DRAFT
                                                                        EPA-822-D-00-001
                         2000 UPDATE OF AMBIENT WATER QUALITY  CRITERIA FOR

                                              CADMIUM
\6
      1
      '•V
          Prepared by:

Great Lakes Environmental Center
 Traverse City,  Michigan 49686
                                           Prepared for:

                               "U.S.  Environmental Protection Agency
                                          Office of Water
                                 Office  of  Science and Technology
                                          Washington,  DC
                                          U.S. EPA Headquarters Library
                                         '       Mail code 3201
                                          1200 Pennsylvania Avenue NW
                                           *  Washington DC 20460
                                   ['
                                   EPA Contract  No.  68-C-98-134
                                      Work  Assignment.No.  1-11

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                                    NOTICES
 This document  has been reviewed by the  Health and Ecological Effects Criteria
 Division,  Office of Science and Technology,  U.S.  Environmental Protection
 Agency,  and approved for publication.

 Mention of trade names or commercial products does not constitute endorsement
 or recommendation for use.

 This document  is available to the public through the National Technical
.Information Service (NTIS), 5285 Port Royal Road, Springfield, VA 22161.
                                        n

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FOREWORD
       V.t: I
    111

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                               ACKNOWLEDGMENTS

 John G. Eaton
 {freshwater author)
 Environmental Research Laboratory
 Duluth, Minnesota

 Charles B. Stephan
 (document coordinator)
 Environmental Research Laboratory
 Duluth, Minnesota
                John H. Gentile
                (saltwater author)
                Environmental Research Laboratory
                Narragansett, Rhode Island

                David J. Hansen
                (saltwater coordinator)
                Environmental Research Laboratory
                Narragansett, Rhode Island
Statistical Support:
Clerical Support:
John W. Rogers
Terry L. Highland
Document Update Effort: June, 2000
 Gregory J. Smith
 (freshwater contributor)
 Great Lakes Environmental Center
 Columbus, Ohio

 Cindy Roberts
 (document coordinator)
 US EPA
 Health and Ecological Effects
 Criteria Division
 Washington, D.C.
                                       IV

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                                    CONTENTS




                                                                             Page




Notices  	'.	•..".".".". ."."	  ii




Foreword		 iii




Acknowledgments	.'	  iv




Contents  	 	   v




Tables  	r	v •,	  vi




Figures  . . . . .	•'•'	vi




Introduction 	;....•	  1




Acute Toxicity to Aquatic Animals 	  2




Chronic  Toxicity to Aquatic  Animals	 10




Toxicity to Aquatic Plants	 17




Bioaccumulation 	•	.-...- .-.-.-•-:•'-:.= :. :	 18




Other  Data 	 19




Unused Data	 22




Summary	'	29




National Criteria  ..;	- - •	 30




References  	 138

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                                    TABLES
la.   Acute Toxicity of Cadmium to Aquatic Animals 	 38
lb.   Results of Covariance Analysis of Freshwater Acute Toxicity
      Versus Hardness 	.-...-.	 65
2a.   Chronic Toxicity of Cadmium to Aquatic Animals	 66
2b.   Results of Covariance Analysis of Freshwater Chronic Toxicity
      Versus Hardness ..	 70
2c.   Acute-Chronic Ratio 	 71
3a.  . Ranked Genus Mean Acute Values with Species Mean Acute-Chronic
      Ratios	 72
3b.   Ranked Freshwater Genus Mean Chronic Values	 83
4.     Toxicity of Cadmium to Aquatic Plants	 85
5.     Bioaccumulation of Cadmium by Aquatic Organisms  	 89
6.     Other Data on Effects of Cadmium on Aquatic Organisms  	 95

                                    FIGURES

l.  Comparison of All Table 1 Freshwater Acute Toxicity Test ECSOs and
    LCSOs with the Hardness Slope Derived CMC	33
2.  Ranked Summary of Cadmium GMAVs (Freshwater)  	 34
3.  Ranked Summary of Cadmium GMAVs (Saltwater)  .....  	 35
4.  Comparison of All Table 2 Freshwater Chronic Values with the
    .Hardness Slope Derived CCC   	 36
5.  Chronic Toxicity of Cadmium to Aquatic Animals   	 37
                                       VI

