^ was 30 percent, and time to 50 percent mortality  was




66 days.




     Two full life-cycle ammonia toxicity  tests (291 and 337 days) were




conducted with fathead minnows (Thurston et al.,  in prep.,  b).   These  tests




began with newly hatched fry and were continued through their  growth,

-------
maturation  and  spawning  stages; progeny were  exposed  from  hatching  through




growth  to 60 days  of  age.  No  statistically significant  difference  was




observed based  on  spawning data (number of egg  lots,  egg lot  size,  egg  lots




per  female, eggs per  female  per day) for  any  of  the concentrations  tested, up




to 0.96 mg/liter NH3•  However, there was a significant  effect  of ammonia




on hatching success (Table 2), with the number of fry hatching  decreasing  as




NH-j  concentration  increased  to 0.187 mg/liter NH^ and higher; no effect




on hatching success was  observed at concentrations of 0.088 mg/liter  and




lower.  Also, there was  some indication that  length of time for incubation




from spawning to hatching increased with  increasing NH3  Concentrations.  No




effect on fish growth was observed for either parental fish or  progeny after




60 days' exposure  and at exposure termination.  Significant mortalities




occurred among  the parental generation at concentrations of 0.9 to  1.0




mg/liter NH3 after 30 and 60 days' exposure,  and there were significant




mortalities among  progeny at concentrations of 0.2 to 0.4 mg/liter NH-j




after 30 and 60 days' post-hatch exposure.




     Tissues from  fathead minnows subjected to prolonged (up  to 337 days)




ammonia exposure were examined, including brain, heart, gills, kidneys,




liver, and gastrointestinal tract (Smith 1981; Thurston  et al., in prep., b).




Growths, some massive, were observed on heads of several fish




exposed to concentrations of 0.425 and 0.955 mg/liter NHg, and  swollen




darkened areas were observed on heads of several fish held at 0.216 and 0.228




mg/liter.   Grossly and histologically the severity of the lesions, which




varied from mild to severe,  was positively correlated with ammonia




concentration.   Lesions appeared to be a connective tissue type that
                                     51

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originated from the raeninx primativa covering the brain.  At the higher




ammonia concentrations proliferated tissue often completely surrounded the




brain but was not observed around the spinal cord.  Pathological changes were




not observed in other tissues.




     An early-Hfe-stage test initiated at the blastula stage of embryo-




genesis and extending through 39 days post-hatching was conducted with green




sunfish by McCormick et al. (in prep.).  Retardation of growth of green




sunfish exposed from embryo through juvenile life stages was found at NHg




concentrations of 0.489 mg/liter and higher, but not at 0.219 rag/liter and




less (Table 2).  In a long-term test on green sunfish, Jude (1973) reported




that for treatments greater than 0.17 mg/liter NHg, mean fish weight




increased less rapidly than controls after introduction of toxicant over the




next four days.  Thereafter, fish exposed to 0.26 and 0.35 mg/liter NHj




grew at an increasing rate while fish exposed to 0.68 and 0.64 mg/liter NH^




remained the same for 12 days before greater increases in growth occurred.




     An early-life-stage test with bluegill from embryo through 30 days




post-hatch was conducted on ammonia by Smith and Roush (in prep.).




Significant retardation of growth due to ammonia exposure was observed at




0.136 mg/liter NH^;  the no-observed-effect concentration was reported to be




0.063 ing/liter NH3 (Table 2).




     Broderius et al. (In prep.) conducted four simultaneous early-life-stage




ammonia tests with smallmouth bass.  These were carried out at four different




pH levels, ranging from 6.6 to 8.7, to examine the effect of pH on partial




chronic ammonia toxicity.  Exposure to ammonium chloride solutions began with




two- to three-day-old embryos and lasted for 32 days.  The effect endpoint




observed was growth,  and ammonia was found to have a greater effect on growth




at lower pH levels than at high.  NH3 concentrations Cound to retard growth

-------
 (table  2)  ranged  from 0.05S8  mg/liter  at  pH  6.60  to  0.865  mg/Liter  at  pH




 8.68.




     Early-life-stage tests (29-31 days'  exposure) on  ammonium  chloride  with




 channel catfish and white  sucker were  conducted by Reinbold  and Pescitelli




 (1982a).   No  significant effect  on percent hatch  or  larval survival was




 observed for  channel  catfish  at concentrations as high as  0.583 rag/liter




 NH3 and for white  sucker as high as 0.239 mg/liter NH3•  Significant




 retardation of growth, however, occurred  for channel catfish at




 concentrations of  0.392 mg/liter NHg and  higher and  for white sucker at




 0.070 mg/liter NH3 and higher (Table 2).  A delay in time  to swim-up stage




 was also observed  for both species at  elevated (0.06 to 0.07 mg/liter Nl^)




 ammonia concentrations.




     Robinette (1976) cultured channel catfish fingerlings for  periods of




 approximately one month at concentrations of 0.01 to 0.16  mg/liter  NH3ซ




 Growth at 0.01 and 0.07 mg/liter NH3 was  not significantly different from




 that of control fish;  growth  retardation at 0.15 and 0.16  mg/liter  NH3 was




 statistically significant.   Colt (1978) and Colt and Tchobanglous (1978)




 reported retardation  of growth of juvenile channel catfish during a 31-day




 period of exposure to concentrations ranging from 0.058 to 1.2 mg/liter




NH3ซ  Growth rate was reduced by 50 percent at 0.63 mg/liter NH3, and no




 growth occurred at 1.2 mg/liter NH3ป  The authors hypothesized  that  growth




may be inhibited by high concentrations of NH^"*" and low concentrations of




Na+ in solution, and/or the NH^/Na4" ratio.




     Ammonia exposure for 30 to 40 days of goldfish and tench resulted in




lesions .and diffuse necrosis of the caudal fin, causing it to degenerate




progressively to the  point  of  breaking off by degrees,  ultimately leaving




only a necrotized stump (Marchetti I960).




                                     53

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     Ver>  li'-cle work has  .n'eii ..uvu"  to  tiwost i ซ!:••; ป•  >vr ; >•ป i •  >•ป  .'. t ', v.- • .• .1.




factors on chronic ammonia to
-------
 (1)  At  pH  7.7  and  near  10  C,  the  threshold  concentration for chronic

     toxicity of ammonia  to rainbow  trout  has  been estimated  to be 0.031

     mg/liter NH-j •  Given the  importance of  this  species,  the FCV must

     be.  at  or below this  number.   The FCV  at pH 7.7 and  10 C  is

     therefore  set  to 0.031  mg/liter NH3•

 (2)  Both the acute data  and the chronic data  of  Broderius et al. (in

     prep.) indicate little  or no  increase of  effect concentrations with

     pH  above 7.7.  The FCV  is therefore set to 0.031 mg/liter for all

     pH  >7.7 at 10 C.

 (3)  At  pH  <7.7, the pooled  salmonid chronic toxicity data (Table 2)

     show trends consistent  with smallmouth bass  chronic  toxicity data

     (Broderius et al., in  prep.)  and the acute data.  Thus,  the  pink

     salmon datum at pH 6.4  is not only important  in its own  right, but

     also likely lies close  to where other trout  and salmon would be  at

     pH  6.4.  The FCV will  therefore be set to 0.0021 (the geometric

     average of the pink  salmon data) at pH 6.4 and  4 C.   Between pH  6.4

     and pH 7.7, the log(FCV) will be assumed  to  vary linearly with  pH,

     a trend which causes it to pass close to  other  salmonid  data and

     which  is similar to  the trend in the acute toxicity pH relationship.

(4)  Temperature effects  will be assumed to be the  same as for acute

     toxicity.   Based on  this assumption, the  pink  salmon  data  point was

     corrected  upward,  to 0.0034 at 10 C, before  computing the  pH

     relationship between pH 6.4 and 7.7.

(5)  Therefore,  for temperatures _>1G C and pH>7.7,

                      FCV10  = 0.031

     For temperatures MO C  and pH <7.7,
                                  0.031
                              100.74(7.7-pH)

                                55

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          For  temperatures <10 C,
                                       (1 +  109.73-PH)
                        FCVT - FCVin x 	A. _ ~x
                                       i 1 -^  1 OP^T P" i


     The criterion based on the FCV will be  considered as  a  30-day  average.

Provisions to  limit excursions above this criterion for  time intervals

shorter than 30 days are believed to be unnecessary, given currently

available data.

(c) Acute/Chronic Ratios

     Although  the ratios of acute effect concentrations  to chronic  effect

concentrations were not used in the derivation of FCV above,  they will  be

used in site-specific applications and therefore merit some  consideration

here.  Acute/chronic ratios are available for ten species  (Table 2).  Because

these ratios vary so widely (2.8-44), their  dependence on  species and

physico-chemical factors should be evaluated so that they  are  properly

applied.