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Introduction1
•.Vtir £'.
      Cadmium is a relatively rare element-^that  is not essential  for any
biological process in plants or animals.   It occurs mainly as a component of
minerals in the earth's crust at an average.concentration of 0.18 ppm  (Babich
and Stotzky 1978).  Cadmium levels in soils usually range from approximately
0.01 to 1.8 ppm  (Lagerwerff and'Specht .1970).  in natural fresh waters,
cadmium sometimes occurs at concentrations of less than 0.01 pg/L, but  in
environments impacted by man, concentrations can be several micrograms  per
liter or greater.  Cadmium can enter the environment from various
anthropogenic sources, such as by-products from  zinc refining, coal
combustion, mine wastes, electroplating processes, iron and steel production,
fertilizers and pesticides (Button 1983) .
      The impact of cadmium on aquatic organisms depends on a variety of
possible chemical forms of cadmium (Callahan et  al. 1979), which  might  have
different toxicities and bioconcentration  factors.  In most well  oxygenated
fresh waters that are low in total organic carbon, free divalent  cadmium will
be the predominant form.  Precipitation by carbonate or hydroxide and
formation of soluble complexes by chloride, sulfate, carbonate, and hydroxide
should usually be of little importance.  In salt waters with salinities  from
about 10 to 35 g/kg, cadmium chloride complexes  predominate.  In  both fresh
and salt waters, particulate matter and dissolved organic material may bind a
substantial portion of the cadmium.
      Because of the variety of forms of cadmium (Callahan et al. 1979)  and
lack of definitive information about their relative toxicities, no available
analytical measurement is known to be ideal for  expressing aquatic life
criteria for cadmium.  Previous aquatic  life criteria for cadmium (U.S.  EPA
1980) were expressed in terms of total recoverable cadmium  (U.S.  EPA 1983a),
but this measurement is probably too rigorous in some situations.  More
      1 An understanding of  the'"Guidelines  for  Deriving Numerical National
Water Quality Criteria for  the Protection,of Aquatic Organisms and Their Uses"
 (Stephan et al.  1985), hereafter  referred to as the Guidelines, is necessary
in order to understand the  following  text;  tables, and calculations.

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recently,  U.S. EPA  (1985) expressed cadmium criteria  as  acid-soluble cadmium
(operationally defined as the cadmium that passes  through a  0.45  ^m membrane
filter after the sample  is acidifie'dr;tb pH =  1.5 to 2.0  with nitric acid).
      The criteria presented herein supersede previous aquatic life water
quality criteria for cadmium  (U.S.- EPA 1976,  1980,  1985,  1995, 1999a)  because
these new criteria were  derived based on the  most .recent literature.   Whenever
adequately justified, a  national criterion may  be  replaced by a site-specific
criterion (U.S. EPA 1994a), which may include not  only site-specific criterion
concentrations  (U.S. EPA 1994b), but also site-specific  durations of averaging
periods and site-specific frequencies of allowed exceedences  (U.S. EPA 1991).
All concentrations  are expressed as cadmium,  not as the  chemical tested.   The
latest literature search for  information for  this  document was conducted in
June, 1999; some newer information was also used.
Acute Toxicitv to Aquatic Animals
      Acceptable  data on the acute  effects  of cadmium in freshwater are
available  for  43  species of invertebrates,  27 species of fish,  one salamander
species, and one  frog species (Table  1).  Although many factors might affect
the results of tests of the toxicity  of  cadmium to aquatic organisms (Sprague
1985),  water quality criteria can quantitatively take into account only
factors for which enough data are available to show that the factor similarly
affects the results of tests with a variety of species.  Hardness is often
thought of as  having a major effect on the  toxicity of cadmium, although the
observed effect may be due to one or  more of a number of usually interrelated
 ions,  such as  hydroxide, carbonate, calcium, and magnesium.  Hardness is used
 here  as a surrogate for the ions which affect the results of toxicity tests on
 cadmium.
       Acute  tests were conducted at three different levels of water hardness
 with Daphnia magma  (Chapman et al.  Manuscript), demonstrating that daphnids
 were at least five  times more sensitive to cadmium in soft than hard water
  (Table 1) .  Data  in Table 1 also indicate that cadmium was more toxic to the
 tubificid worm Limnodrilus hoffmeisteri, Ceriodaphnia reticulata, Daphnia
 pulex, chinook salmon, goldfish., fathead minnow, green sunfish, striped bass
                                        2

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and bluegill in soft Chan  in hard water.  Other  species  tested at  different
hardness levels (e.g., rainbow  trout) ,;did.not  show the  same consistent  water
hardness to acute toxicity relationship as discussed above, possibly due  to
differences in the various test conditions.  Carroll et  al. (1979)  found  that
calcium, but not magnesium, reduced the acute" toxicity  of cadmium.
      Other water quality  characteristics could  potentially influence the
toxicity of cadmium to aquatic  species.  Giesy et  al.  (1977)  found that
dissolved organics substantially reduced the toxicity of cadmium to daphnids,
but had little effect on its toxicity to fish.   No consistent relationship
between toxicity and organic particle size was observed.  Development of  the
"biotic ligand model"  (BLM - formerly the "gill  model")  in recent  years has
attempted to better account for the bioavailability of  metals to aquatic  life.
The BLM, which quantifies  the capacity of metals to bind to the gills of
aquatic organisms, can be  used  to calculate  the  bioavailable portion of
dissolved metals in the water column  based on  site-specific water  quality
parameters such as alkalinity,  pH and dissolved  organic carbon (U.S. EPA
1999b).  Future development of  the BLM for cadmium will  help better quantify
the bioavailable fraction  of cadmium.
      A tendency for  increasing resistance to  toxicity  with increasing  size or
age has been reported (Table  1) in the snails, Ami cola sp. (Rehwoldt et  al.
1973) and Physa gyrina  (Wier and Walter  1976), the coho salmon (Chapman 1975),
a-nd the common carp  (Suresh et  al.  (1993) .   No such effect was observed with'
increasing age  (Table 1)  in the cladoceran,  Daphnia magna  (Stuhlbacher  et al.
1993),  the rainbow  trout   (Chapman  1975,  1978), or  in the striped bass (Hughes
1973;  Palawski et al.  1985).  Data are unavailable for  a sufficient number of
species and  life stages  to allow general adjustment of  test results or
criteria on  the basis of  size  or  life stage,   where relationships  were
apparent between  life-stage and sensitivity,  only values for the most
sensitive  life-stage  were considered.
       Currently,  the  primary quantitative correlation used' to modify metal
toxicity estimates  is water hardness (viz.  the USEPA 1984 water quality
 criteria  for cadmium).   Hardness (as calcium or magnesium  ions) almost
 certainly has some direct effect on cadmium £oxicity (e.g. by influencing
                                       3

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membrane integrity),  but it also serves as a general surrogate for pH,
alkalinity, and ionic strength, becaus'e waters of higher hardness usually  have
higher pH,  alkalinity, and ionic strength.
      Although past water quality criteria for cadmium  (and other metals)  have
been established upon the loosely defined term of "acid soluble metals,"' U.S.
EPA made the decision to allow the expression.of metal criteria on the  basis
                                              i
of dissolved metal (U.S. EPA 1994), operationally defined as  that metal that'
passes through a 0.45 micron filter.  Because most of the data in existing
databases are from tests that were either nominal concentrations, or provided
only total cadmium measurements, some procedure was required  to estimate their
dissolved equivalents.  U.S. EPA evaluated the data on dissolved-total
relationships from existing data, and then had a number of tests conducted
under conditions  (static, flow-through, fed, and unfed) that  typified standard
acute and chronic toxicity tests from which  criteria are derived.  These
studies were used to  derive conversion  factors  (CFs)  (Stephan 1995; Lussier et
al. 1995; Univ. of Wisconsin-Superior 1995).  For certain metals like cadmium,
these CFs are hardness dependent.
      Based upon the  results of these studies, acute freshwater total cadmium
concentrations were converted  to dissolved concentrations using the factor of
0.97 at a  total hardness level of SO mg/L as CaC03, 0.94 at a total hardness
level of 100 mg/L  as  CaC03r  and 0.92  at  a  total  hardness level of 200  mg/L as
CaC03.   Acute saltwater total cadmium values were converted to dissolved using
the  factor of 0.994.   For the  final  criterion values, conversion from total to
dissolved  was used because hardness  relationships were  established based upon
total cadmium concentrations as this minimized  the  number of  conversions
required,  and because of the uncertainty  of  the  conversion factor  in  tests
 reporting  acute  toxicity at higher cadmium concentrations.   In cases  where
only dissolved cadmium was  reported  in  freshwater,  conversion to total  used
 the  same  appropriate  factor.
      To  account for  the apparent  relationship  of  cadmium acute  toxicity  to
 hardness,  an analysis of covariance  (Dixon and  Brown 1979; .Neter and  Wasserman
 1974)  as  noted in the guidelines  (Stephan 1985}  was performed using the
 Statistical Analysis  System.(SAS line.,  Cary, NO  software program  to  calculate