     A clear trend exists in the species-dependence of the ratios.  Of  the

ten species, the five most acutely sensitive have ratios >11 and the five

least acutely  sensitive have ratios <6.  This has considerable  significance

because ratios are applied to FAVs which reflect species with  high  acute

sensitivities.  To produce an appropriate FCV therefore  requires using  a

ratio appropriate to such species.  The question thus arises as to  the  range

of acute sensitivity from which ratios can be justifiably  used.  The five

highest ratios here cover the LIMSMAV range  <1.8.  This range  includes  a

large number of diverse families and the FAV is unlikely ever  to exceed this

value;  thus there is little reason ever to consider the five  lowest ratios.

Within the LIMSMAV range <1.8, there is no clear trend of  ratio with
                                     56

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 sensitivity;  thus  there  is  also  no  good  reason to  try to limit further this




 set  of  ratios based  on proximity to the  FAV.




     The  sraallmouth  bass data  in Table 2 indicate  that acute/chronic ratios




 may  increase with  decreasing pH.  This is consistent  with the comment earlier




 that the  effect of pH on chronic  toxicity in  the 6.5-7.5 pH range is greater




 than the  effect of pH on acute toxicity.   The large  ratio for pink salmon




 also suggests such a pH dependence  of the ratio, if  it is assumed all




 salmonids have similar chronic and  acute  sensitivities.   The paucity of  data




 makes firm conclusions impossible,  but it is  probably inappropriate  to apply




 the  pink  salmon ratio (measured  at  pH 6.4)  to the  pH  range (>7.5) at which




 other ratios were  measured.




     Eliminating the pink salmon  ratio from the five  highest ratios  leaves




 four ratios for diverse fish species (white sucker,  fathead  minnow,  bluegill,




 rainbow trout) that were measured at similar  pHs (7.7-8.3).   These four




 ratios have a geometric average of  16 and  have no  apparent  trends with pH,




 temperature, or species sensitivity.  This average is  therefore  as




 appropriate as can be derived  from  the available data  for  application to  FAVs




 derived in the pH >7.5 range (this  range  being based  on  the  range in which




 the ratios were measured (7.7-8.3), with  extrapolation to  lower  pH being




modest because of  the indications of effects  on ratios of  pHs  <7.5,  and




extrapolation to higher pH being liberal because of the  lack  of  significant




effects of pH in this range).  No recommendation is made  here  about




appropriate ratios for lower pHs, except  that they should  probably be  higher




than 16 and will require further testing.




     Use of a constant ratio, of course,   implicitly makes  the  assumption  that




the pH dependency of  chronic toxicity is  the  same,  on a  relative  scale, as




for acute toxicity.  Although this assumption was  rejected earlier,  the




                                     57

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rejection was based on a broad pH range  that extends below  pH  7.5.   For  a




limited pH range and/or one  restricted to pHs above 7.5,  this  assumption




should cause only minor errors as long as the ratios are, as advised  above,




derived from data within the range  in question.




     An acute/chronic ratio  should  also be based on acute data consistent




with the FAV it is applied to.  Specifically, if sensitivity varies  with




organism size, the ratio should be  modified based on this variation  so that




it reflects acute data from  the same size range as the acute data on which




the FAV Is based, or the original ratio should be applied to the value the




FAV would have had if a sensitive size range had not been used.







Saltwater Animals




     Little information is available on long-term effects of sublethal




ammonia exposures on saltwater species, and no chronic data are available for




any saltwater fish species.




     Three-week exposure (Wickins 1976) of P_. setiferus to NH^Cl yielded an




EC50 value (Table 5), based  on growth reduction, of 0.72 mg/liter ^3.   A




six-week test (Table 5) with M. rosenbergii resulted in reduction in  growth




to 60-70 percent that of controls for prawn exposed to concentrations above




0.12 mg/Hter NH3 •  A "maximum acceptable level" was estimated to be  0.12




mg/liter NH3•  Armstrong et  al. (1978) conducted growth tests  (Table  5)  on




NH^Cl using prawn larvae (M. rosenbergii).  Retardation in  growth was




observed at sublethal concentrations (0.11 mg/liter NH-j at  oH  6.83 and 0.63




mg/liter NH-j at pH 7.60), and this  effect was greater at  low pH.
                                     58

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                              TOXICITY TO PLANTS


Bacteria  and  Freshwater  Plants

     Ammonia  is  known  to play an  important  part in the nitrogen metabolism of

aquatic plants.   In  the  aquatic environment,  nitrogen  plays  an important role

in  determining the composition of phytoplankton and vascular plant communi-

ties and  in some  cases can act as a  limiting  nutrient  in  primary production.

Ammonia can also  be  toxic at  certain concentrations.   Data concerning the

toxicity  of ammonia  to freshwater vascular  plants  and  phytoplankton are

contained in  Table 4.  Few of the papers  examined  contained  sufficient

information to enable calculation of  un-ionized  ammonia concentrations,

altough total ammonia solutions were  more toxic  at  high than at  low pH,  indi-

cating that toxicity was  likely due  primarily to NHj rather  than NH^ .

Some information  on ammonia effects  on bacteria  is  also included here.

     The bacterial species Escherichia coll and  Bacillus  subtills were found

to be sensitive  to NH^Cl  (Deal et  al. 1975);  1100 mg/liter NH3 killed 90

percent of an ]5.  coli population  in  78 minutes.  _B. sub tills,  an aerobic,

spore-forming bacterium, was destroyed in less  than two hours  in 620 mg/liter

NH-j •  NH-} inhibition of  the bacteria  Nltrosomonas  (that convert  ammonium

to nitrite) and the bacteria Nltrobacter  (that convert nitrite to nitrate)

was studied by Anthonisen et al.   (1976) and Neufeld et al. (1980).   NH3

inhibited the nitrification process at a concentration of 10 mg/liter

(Neufeld et al. 1980).  The NH-j concentrations that inhibited  nitrosomonads

(10 to 150 rag/liter)  were greater  than those  that  inhibited  nitrobacters  (0.1

to 1.0 mg/liter), and NH^, not NH^"1", was reported  to be the  inhibiting

species (Anthonisen et al. 1976).  Acclimation of  the nitrifiers  to  NH3>

temperature,  and  the  number of active nitrifying organisms are factors that

may aftect the inhibitors' concentrations of NH-j in a notification  system.
                                      59

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     Langowska and Moskal  (1974) investigated  the  inhibitory  effects  of




24-hour exposures to NH-j on pure cultures of ammonifying  and  denitrifying




bacteria.  Effects examined were based on ability  of  the  bacteria  to  produce




some specific metabolic processes, such as proteolysis, aramonification,




denitrification, and nitrification.  Ammonifying and  denitrifying  bacteria




were most resistant to MR-}', proteolytic and nitrifying bacteria were  the




most sensitive.  Concentrations ranging from 0.8 to 170 mg/liter NH-j  did




not adversely affect denitrifying and ammonifying bacteria; 220 mg/liter




caused reduction of the examined metabolic processes.  Proteolytic bacteria




were unaffected at 0.8 nig/liter NH^, but were  reduced to  zero  at 4.2




mg/liter;  nitrifying bacteria were unaffected  at 2.6  to 5.1 rag/liter  and




reduced to zero at 13 to 25 mg/liter.




     Experimental data concerning the toxicity of  ammonia to  freshwater




phytoplankton are limited.  Przytocka-Jusiak (1976) reported ammonia  effects




(Table 4) on growth of Chlorella vulgaris with 50 percent inhibition  in five




days at 2.4 mg/liter NH-j, and complete growth  inhibition  in five days at




5.5 rag/liter.  The MR-} Concentration resulting in 50 percent survival of C_.




vulgaris after five days was found to be 9.8 mg/liter NH^•  In a separate




study, Przytocka-Jusiak et al. (1977) were able to isolate a C. vulgaris




strain with enhanced tolerance to elevated ammonia concentrations, by




prolonged incubation of the alga in ammonium carbonate solutions.  C_.




vulgaris was reported to grow well in solutions containing 4.4 mg/liter




NH-}, but growth was inhibited at 7.4 mg/liter  (Matusiak 1976).  Tolerance




to elevated concentrations of NH-j Seemed to show a slight increase when




other forms of nitrogen were available to the alga than when ammonia was the




only form of nitrogen In the medium.  The effects of ammonia on growth of the




algal species Ochromonas sqciabii is was studied by Brettliauer  (1978).  He




                                     60

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 fpund  that  concentrations  (assuming  pH 6.5  and  30  C)  of  0.6  mg/liter NH3




 killed  the  organisms, and  at  0.3 mg/liter development  of the population  was




 reduced.  Concentrations of 0.06 to  0.15 mg/liter  NHg  had insignificant




 effect  on growth, and 0.015 to  0.03  mg/liter  enhanced  growth.