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the pooled slope for hardness using the natural  logarithm of the.acute value
as the dependent variable, species as the treatment  or grouping variable, and
the natural logarithm of hardness as the covariate or independent variable.
The pooled slope is a regression slope from  a pooled data set,  where  every
variable is adjusted relative to it's mean. . The species  are adjusted
separately, then pooled for a single conventional least squares regression
analysis.  The slope of the regression line  is the best estimate of the
all-species relationship between toxicity and hardness,   with analysis of
covariance, different species will be weighted relative to the  number of data
points they have.  In this case, the fathead minnow  has 29 data points out of
the total of.69, and the next most frequent  species  has just 6  data points.
      This analysis of covariance model was  fit  to the data in  Table  1 for the
10 species for which definitive acute values are available over a range of
hardness such that the highest hardness is at least  three times the lowest,
and the highest is also at least 100 mg/L higher than the lowest (other
species in Table 1 either did not meet these criteria or  did not show any
hardness-toxicity trend due to differences in exposure methods,  species age,
etc.).  For D. magna, only acute toxicity tests  that were initiated with less
than 24-hr old neonates were used to estimate the hardness slope.  For the
striped bass, the data from Rehwoldt et al.  (1972) were not used because the
data were too divergent.  The slopes for all 10  species ranged  from 0.1720 to
1.535, and the pooled slope for these 10 species was 0.9931 (see Table Ib) .
An F-test was used to test whether a model, with  separate  species slopes for
each species gives significantly better fit  to the data than the model with
parallel slopes.  This test showed that the  separate slopes'model is  not
significantly better, and therefore the slopes are not significantly  different
than the overall pooled slope  (P=0.66).  The slopes  and confidence intervals
associated with the  10 species  indicated that D. magna (all available data)
had a very flat slope and a  large confidence interval (and large standard
error).  If  only  the D. magna data from Chapman  et al. (Manuscript) were used,
the resultant D.  magna slope was  1.1324, with smaller confidence intervals
than  for the all  D.  magna slope.   If this  reduced data set is used (all
species  but  using only data  from  Chapman et  al.  (Manuscript) for D. magna),

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 the pooled slope for these-species was 1.2049 (see Table Ib).  The test for
 equality of the 10 slopes using the reduced data set (all species but only
 Chapman D.  magna data)  produced P=0.99.  It therefore is,reasonable to assume
 that the slopes for these 10 species are the same, and that the overall slope
                                                                    '  ,'  •
 is a -reasonable- estimate of the average relationship between hardness.and
-toxicity.   Either'p value indicated that it was reasonable to assume that the
 slopes were the same,  however, the second model'was considered the better
                                             t
 model  and was therefore selected.  The pooled slope of 1.2049 is close to the
 slope  of 1.0 that is expected on the basis that cadmium, calcium, magnesium,
 and carbonate all have a charge of two.  A plot of the acute effect level
 (EC50  or LC50)  versus total hardness is provided in Figure 1.
       The potential for a back-transformation bias associated with the
 hardness slope adjustment has been noted by Newman (1991).   However,  the bias
 discussed by the author reviews bias for single species in least squares
 regression, rather than ANCOVA used here, so it is not clear how biases may
 accumulate (or cancel)  with combined,species and a combined slope.
       The pooled slope of 1.2049 was used to adjust the freshwater acute
 values in Table 1 to hardness a so mg/L,' except where it was not possible
 because no hardness was reported.  Species Mean Acute Values (SMAV) were
 calculated as geometric means of the adjusted acute values.  As stated in the
 guidelines (Stephen 1985),  flow-through measured study data are given
 preference over rion-flow-through data for a particular species.  In certain
 cases  flow-through measured results were available,  yet preference was given
 to the sensitive life stage for certain species in calculating SMAVs.  Only
 data from Chapman (1975) were used for coho salmon'and only data from Rehwoldt
 et al. (1972) were used for the common carp to avoid using test results from
 studies in which the life stage tested is known to be less sensitive, or in
 which the life stage tested is unreported and the' higher LCSOs may be due
 primarily to the use of less sensitive life stages.   The available acute
 values for U. imbecilis, striped bass and brook trout covered a wide range.
 The data for Palawski et al.  (1985) were used for striped bass because they
 were considered better data than those given in U.S.  EPA (1985),  although  the
 data from Hughes (1973) support:Jthe newer data^   Only some of the Keller