     Effects of ammonia on four algal  species (Table 4)  were studied by




 Abeliovich  and Azov  (1976).   Ammonia at concentrations over  2.5 mg/liter




 NH3 inhibited photosynthesis  and growth of  the  algal  species Scenedesmus




 obliquus and inhibited photosynthesis  of the  algae Chlorella pyrenoidosa,




 Anacystis nidulans,  and Plectonema boryanum.  Hosier  (1978)  reported that




 NH^ concentrations causing 50 percent  reduction  in oxygen production by  the




 green alga  Chlorella ellipsoidea and blue-green  alga Anabaena  sub cy1indrica




 were 16.0 x 10~8 and 251.0 x  10~8 Vig NH3-N/cell, respectively.




     The rate of photosynthesis in the blue-green  alga _P. boryanum was




 observed to be stimulated  by  NH^+, but inhibited by NH3  (Solomonson




 1969);  the  magnitude of these effects was dependent on the sodium-potassium




 composition of the suspending media.  NH3 inhibition of  photosynthesis was




 associated  with a conversion of inorganic polyphosphate  stored in the cells




 to orthophosphate.




     Champ  et al. (1973) treated a central Texas pond  with ammonia to a  mean




 concentration of 25.6 mg/liter Nl^.  A diverse population of dinoflagel-




 lates, diatoms, desmids, and blue-green algae were present before ammonia




 treatment.  Twenty-four hours after  treatment the mean number of




 phytoplankton cells/liter  was reduced by 84 percent.   By  the end of  two weeks




 (NH-j = 3.6 mg/liter) the original concentration of cells  had been reduced




by 95 percent.




     Much of the work concerning the response of freshwater  vegetation to




high ammonia concentrations is only descriptive or is a result of research




exploring the possible use of ammonia as an aquatic herbicide.



                                     61

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     Champ et al. (1973) reported virtually complete eradication of rooted

aquatic vegetation (water shield, Brasenia schreberi, and American lotus,

Nelumbo sp.) in a central Texas pond within two weeks after treatment with

anhydrous ammonia; NHg concentration was 25.6 rag/liter 24 hours after

ammonia addition, and 3.6 rag/liter two weeks later.  In experiments with

Potamogeton lucens, Litav and Lehrer (1978) observed that ammonia, which

forms a readily available nitrogen source for the plant, can be toxic when

present at high concentrations, with ammonia causing appreciable injury to

detached branches.  Ammonia inhibition of growth of Eurasian watermilfoil

(Myriophyllum spicatum) affected length and weight similarly and affected

roots and shoots similarly (Stanley 1974).

     Litav and Agami (1976) studied changes in vegetation in two rivers

subject to increased pollution from agricultural fertilizers, urban sewage,

and industrial wastes, and attributed the changes in plant species

composition primarily to ammonia and detergents.  Agami et al. (1976)

transplanted seven species of "clean water" macrophytes to various sections

of river, and found that ammonia affected only Nymphaea caerulea.   Use of

high concentrations of ammonia to eradicate aquatic vegetation was described

by Ramachandran (1960), Ramachandran et al. (1975), and Ramachandran and

Ramaprnbhu (1976).

Saltwater Plants

     Data concerning the toxicity of ammonia to saltwater phytoplankton are

presented in Table 4.  Ten species of estuarine benthic diatoms were cultured

for ten days in synthetic media at a range of NH3 Concentrations from 0.024

to 1.2 mg/liter NH3 (Admlraal 1977).  A concentration of 0.24 mg/liter

NH-j retarded the growth of most of the tested species (Table 4) .  Relative

tolerance to ammonium sulfate of five species of chrysomunads was  studied by
                                     62

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 Pinter and Provasoli (1963).   Coccolithus huxleyi was most sensitive, and




 Pavlova gyrans and  Hymenomonas sp.  were most tolerant,  with intermediate




 tolerance  exhibited by  Syracosphaera  sp. and Ochrosphaera neapolitana.




      Shilo and Shilo (1953, 1955)  reported  that  the  euryhaline algae




 Prynmesium parvum was effectively  controlled with applications of  ammonium




 sulfate, which exerted  a  lytic effect.   Laboratory and  field  tests showed




 that  the concentration  of  ammonium  sulfate  necessary for  cell lysis decreased




 with  increasing pH,  indicating that un-ionized ammonia  and not the ammonium




 ion is  responsible  for  the lytic activity of ammonium sulfate on P^. parvum.




 Effect  of  ammonia on the dinoflagellate  Amphidinium  carterae  was studied by




 Byerrura and Benson  (1975), who reported  that added ammonium ion at




 concentrations  found to stimulate  the photosynthetic rate also caused the




 algae  to release up  to 60  percent of  fixed    CC>2  to  the medium.




     Natarajan  (1970) found that the concentrations  of  fertilizer  plant




 effluent toxic  to natural  phytoplankton  (predominantly  diatoms)  in Cook




 Inlet,  Alaska,  were  between 0.1 percent  (1.1 mg/liter NH^)  and 1.0 percent




 (11 mg/liter NH-j).   At 0.1 percent effluent  concentration * C uptake




 was reduced only 10  percent, whereas at  1.0  percent  effluent  concentration a




 24-33 percent  reduction in the relative   C  uptake was  observed.   Effects




 of ammonium sulfate  on growth  and photosynthesis  of  three diatom and two




 dinoflagellate  species were reported by  Thomas et  al. (1980),  who  concluded




 that increased  ammonium concentrations found  near  southern  California sewage




outfalls would  not be inhibiting to phytoplankton  in the  vicinity.   Provasoli




and McLaughlin  (1963) reported  that ammonium  sulfate  was  toxic  to  some marine




dinoflagellates only at concentrations far exceeding  those  in seawater.




     No data were found concerning the toxicity of ammonia  to  saltwater




vegetal ion.




                                     63

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                               BIOACCUMULATION




     No data are available concerning the accumulation of ammonia by aquatic




organisms.
                                     64

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                                  OTHER DATA




      A  number  of  investigators  have  studied effects  of  ammonia  on behavior




 and  various  metabolic  processes of exposed  animals,  or  have conducted field




 studies.   This research  has  dealt predominantly  with freshwater fishes.






 Freshwater Invertebrates




      The  effect of  ammonia (Table 5)  on the ciliary  beating rate of  clam




 gills was  investigated by Anderson et  al. (1978).  Concentrations of 0.036 to




 0.11 ing/liter  NHj caused a reduction  in ciliary  beating rate of fingernail




 clams;  the effect of these concentrations ranged from 50 percent reduction in




 beating rate to complete inhibition of  cilia.  Adult clams  (>5  mm) were  more




 sensitive  than juveniles (<5 mm); adults were also slightly more




 sensitive  than the  unionid mussel (Elliptic complanata)  and the Asiatic  clam




 (C. manilensis).  Shaw (1960) investigated  effects of ammonium  chloride  on




 sodium  influx  in the freshwater crayfish, Astacus pallipes.   Ammonia produced




 an inhibition  of sodium influx; a concentration  of 18 rag/liter  NH^+




 reduced the influx  to about 20  percent  of its normal value,  and  influx




 reduction  was  related to greater ammonia concentration.   This effect was




 attributed to  NH^   ions and not to any  toxic effect  exerted  on  the




 transporting cells by un-ionized ammonia.   NH^"1"  did  not  affect  chloride




 influx nor the  rate of sodium loss.




     Ammonia was added to a Kansas stream at a 24-hour  average  concentration




 of 1.4 mg/liter NH3, and a 24-hour drift net sampling was conducted




 (Liechti and Muggins 1980).   No change  in diel drift pattern was  observed,




but there was  an increase in magnitude  of drift, a shift  in  kinds of




 organisns present, and changes in benthic standing crop  estimates; the




 ammonia concentration was concluded to be nonlethal.

-------
Fre shwa te r Fi she s




     Herbert and Shurben (1963) Investigated the effect on susceptibility to




ammonium chloride solutions of rainbow trout forced to swim continuously




against water currents of different velocities prior to ammonia exposure.




Forcing rainbow trout to swim for one to two days at 85 percent of the




maximum velocity they could sustain increased their susceptibility only




slightly, corresponding to a 20 to 30 percent reduction in the 24- or 48-hour




LC50.




     The behavioral response of blacknose dace (Rhinichthys atratulus) to




ammonium chloride solutions has been studied (Tsai and Fava 1975;  Fava and




Tsai 1976);  the test fish did not avoid concentrations of 0.56 or 4.9




mg/liter NH^, nor did these concentrations cause significant changes in




activity.  Avoidance studies were conducted by Westlake and Lubinski (1976)




with bluegill using ammonium chloride solutions.  Bluegill detected




concentrations of approximately 0.01 to 0.1 mg/liter NH3ป and evidenced a




decrease in general locomotor activity.  No apparent avoidance of ammonia was




observed, and there was some indication of an attraction.  Behavioral




responses of bluegill to a five-hour exposure to 0.040 mg/liter XH-j>




although variable,  were related to at least a small amount of physiological




stress either at the gill or olfactory surfaces.  At a concentration of 0.004




mg/liter NH^, bluegill evidenced slight temporary increases in both




activity and turning behavior;  no preference or avoidance was demonstrated,




with responses seemingly exploratory (Lubinski et al.  1978, 1980).  Wells




(1915) investigated the avoidance behavior of bluegill to ammonium hydroxide




solutions and reported that fishes did not avoid ammonia prior to being




killed by it.   In a study of the repelling ability of  chemicals to green




sunfish,  Summerfelt and Lewis (1967) concluded that concentrations of ammcaia




                                     66

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'high enough to repel fish would ho rapidly tcitป .-ปv>>tJtn, ,-i <-ซ.ป>..-i f.-u->u i




 with threespine stickleback, solutions of ammonia concentration O.J/ mg/liter




 NH-j  elicited a positive (attraction)  response from the test fish (Jones




 1948).




      Weltering et  al.  (1978),  in tests with largemouth bass and mosquitofish,




 demonstrated that  predator-prey interactions  were sensitive to sublethal




 concentrations of  NHj-   Ammonia concentrations of 0.63 and 0.86 mg/liter




 NH3  decreased  prey consumption and bass growth;  bass  were reported to be




 more sensitive than mosquitofish to NH3ซ   The effect  of ammonium chloride




 on consumption of  juvenile  chinook salmon by  brook trout  was studied by




 Hedtke  and  Norris  (1980).   At  the lowest  test concentration of 0.29 mg/liter




 NH3>  trout  consumption  rates decreased as much as 65  percent.   As  ammonia




 concentration  increased, however,  consumption of prey increased  and was




 double  that  of controls at  the highest tested concentration of 0.76 mg/liter




 NH-j-  Increased  consumption rate was  related  to  both  increased NH-j




 concentration  and  increased  prey density.   The magnitude  of  the  effect of




 ammonia was  not  the same at  all  prey  densities,  having a  greater effect on




 consumption  rate at high than  at low  prey densities.   Mortalities  were




 observed among prey salmon  at  the  highest  NH3  levels,  and these  were




 attributed  to  the  combined  effect  of  NH^  and  stress from  presence  of the




 predator.   Brook trout  exhibited toxic  effects due to  NH-j •




     NH/^Cl and NH^HcOg  Solutions were  injected intraarterially into




 rainbow trout  (Hillaby  and  Randall  1979).   The same dose  of  each compound was




 required to  kill fish,  but  there was  a  more rapid  excretion  of NH3  after




 NH4HC03 infusions,  resulting in  higher  NH3  Concentrations in blood,




 than after NH^Cl infusions.  Ammonium  acetate  solutions of different




 concern rations were  injected intraperltoneally into three  species  of fishes




                                     67

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(Wilson 1968; Wilson et al. 1969).  LD50 values (mmoles/kg body weight)  for




channel catfish for one to four hours was 26.7 to 18.7, for goldfish  for  one




hour were 29.3 and 29.6 in two separate tests, and for rainbow trout  for  one




hour was 17.7.  Goldfish was the most resistant species tested and rainbow




trout the least resistant.  Nehring (1964) compared toxicity of ammonia  in




the water to toxicity of ammonia administered orally and concluded that  the




threshold and lethal concentration values were considerably lower for  ammonia




in water than for ammonia administered orally.




     Acute symptoms of NH-j toxicity to brown trout sac fry and 12-day-old




fry were described by Penaz (1965), who exposed fry to concentrations  ranging




from 0.08 to 50.0 mg/liter NH^-  Symptoms caused by Ntt-^ exposures were:




rapid spasm-like movements at concentrations of 2.0 mg/liter NH-j and  higher




within 16-17 minutes of exposure;  after 40 minutes these symptoms were also




observed at 0.4 mg/liter NH-j >  After 2.5 hours these abnormal movements




ceased, and at 10 hours heart activity was decreased and fish lost movement




ability at the higher (>2.0 mg/liter N^) concentrations.  Other symptoms




included Inability to react to mechanical stimulation and disorders in rhythm




of mouth movements culminating in the mouth's staying rigidly open.  Thumann




(1950), working with rainbow trout and brook (=brown?) trout, described




observed symptoms of ammonia poisoning to fishes to be convulsions and




frequent equilibrium and positional anomalies.




     Smart (1978) reported that exposure of rainbow trout lo an acutely




lethal concentration of 0.73 mg/liter NH^ resulted in an increase in oxygen




consumption, increase in ventilation volume, decrease in percent oxygen




utilization, increase in respiratory frequency and amplitude (buccal




pressure), decrease in dorsal aortic blood PQ? ป increase in dorsal aortic




blood pressure, and increase in mean heart rate.  Physiological parameters




                                     68

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 not  significantly  affected  by  NH_-j  exposure were "cough" race,  dorsal aort li-




 blood  pH, blood  P50,  erythrocyte count,  heraatocrit,  and hemoglobin




 concentration.   Coho  salmon exposed  to  concentrations  ranging  from 0.094 to




 0.162  rag/liter Nll-$ (Souea and  Meade  1977)  exhibited  hyperexcitability,




 hyperventilation,  ataxia, and  progressive  acidemia;  metheraoglobin




 concentrations in  blood  of  exposed fish did not differ  significantly from




 those  of controls.  Effects  on trout  (species not  specified) blood with




 exposure to accumulated  excreted NHg  were  investigated  by  Phillips et al.




 (1949) and were  reported to  include  an  increase in blood carbon  dioxide




 content and a decrease in oxygen content.




     Arillo et al.  (1979d) measured  gill sialic acid content in  rainbow  trout




 exposed to NH^Ofl or NH^Cl solutions  ranging from 0.05 to 0.5 mg/liter




 NH-j, and reported  that increasing  NH^ concentrations produced  increasing




 gill sialic acid content.  Elevated gill sialic  acid levels were  also




 produced by higher  ammonium  ion (NH^  )  concentrations at identical NH^




 concentrations, and the authors concluded  that  NH^  was  a  stressor




 causing elevated sialic acid levels.   Exposure  of  rainbow  trout  (14-cm




 length) for four hours to NH^Cl and NH^On  solutions of  concentrations




 ranging from 0.094  to 0.50 mg/ltter NH^ resulted in increased  proteolytic




 activity and free amino acid levels in  the  fish  livers, but no statistically




 significant change in fructose 1,6-biphosphatase enzyme activity  (Arillo  et




 al. 1978,  1979a).  Renal renin activity was reported (Arillo et al.  1931b)  to




 increase in rainbow trout exposed  to concentrations of 0.048 to 0.61  rag/liter




 NH-j •   A significant decrease in liver glycogen and increase in free  glucose




were  observed in rainbow trout exposed to NH^Cl  solutions  for  four  hours  at




 a concentration of 0.048 mg/liter NH3, and a decrease in total carbo-




hydrates was observed at 0.12 mg/liter NH3 (Arillo et al.  1979b).   For




                                     69

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trout similarly treated with .umnonlum hydroxide,  .-slgivif ti-aui.  Uei-redsiea  ii>
glycogen and carbohydrates, and increase in glucose occurred  at 0.097
mg/liter N!^.
     A statistically significant increase in rainbow  trout  liver
concentrations of cyclic-31,5'-adenosine-raonophosphate  (cAMP) was  reported by
Arillo et al. (1979c) to be induced by a four-hour exposure to elevated
ammonia concentrations of 0.011 to 0.124 mg/liter NH3ซ  Decreases  in liver
glycogen levels were also measured and were significantly different from
controls only in the trout  exposed to 0.048 mg/liter  NHj, the highest
exposure used for glycogen measurements.  The authors concluded that cAMP
measurements provided a very sensitive means of discerning  fish stress even
at very low toxicant concentrations, although quantitative measurement of
stress intensity was not possible.  Lysosoraal lability  was  also investigated
as an indicator of stress in rainbow trout due to ammonia exposure (Arillo et
al. 1980) , and was reported to increase significantly for fish subjected to
concentrations of 0.048 to 0.61 mg/liter NHj.  Exposure of  rainbow trout
for four to 48 hours to 0.024 to 0.61 mg/liter NH3 resulted in changes in
various brain and liver metabolites; the magnitude of the changes was depen-
dent on both exposure time  and NH-j concentration  (Arillo et al- 1981a) .
     Exposure of walking catfish (Clarias batrachus)  to ammonia caused
inhibition of fish brain cholinesterase and kidney peroxidase activity
(Mukherjee and Bhattacharya 1974, 1975a).  Plasma corticosteroid
concentrations were measured (Tomasso et al. 1981) in channel catfish exposed
to 1.1 mg/liter NH^ for 24 hours; corticosteroid  levels increased
initially, peaked after eight hours, then decreased.  The overall  increase
was approximately tenfold over normal levels.
     Korting (I969b) reported that carp exposed to 1 mg/liter NH-j exhibited
an increase in number of blood erythrocytes, reaching an initial maximum
                                     70