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unpublished data were used to calculate the SMAV for  U.  intbecilis.  The  data
for. brook trout were not used in the calculation of the  Final Acute Value.
Drumraond and Benoit  (Manuscript) reported that stress greatly affected the
sensitivity of brook trout to cadmium.
      The SMAV for freshwater invertebrates ranged from  12.00 ng/l> total
cadmium for the mussel, Anodonca coupierana to 78,579 ^g/L total  cadmium for
the midge, Chironomus riparius.  Of the fish  species  tested, the  brown trout,
Salmo tructa, had the lowest SMAV of 1.656 fig/L total cadmium,  and the
tilapia, Oreochromis mossambica, recorded the highest fish SMAV of 11,861 iig/L
total cadmium.  As indicated by the data, both invertebrate and fish  species
display a wide range of sensitivities to cadmium.
      Fish species represent eight of the nine most sensitive species to
cadmium (Table 3).  Salmonids  (Salmo truCta,  Oncorhynchus kisutch,
Oncorhynchus mykiss, and Oncorhynchus tsnawytscha) are four of  the five  most
sensitive species listed in Table 1, and thus are more sensitive  to cadmium
than any other freshwater animal species thus far tested (Chapman 1975,  1978,
1982; Cusimano et al. 1986; Davies et al. 1993; Finlayson and Verrue  1982;
Phipps and Hoicombe  1985; Spehar and Carlson  1984a,b).   The mussel, Anodonta
coupierana, is the sixth most  sensitive species to cadmium, and thus  the most
sensitive invertebrate species  tested thus far  (Keller Unpublished).
      Genus Mean Acute Values  (GMAV) at a hardness of SO mg/L were then
calculated  (Table 3) as geometric means of the available freshwater Species
Mean Acute Values and ranked.   Of the 59 genera for which acute values are
available, the most  sensitive  genus, Salmo,  is over 47,451 times  more
sensitive than the most resistant,  Chironomus.  The first through fourth most
sensitive genera  (and a total  n of  59 were considered) in the computation of
the final acute value.  Because there are  59  GMAVs, the  four lowest GMAVs were
selected  as being closest  to the  fifth percentile of  toxicity,  even though  the
second  through the sixth values were also  equally as  close to the fifth
percentile.   The  sensitivity of these  four most sensitive genera  are  within a
factor  of 7.2, and except  for  the fourth genus  (Anodonca), all  are fish. Of
the  ten most.sensitive genera,  seven  are  fish, one  is a  mussel, one is a
cladoceran,  and  one  is a  bryozoan (Figure  2;  Table  3).   Hardness-adjusted
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acuce values are available for more than one species  in  10 genera,  and the
range of SMAVs within each genus is less than a  factor of 4.0  for eight of the
10 genera.  The ninth genus, Ptychocheilus. has  two SMAVs that differ by a
factor of 146, possibly due to differences  in the  test conditions between
species.  The tenth genus, Morons, has SMAVs that  differ*by  a  factor of 2,954,
but only the most sensitive species was used because  the two species values
are too divergent to use for the genus value.   ;-',.
      The freshwater Final Acute Value  (FAV) for total cadmium at a hardness
of 50 mg/L was calculated to be 5.995 ng/l>  total cadmium from  the Genus Mean
Acute Values in Table 3 using the procedure described in the Guidelines.  The
Species Mean Acute Values for four salmonids and the  striped bass are lower,
but the acute value for the brown trout and striped bass are from static
tests, whereas flow-through measured tests  have  been  conducted with the
remaining .three salmonid species.  The  freshwater  Final  Acute  Value for total
cadmium at a hardness of 50 mg/L was lowered to  4.296 ^g/L to  protect the
important rainbow trout  (Table 3).  This value  is  above  the  SMAV of 1.656 M9/&
for the brown trout and 2.535 for striped bass,  but below all  other SMAVs
listed  in Table 3  (Figure 2).  The resultant freshwater  Criterion Maximum
Concentration  (CMC) at a hardness of 50 mg/L for total cadmium (in j«g/L) =
eU.205(ln(hardnesS}]-3.949)_   If the CMC baged Qn  total cadmiura values  is
converted to  dissolved cadmium using the 0.97  factor  at  a hardness of 50 mg/L.
determined by EPA  (Stephan  1995;  Lussier et al.  1995; Univ.  of Wisconsin-
Superior  199S), the freshwater CMC for  dissolved cadmium, (in M9/L)  = 0.97
 [e (1.205 (in (hardness)] -3.949), _  ThuSj the 2 . t Mg/L CMC  for digsoived  cadmium at
a hardness  of 50 mg/L  is below all of  the  SMAVs but  the  brown  trout presented
in Table  3  (Figure 2).
       Tests  of  the acute  toxicity of  cadmium  to saltwater  organisms have been
conducted with 50  species  of invertebrates  and 11  species  of fish  (Table 1}.
 The SMAVs for saltwater invertebrate  species  range from 41.29  ^ig/L for a mysid
 to 133,000 /ug/L for an oligochaete worm (Tables 1  and 3).   The acute values
 for saltwater polychaetes range  from 200  ng/L for Capitella capicata to 14,100
      for Neanches arenaceodentata (Reish and LeMay 1991),  but  the larvae of C.