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 after  several hours  followed by a gradual  decrease;  after  50  hours  the  number




 was  less  than the  average  for  non-exposed  fish.  Other blood  changes  from the




 ammonia exposure were: thickening of individual erythrocytes,  reduction of




 osmotic resistance of erythrocytes, increase  in concentrations  of urea  and




 lactic acid, and decrease  in ATP concentration.  Levi et al.  (1974) reported




 that goldfish exposed for  24 hours  to NH^Cl solutions exhibited  increases




 in cerebral and blood concentrations of glutamine and in other amino  acids,




 with changes most  pronounced in the brain.  Concentrations of  free  amino




 acids  in  livers showed only slight increases  of a few amino acids,  including




 glutamine, and the concentration of lysine decreased.  No change in




 concentrations of  free amino acids was observed in kidneys.   Rainbow  trout




 exposed to 0.33 mg/liter NH^ had significantly higher packed  cell volumes;




 exposures to concentrations of 0.24 rag/liter  NH^ and higher resulted  in




 significantly raised blood glucose and plasma cortisol concentrations (Swift




 1981).




     Diuretic response of  rainbow trout exposed to concentrations of 0.09 to




 0.45 mg/liter NH3 was studied by Lloyd and Orr (1969).  After an initial




 lag period, urine production increased rapidly during exposure then returned




 to normal within a few hours after discontinuation of NH3 exposure.  A




 no-obs^rved-effect concentration was reported to be 0.046 mg/liter NHo -




 Goldfish were exposed to solutions containing 1.0 to 1.9 mg/liter NH^




 (Fromm 1970;  Olson and Fromm 1971);  onset of death was characterized by  a




 gradual cessation of swimming movements and settling to the bottora of the




 tank.  Some goldfish near death were returned to ammonia-free water in which




 they recovered to at least some degree.  In similar experiments (Fromm 1970;




Olson and Fromm 1971) rainbow trout  were exposed to ambient total ammonia




concentrations of 0.04 to 0.2 mg/litei  NH3•  There wad a decrease in total




                                     71

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nitrogen excreted with increase in ambient NH-p and a concomitant decrease




in the. NH^ portion of total nitrogen excreted; urea and protein nitrogen




excretion rates showed no changes as ambient NH-j increased.  Onset of death




for trout was characterized by violent thrashing movements.




     Exposure of rainbow trout to solutions of NH^Ci for 24 hours (Froram




and Gillette 1968;  Fromm 1970) showed that an increase in ambient water NH3




concentration resulted in a corresponding increase in blood NH-j




concentrations, and a decrease in total nitrogen and NH-j excretion.  The




decrease in NH^ excretion accounted for half or less of the total nitrogen




excretion, depending on the water NH^ concentration, indicating that the




reduction in NH3 excretion was to some extent compensated for by increased




excretion of some other nitrogenous compound(s).




     Young fry (2-20 days old) of loach (Misgurunus anguilicaudatus) and carp




were exposed for five to 70 hours to   N-labeled ammonium chloride




solutions at six concentrations from 0.002 to 0.064 mg/liter NH-j (ito




1976), and the proportion of ^N relative to total N in the fishes




determined.  Ammonia was shown to be directly absorbed by the fry; nitrogen




conversion rate increased with increasing ammonia concentration and exposure




time.  Nitrogen conversion rates for carp fry decreased as fry age increased




from 3 to 20 days.   After 48 hours of exposure to 0.064 mg/liter NH-j




followed by transfer to ammonia-free water, rapid excretion (15-20 percent)




of the absorbed   N occurred during the first hour in ammonia-free water.




Excretion rate then slowed, with about 50 percent of the absorbed   N




being retained after 48 hours in ammonia-free water.  Comparison of   N




absorption rates between live and sacrificed three-day-old carp fry showed




one-third to one-half the uptake of   N by dead fry compared with live,
                                     12

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 indicating  that  the  uptake  of  ammonia from water  by  live  fish  occurs  not  only




 by  simple membrane permeation  but  also by  metabolic  action.




     Flagg  and Hinck (1978)  reported  that  exposure to NH3  lowered  the




 resistance  of channel catfish  to the  pathogen Aeromonas hydrophila.   In 17-




 and 28-day  tests, increasing exposure concentrations from  0.02  to  0.04




 rag/liter NH^ resulted in  increasing numbers of bacteria in host livers.




 Schreckenbach et al. (1975)  reported  that  ammonia in pond  water leads to




 outbreaks of gill necrosis  in  carp, accompanied by an increase  in  ammonia




 concentration in serum of the  fish.   This  is aggravated at elevated pH  levels




 due to increasing inhibition of ammonia excretion at increasing pH levels,




 with ammonia excretion being almost totally blocked  at pH values above  10.5.




 After investigating  the possible role  of parasites,  bacteria, viruses,  and




 other ultramicroscopic agents  in causing gill necrosis, the authors concluded




 that pH-dependent intoxication or  autointoxication with ammonia was the sole




 cause of the gill damage.   Studies of  the  treatment  and prophylaxis of gill




 necrosis using 28 different  therapeutical  preparations led to the  conclusion




 that only those preparations that  lowered  the water  pH level and/or ammonia




 concentrations resulted in an  improvement  in clinical symptoms.




     Increase in frequency of opercular rhythm in fishes was monitored as a




 means to measure fish response to  sublethal concentrations of ammonia (Morgan




 1976, 1977).  Ammonia threshold detection  concentration (Table  5)  for




 largemouth bass was approximately 30 percent of the  lethal concentration




 (LC50) for that species.  Increases in largemouth bass opercular rhythms  and




activity were electronically monitored (Morgan 1978,  1979) to determine




 threshold effect ammonia concentrations (Table 5); for a 24-hour exposure the




effect concentration for opercular rhythms was 0.028  mg/liter NH-j  and for




activity was 0.0055 mg/liter.  Lubinski et al. (1974) observed  that ammonia




stress apparently caused bluegill to consume more oxygen.




                                     73

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     In field experiments in an Arizona mountain lake, mortalities of oa




rainbow trout were attributed to high un-ionized ammonia concentrations and




high pH levels; 20 to 100 percent of test fish died in 24 hours at NH^




concentrations of 0.109 to 0.225 mg/liter (Fisher and Ziebell 1980).  Ammonia




added to a Kansas stream at 24-hour average concentrations of 1.4 mg/liter




NH3 resulted in fry of slender madtora (Notorus exllis), Notropis sp., and




orangethroat darter being collected in large numbers in a 24-hour drift net




sampling;  these fishes are not normally found in drift net samples, and their




presence was attributed to toxic effects of the ammonia (Liechti and Huggins




1980).







Saltwater Invertebrates




     Sublethal toxicity of NH^Cl to the quahog clam and eastern oyster was




studied by Epifanio and Srna (1975) who measured the effect of ammonia over




20 hours on the rate of removal of algae (Isochrysis galbana) from suspension




(clearing rate) by the clams and oysters.  Concentrations of 0.06 to 0.2




mg/liter MH3 affected clearing;  no difference was observed between




juveniles and adults.  The effect of ammonia on the ciliary beating rate of




the mussel Mytilus edulis was studied by Anderson et al. (1978).




Concentrations of 0.097 to 0.12 mg/liter NH3 resulted in a reduction in




ciliary beating rate from 50 percent to complete inhibition (Table 5).




     Exposure of unfertilized sea urchin (Lytechinus pictus) eggs to NH^Cl




resulted in stimulation of the initial rate of protein synthesis, an event




that normally follows fertilization (Winkler and Grainger 1978).  NH^l




exposure of unfertilized eggs of St rongylocent rot us purpuratus, L^. pictus,




and Strongylocentrotus drobachiensis was reported (Paul et al. 1976; Johnson




et al. 1976) to cause release of "fertilization acid", more rapidly and in




greater amounts than after insemination.  Activation of unfertilized L_.




pictus eggs by NH/jCl exposure was also evidenced by an increase in




                                     74

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 Intracellular  pH  (Shen ami  Stelnhardt  1978;  Steinhardt  and  Mazia  1973).




 Ammonia  treatment was also  reported  to activate  phosphorylation of  thymidirce




 and  synthesis  of  histones in unfertilized eggs of  the sea urchin  S^.




 purpuratus (Nishioka 1976).  Premature chromosome  condensation was  Induced by




 ammonia  treatment  of eggs of L. pictus and S. purpuratus (Epel et al.  1974;




 Wilt and Mazia 1974; Krystal and Poccia 1979).   Ammonia treatment of S^.




 purpuratus and S. drobachiensis fertilized eggs  resulted in absence of  the




 normal uptake of calcium following insemination, but did not inhibit calcium




 uptake if ammonia  treatment preceded insemination  (Paul and Johnston 1978).




     The polychetous annelid (Nereis succinea),  the channeled whelk (Busycon




 canalieulaturn), and the brackish water clam  (Rangia cuneata) were subjected




 to ammonia concentrations of 0.85, 0.37, and 2.7 mg/liter NH3 and ammonia




 excretion measured (Mangura et al. 1978).  The excretion of  ammonia in  these




 species was inhibited by non-lethal concentrations of ammonia;  the authors




 concluded that ammonia crosses the excretory epithelium in  the ionized form,




 and  that the process is linked to the activity of  the Na+ + K+ ATPases.




When blue crab (Callinectes sapidus) were moved  from water  of 28 ppt salinity




 to water of 5 ppt, a doubling of ammonia excretion rate occurred;  addition of




 excess NH^Ci to the low salinity water inhibited ammonia excretion and




 decreased net acid output (Mangura et al. 1976).  The effect of gaseous NH3




on hemoglobin from blood of the common marine bloodworm (Glyeera dibrachiata)




was  examined (Sousa et al. 1977) in an attempt to determine whether there was




competition between NH^ and oxygen in binding to hemoglobin; such an




NH-j/02 relationship was not found.







Saltwater Fishes




     Mo other data were found for saltwater fish species.

-------
                                 UNUSED DATA




     Many references cited In the References section were not used in the




text or tables, for a variety of reasons.  Listed below are unused




references, with a brief explanation for their being relegated to this




category.  For those several cases where more than one reason applies to a




given paper, it is listed only under the principal reason for its not being




used.




     The following references were not used because the research they




reported was conducted using aquatic organisms not resident in North America:




Alderson (1979), Arizzi and Nicotra (1980), Brown and Currie (1973), Brownell




(1980), Chin (1976), Currie et al. (1974) Dockal and Varecha (1967), D'Silva




and Verlencar (1976), Giussani et al. (1976), Greenwood and Brown (1974),




Grygierek et al. (1978), Inamura (1951), Nicotra and Arizzi (1980),




Orzechowski (1974), Reddy and Menon (1979), Sadler (1981), Saha et al.




(1956), Shaffi (1980b), Singh et al.  (1967), Stroganov and Pozhitkov (1941),




Thomas et al.  (1976), Turoboyski (1960), Vailati (1979), Woker (1949),




Wuhrmann (1952), Wuhrmann and Woker (1953), Wuhrmann and Woker (1955),




Wuhrmann and Woker (1958), Yamagata and Niwa (1982).




     The following references were not used because insufficient water




chemical composition data were provided to permit calculation of NH-j:




Belding (1927), Binstock and Lecar (1969), Chu (1943),  Danielewski (1979),




Das (1980), Ellis (1937), Hepher (1959), Joy and Sathyanesan (1977), Kawamoto




(1961), Mukherjee and Bhattacharya (1978), Oshima (1931), Oya et al. (1939),




Patrick et al. (1968), Rao and Ragothaman (1978), Roberts (1975), Rushton




(1921), Scidmore (1957), Shelford (1917), Shevtsova et  al. (1979), Sigel et




al. (1972), Southgate (1950), Wolf (1957a), Wolf (1957b), Zgurovskaya and




Kustenko (1968).




                                     76

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      The  following  references  were  not  used  because  they  report or)




 published  elsewhere which  was  cited  in  this  Jiviuarut  ! row  t hซป  .ปthoi




 publicatlon(s):  Burkhalter  (1975),  Colt  (1974), Dept. of  Environment,  U.K.




 (1972), Herbert  (1955),  Hillaby  (1978), Larmoyeux and  Piper  (1973), Ministry




 of Technology, U.K. (I960),  Ministry of Technology,  U.K. (1966), Rice  (1971),




 Smart  (1975), Wilson  (1974).




     The  following references  were not used  because  they were  foreign-




 language papers  for which  no translation was available, and  no useful




 information could be  obtained  from the abstract:  Desavelle  and Hubault




 (1951), Fedorov  and Smirnova (1978), Frahm (1975), Garcia-Romeu and Motais




 (1966), Guerra and Comodo  (1972), Guseva  (1937), Hubault (1955), Jocque and




 Persoone  (1970), Kawamoto  (1958), Korting (1976), Krauss (1937), Kuhn and




 Koecke (1956), Leclerc and Devlaminck (1950), Maraontova (1962), Oya et al.




 (1939), Pequignot and Moga (1975), Pora and Precup (1971), Revina (1964),




 Saeki  (1965), Schaperclaus (1952), Scheuring and Leopoldseder  (1934),




 Schreckenbach and Spangenberg  (1978), Steinmann and  Surbeck  (1922a),




 Steinmann and Surbeck (1922b), Svobodova (1970), Svobodova and Groch (1971),




 Teulon and Simeon (1966), Truelle (1956), Varaos and  Tasnadi  (1962a), Vamos




 and Tasnadi (1962b), Vamos et  al. (1974), Yasunaga (1976), Yoshihara and Abe




 (1955).




     The following references  were not used because  they relate more to




 ammonia metabolism in fishes,  than to ammonia toxicity:  Bartberger and




 Pierce (1976), Becker and Schmale (1978), Brett and  7,ala (1975), Cowey and




 Sargent (1979),  Creach et al.  (1969), Cvancara (1969a), Cvancara (1969b), De




and Bhattacharya (1976), De Vooys (1968), De Vooys (1969), Driedzic and




Hochachka (1978), Fauconneau and Luquet (1979), Fechter (1973), Fellows and




Hird (1979a),  Fellows and Hird (1979b), Flis (1968a), Flis (1963b), Florkin




and Duchateau (1943),  Forster and Goldstein (1966),   Forster and Goldstein



                                     77

-------
(I960), Fromm (1963), Cirard and Payan (1980), Goldstein and Forster (1961),


Goldstein and Forster (1965), Goldstein et al.  (1964), Gordon (1970),


Gregory (1977), Grollman (1929), Guerin-Ancey (1976a),  Guerin-Ancey (1976b),


Guerin-Ancey (1976c), Guerin-Ancey (1976d), Hays et al.  (1977), Hoar (1958),


Muggins et al. (1969), Janicki and Lingis (1970),  Katz  (1979), Kaushik and


Luquet (1977), Kloppick et al. (1967), Kutty (1978), Lawrence et al. (1957),

                                                   /ป
Lum and Hatnraen (1964), Maetz (1973),  Maetz and Garcia-Roraeu (1964),


Makarewicz and Zydowo (1962), Mason (1979a), Mason (1979b), Matter (1966),


McBean et al. (1966), McKhann and Tower (1961), Moore et al. (1963), Morii et


al. (1978), Morii (1979), Morii et al. (1979), Mukherjee and Bhattacharya


(1977), Nelson et al.  (1977), Payan  (1978), Payan and  Maetz (1973), Payan


and Matty (1975), Payan and Pic (1977), Pequin and Serfaty (1963), Pequin and


Serfaty (1966), Pequin and Serfaty (1968), Pequin  et al. (I969a),  Pequin et


al. (1969b), Raguse-Degener et al. (1980), Ray and Medda (1976), Read (1971),


Rice and Stokes (1974), Rychley and Marina (1977), Savitz (1969),  Savitz


(1971), Savitz (1973), Savitz et al.   (1977), Schooler  et al. (1966), Smith


(1929), Smith (1946), Smith and Thorpe (1976), Smith and Thorpe (1977),


Storozhuk (1970), Sukumaran and Kutty (1977), Tandon and Chandra (1977),


Thornburn and Matty (1963), Vellas and Serfaty (1974),  Walton and Cowey


(1977), Watts and Watts (1974), Webb  and Brown (1976),  Wood (1958), Wood and


Caldwell (1978).


     The following references were not used because the material the authors


used was a complex compound or had an anion that might  in itself be toxic:


Curtis et al. (1979), Johnson and Sanders (1977),  Simonin and Pierron


(1937), Vallejo-Freire et al. (1954).


     The following references were not used because they dealt with complex


effluents or waste waters,  of which ammonia was a  primary component:  Brown


                                     78

-------
•'et  al.  (1970),  Calamari  and Marchetti (I1)?1*).  <7upM et al. (t, Twan ami




 Cella  (1979), Janicke  and  Liidemann (1967),  Lloyd and Jordan (196J), Lloyd and




 Jordan  (1964),  Martens and Servizi (1976),  Matthews and Myers (1976), Mihnea




 (1978),  Nedwell (1973),  Okaichi  and  Nishio  (1976),  Perna (1971),  Rosenberg et




 al.  (1967),  Ruffier  et al. (1981),  Sahai  and  Singh  (1977), Shaffi (1980a),




 Vamos  (1962), Vamos  and  Tasnadi  (1972), Ward  et  al. (1982).




     Three  references consisted  only of an  abstract,  providing  insufficient




 information  to  warrant their  use:  Liebmann and  Reichenbach-Klinke (1969),




 Mukherjee and Bhattacharya (1975b),  and Redner et al.  (1980).
                                     79

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                                   SUMMARY




     All concentrations used herein are expressed as un-ionized ammonia




(Nil-}), because NH^, not the ammonium ion (NH^ ) has been demonstrated




to be the principal toxic form of ammonia.  The data used in deriving the




criteria are predominantly from flow-through tests in which ammonia




concentrations were measured.  Ammonia was reported to be acutely toxic to




freshwater organisms at concentrations (uncorrected for pH) ranging from 0.53




to 8.00 rag/liter NH-j for 12 Invertebrate species representing nine families




and from 0.083 to 4.60 mg/liter NH3 for 23 fish species from nine families.




Among fish species, reported 96-hour LC50 values ranged from 0.083 to 1.09




mg/liter for salmonids and from 0.14 to 4.60 mg/liter NH-j for non-salmonids.




Reported data from chronic or partial chronic tests on ammonia with two




freshwater invertebrate species, both daphnids, showed effects at concentra-




tions (uncorrected for pH) ranging from 0.304 to 1.2 mg/liter Nt^i and with




nine freshwater fish species, from five families, ranging from 0.0017 to




0.612 mg/liter NH3-




     Concentrations of ammonia acutely toxic to fishes may 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 kidneys.




     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 affect




the toxicity of un-ionized ammonia itself, either ameliorating or exacer-




bating the effects of ammonia.  Factors that have been shown to affect




                                     80

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 ammonia toxicity include dissolved oxygen concentration,  temperature, pH,




 previous  acclimation  to  ammonia,  fluctuating or intermittent exposures,




 carbon  dioxide  concentration,  salinity,  and the presence  of other toxicants.




      The  most well-studied  of  these is  pH;  the  acute toxicity of NH-j has




 been  shown  to increase as pH decreases.   Sufficient  data  exist from toxicity




 tests conducted  at  different pH values  to formulate  a mathematical expression




 to  describe  pH-dependent acute NH^ toxicity. The  very limited amount of




 data  regarding  effects of pH on chronic  NH^ toxicity also indicate




 increasing NH-j  toxicity  with decreasing  pH, but the  data  are insufficient




 to  derive a  broadly applicable toxicity/pH relationship.   Data on temperature




 effects on acute  NH^  toxicity are  limited,  and  somewhat variable, but




 indications  are  that  NH3 toxicity  is  greater at low  (<10  C)  temperatures.




 There is no  information  available  regarding temperature effects  on chronic




 NH3 toxicity.




      Examination  of pH-corrected acute NH-j  toxicity  values  among families




 of  freshwater organisms  tested showed that  the  most  sensitive  families  are




 Percidae and Salraonidae, with walleye and  rainbow  trout being  the most




 sensitive tested  species in  each of these  families;  invertebrates are




 generally more tolerant  than fishes.  Available  chronic toxicity data for




 freshwater organisms  show that the most  sensitive  families among  those  tested




 are Salmonidae and  Catostomldae (suckers),  with  pink salmon  and  white sucker




being the most sensitive tested species  in  these families.   Limited  data  for




 invertebrates, mostly cladocerans  and one  insect species,  indicate  they are




generally more tolerant  than fishes; however, the  fingernail clam appears  to




be as sensitive as  salmonids.  The range of  acute/chronic  ratios  for  ten




species from six  families was 2.8  to 44, and acute/chronic ratios were  higher




for the species  found to be the most sensitive on  -in  acute (pH-corrected)




                                     81

-------
basis.  Available data indicate  that differences in sensitivities between




warm and cold water families of  aquatic organisms are inadequate to warrant




discrimination in the national ammonia criterion between bodies of water  with




"warm" and "cold" water fishes;  rather, effects of organism 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 NH^ than are invertebrates or fishes.  The




ammonia criterion appropriate for the protection of aquatic animals will




therefore in all likelihood be sufficiently protective of plant life.




     Available acute and chronic data for ammonia with saltwater organisms




are very limited, and insufficient to derive a saltwater criterion.  A few




saltwater invertebrate species have been tested, and the prawn Macrobrachium




rosenbergii was the most sensitive.  Acute toxicity of NH^ appears to be




greater at low pH values, similar to findings in freshwater.  Data for




saltwater plant species are limited to diatoms, which appear to be more




sensitive than the saltwater invertebrates for which data are available.




     Although a great deal of information has been published about ammonia




toxicity to aquatic life, much of it provides little,  if any, quantitative




data.   There are some key research needs that need to be addressed in order




to provide a more complete assessment of ammonia toxicity.  These are:




(1) acute tests with saltwater fish species, and additional saltwater




invertebrate species;  (2) life-cycle and early-life-stage tests with




representative freshwater and saltwater organisms from different families,




with investigation of pH effects on chronic toxicity;  (3) fluctuating or




intermittent exposure tests for a variety of species anJ exposure patterns;




                                     82

-------
(4) tests at cold water temperatures, both acute and chronic; (5) more




hlstopathological and histochemical research with fishes, which would provide




a rapid means of identifying and quantifying sublethal ammonia effects; (6)




studies on effects of dissolved and suspended solids on acute and chronic




toxicity.
                                    83

-------
                              NATIONAL CRITERIA

     To protect freshwater aquatic life,  the criteria  for ammonia  (in

rag/liter un-ionized NH-j) are:

     (1) the concentration at all times should be  less than  or  equal  to  the

         numerical value given by 0.15 x  f(T)/f(pH), where

             f(pH) - 1 + 10l-03(7.32-pH)

             f(T)  - I, T _> 10 C

                   * 1 4- 109-73-pH
                           pKT-pH
                                  ' T < 10 C
     (2) The average concentration over any 30 consecutive days  should be  less

         than or equal to 0.031 x f(T)/f(pH), where

             f(pH) = 1, pH _> 7.7

                   ป 100.74(7.7-pH)> pH < 7.7

             f(T)  ซ 1, T >_ 10 C
                     ! + iQ9.73-pH
                   * 1 + iOPKT~pH • T < 10 C

Criteria values for the pH range 6.5 to 9.0 and the temperature  range 0 C  to

30 C are provided in the following tables.  Total ammonia concentrations

equivalent to each NH-j criterion are also provided in these tables.

     Data available for saltwater species are insufficient to derive a

criterion for saltwater.
                                     84

-------
                   (1) Maximum allowed concentrations for ammonia.*
pH
0 C
5 C
10 C
15 C
20 C
25 C
30 C
Un-ionized Ammonia (rag/liter NHj)
6.50
6.75
7.00
7.25
7.50
7.75
8.00
8.25
8.50
8.75
9.00

6.50
6.75
7.00
7.25
7.50
7.75
8.00
8.25
8.50
8.75
9.00
0.008
0.014
0.021
0.030
0.040
0.049
0.056
0.061
0.065
0.068
0.071

31.9
29.5
25.7
20.8
15.5
10.6
6.84
4.22
2.54
1.53
0.94
0.013
0.021
0.032
0.046
0.061
0.074
0.084
0.091
0.096
0.100
0.104

31.9
29.5
25.7
20.8
15.5
10.6
6.84
4.22
2.54
1.53
0.94
0.019
0.031
0.048
0.069
0.091
0.110
0.125
0.135
0.141
0.145
0.147
Total Ammonia
31.9
29.5
25.7
20.8
15.5
10.6
6.84
4.22
2.54
1.53
0.94
0.019
0.031
0.048
0.069
0.091
0.110
0.125
0.135
0.141
0.145
0.147
(mg/liter
21.8
20.1
17.6
14.2
10.6
7.29
4.71
2.92
1.78
1.09
0.69
0.019
0.031
0.048
0.069
0.091
0.110
0.125
0.135
0.141
0.145
0.147
NH3)
15.0
13.9
12.1
9.84
7.34
5.06
3.28
2.05
1.27
0.80
0.52
0.019
0.031
0.048
0.069
0.091
0.110
0.125
0.135
0.141
0.145
0.147

10.5
9.69
8.48
6.89
5.15
3.56
2.33
1.47
0.93
0.60
0.41
0.019
0.031
0.048
0.069
0.091
0.110
0.125
0.135
0.141
0.145
0.147

7.41
6.86
6.00
4.88
3.66
2.55
1.68
1.08
0.70
0.47
0.33
* To convert these values to mg/liter N,  multiply by 0.822.
                                          85

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               (2) 30-day average allowed concentrations for ammonia.*
pH
0 C
5 C
10 C
15 C
Un- ionized Ammonia (mg/liter
6.50
6.75
7.00
7.25
7.50
7.75
8.00
8.25
8.50
8.75
9.00

6.50
6.75
7.00
7.25
7.50
7.75
8.00
8.25
8.50
S.75
9.00
0.0018
0.0027
0.0042
0.0064
0.0098
0.0138
0.0139
0.0140
0.0142
0.0145
0.0150

6.82
5.87
5.06
4.36
3.77
2.99
1.70
0.97
0.56
0.33
0.20
0.0027
0.0041
0.0063
0.0096
0.0148
0.0208
0.0209
0.0210
0.0211
0.0214
0.0219
Total
6.82
5.87
5.06
4.36
3.77
2.99
1.70
0.97
0.56
0.33
0.20
0.0040
0.0061
0.0094
0.0144
0.0220
0.0310
0.0310
0.0310
0.0310
0.0310
0.0310
Ammonia
6.83
5.89
5.07
4.37
3.78
3.00
1.70
0.97
0.56
0.33
0.20
0.0040
0.0061
0.0094
0.0144
0.0220
0.0310
0.0310
0.0310
0.0310
0.0310
0.0310
(mg/liter NH3
4.65
4.01
3.45
2.98
2.58
2.05
1.17
0.67
0.39
0.23
0.14
20 C
NH3)
0.0040
0.0061
0.0094
0.0144
0.0220
0.0310
0.0310
0.0310
0.0310
0.0310
0.0310
)
3.21
2.76
2.38
2.06
1.78
1.42
0.81
0.47
0.28
0.17
0.11
25 C

0.0040
0.0061
0.0094
0.0144
0.0220
0.0310
0.0310
0.0310
0.0310
0.0310
0.0310

2.24
1.93
1.67
1.44
1.25
1.00
0.58
0.34
0.20
0.13
0.09
30 C

0.0040
0.0061
0.0094
0.0144
0.0220
0.0310
0.0310
0.0310
0.0310
0.0310
0.0310

1.58
1.37
1.18
1.02
0.89
0.72
0.42
0.25
0.15
0.10
0.07
* To convert these values to rag/liter N,  multiply by 0.822.
                                          86

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                      KXAMP1.KS  OF  SITR-SrKOU'lO  CR




      National  criteria  are  subject  to modification,  if  appropriate,  to




 reflect  local  conditions.   One method provided  in  the Site-Specific  Criteria




 Guidelines  (U.S. Environmental Protection Agency 1982b)  for  such modification




 is  to base  certain  calculations only on  those species that occur in  the body




 of  water of  interest.   As an example of  how  site-specific criteria for




 ammonia may  differ  from the national criteria,  calculations  were performed




 for several  sites.




     The sites were chosen  on  the basis  of readily available information on




 the presence of fish and invertebrate species and on a  reasonable diversity




 between sites.  The sites were:




     (1) St. Louis  Bay  - mouth of the St. Louis River on Lake  Superior at




         Duluth, Minnesota/Superior, Wisconsin  (A. R. Carlson, pers. comra.




         1982)




     (2) Straight River at Owatonna, Minnesota  (A. R. Carlson, pers. comm.




         1982)




     (3) Piceance Creek, Colorado (Goettl and Edde 1978; Gray  and Ward 1978)




     (4) Poudre River at Fort Collins, Colorado (W. T. Willingham, pers.




         comm. 1982)




     (5) Colorado River at Utah/Colorado border (W. T. Willingham, pers.




         comm. 1983)




The calculations here are limited to temperatures of 15 C and  to a single pH




typical of the site (see following table).  Therefore, the criteria here




should not be considered as final criteria for the sites because variation




with pH and temperature was not explored.




     For each site, available surveys of species occurrence were used  to




identify which of the families  tested for acute toxicity (Table 3) were

-------
present.  Minimum data set requirements for the diversity of  families were




met except where inappropriate to a site (U.S. Environmental  Protection




Agency 1982b).  The national LIMFMAV (Table 3) was used for each family,  even




if it is based in part or whole on species not occurring at the site, unless




LIMSMAVs were available for all species in the family at the  site, in which




case a site-specific LIMFMAV was computed that was based only on site




species.  The LIMFMAVs so developed for each site are listed  in the following




table.




     The Guidelines method for estimating the FAV as the fifth percentile of




FMAVs was applied to the set of LIMFMAVs selected for each site.  If the




LIMFAVs so computed exceeded the LIMSMAV of an important species at a site,




or of an important size class of an important species, the LIMFAV was lowered




to the lowest such LIMSMAV.  The FAVs at each site were then  computed by




adjusting the LIMFAVs to the site pHs, using the acute pH relationship




already discussed.  The maximum criterion at each site was set to one-half of




the site FAV.




     In all cases, the site pH was >7,5, so an acute/chronic  ratio of 16 was




used.  At sites 2, 4, and 5, this ratio was applied directly  to the FAV to




obtain an FCV.  At sites 1 and 3, where the LIMFAV was lowered to reflect an




age/size class, the ratio was applied to the value the FAV would have had if




the LIMFAV was only lowered to LIMAVs averaged over all data  for a species,




not just an age/size class.  If the resultant PCV at a site exceeded the




average chronic value (measured at pH >7.5) of an important species present




at the site, the FCV was lowered to this species' chronic value.  The 30-day




average concentration was set to the FCV.




     If rainbow trout are present at a site, the maximum allowable




concentration is the s^uie as the national value.  If raiubow  trout are




                                     88

-------
 ab,sent, but  percids  or  other  salmonids  are  present,  the maximum allowable




 concentration  Is  about  twofold higher than  the  national number.   This is  not




 due  to differences in general species sensitivities, but rather reflects




 unavailability of information on  the sensitivity of different  age/size




 classes for  species  other  than rainbow  trout; these higher  numbers  should




 therefore be treated with  caution as perhaps being relatively  underprotective




 compared to  the national criterion.  Only where both Percidae  and Salmonidae




 are absent (site  5)  is  the maximum allowable concentration  greatly  higher,




 about fourfold, than the national number.   Again, however,  a large  part of




 this difference may be  due to the dearth of toxicity information on sensitive




 age/size classes.




     The 30-day average criterion shows even less variation among sites than




 the maximum criterion,  due to the size class issue not  affecting sites




 differently.  If rainbow trout are present, the average  criterion is  near the




 national value.  If rainbow trout are absent, but percids or other  salmonids




 are present, the average criterion is only  slightly (<20%)  higher than the




 national value.  Only if both Percidae and  Salmonidae are absent from  the




 site is the number substantially higher than the national value and  the




difference is little more than twofold.
                                     89

-------
           Site-specific aramoni.* criteria examples  for  t lv
Site Number
LIMFAVs:
Elmidae
Ephemerellidae
Asellidae
Actacidae
Tubificidae
Tctaluridae
Perlodidae
Poeciliidae
Baetidae
Percichthyidae
Cyprinidae
Daphnidae
Cottidae
Dendrocoelidae
Catostomidae
Centrarchidae
Salraonidae
Percidae
pH
Temperature (C)
1

a
6.22
3.86
3.46
3.03
2.87
2.70
a
a
a
1.92
1.84
1.65
a
1.57
1.49
0.90
0.82
7.6
15
2

8.30
6.22
a
3.46
3.03
2.87
a
a
2.36
a
1.92
a
a
a
1.57
1.49
a
0.82
7.5
15
3

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6.22
3.86
a
3.03
a
2.70
a
2.36
a
1.92
a
1.65
a
1.57
a
0.90
a
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4

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a
3.86
3.46
3.03
2.87
a
a
2.36
a
1.92
1.84
a
a
a
1.30b
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7.9
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5

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6.22
3.86
3.46
3.03
2.37
2.70
a
2.36
a
1.92
1.84
a
a
1.57
1.49
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a
8.1
15
Maximum allowed concentration
(ing/liter NH3):
as un- ionized ammonia
(as total ammonia)
30-day average concentrat
(mg/liter NH3) :
as un-ionized ammonia
(as total ammonia)

0.10C
(9.3)
ion

0.028
(2.6)

0.21
(25)


0.026
(3.0)

0.13C
(3.9)


0.03ld
(0.9)

0.27
(13)


0.034
(1.6)

0.5S
(18)


0.073
(2.2)
(a) Family not present at site.




(b) LIMFAV adjusted to reflect only site species.




(c) LIMFAV lowered to L1MAV of adult rainbow trout.




(d) FCV lowered to CV of rainbow trout.






                                      90

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Larmoyeux, J.D. and R.G. Piper.  1973.  Effects of water reuse on rainbow



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-------
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(• , .   ), iiiificjis  60604

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