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capitata are 38 times more  sensitive  than  the  adults.   Saltwater molluscs have
Species Mean Acute Values from  227.9  /ug/L  for  the  Pacific oyster to 19,170
Mg/L for the mud snail.
      Frank and Robertson  (1979)  reported  that the acute toxicity to juvenile
blue crabs was related  to salinity.   The 96-hr LCSOs were 320,  4,700,  and
11,600 M9/L at salinities of  1,  15, and 35 g/kg, respectively.   The LC50 at
the very low salinity is in Table 6 and was not used in deriving criteria.
Studies with Americsuaysia bahia (formerly  Mysidopsis bahia)  by Gentile et al.
(1982) and Nimmo et al.  (1977a)  also  support a relationship between salinity
and the acute toxicity  of cadmium.  O'Hara (1973a)  investigated the effect of
temperature and salinity on the toxicity of cadmium to the fiddler crab.  The
LCSOs at 20°C were 32,300,   46,600, and 37,000  ^g/L  at  salinities of  10, 20,
and 30 g/kg, respectively.  Increasing the temperature from 20 to 30°C lowered
the LC50 at all salinities  tested.  Toudal and Riisgard (1987)  reported that
increasing the temperature  from 13 to 21°C at  a salinity of  20  g/kg  also
lowered the LCSO value  of cadmium to  the copepod,  Acartia tonsa.
      Saltwater fish species  were generally more resistant to cadmium than
freshwater fish species with  SMAVs ranging from 75.0 uglli for the striped bass
(at a salinity of  1 g/kg) to  50,000 j/g/L for the sheepshead minnow.   In a
study of the interaction of dissolved oxygen and salinity on the acute
toxicity of cadmium to  the  mummichog, Voyer  (1975)  found that the 96-hr LCSO
at a temperature of 18-20°C and a salinity of  32 g/kg was about one-half what
it was at 10 and 20 g/kg.   Sensitivity of  the  mummichog to acute cadmium
poisoning was not  influenced  by reduction  in dissolved oxygen concentration to
4 mg/L.  This increase  in  toxicity with  increasing salinity conflicts with
other data reported  in  Tables 1 and  6.
      Of the 54 saltwater genera for  which acute values are available, the
most  sensitive, Americamysis, is 3,270 times more  sensitive than the most
resistant, Monopylephorus   (Table 3) .  Acute values are available for more than
one species  in  each  of  seven genera,  and the  range of Species Mean Acute
Values within each genus  is no more  than a factor  of 3.6 for six of the seven
genera.  The  seventh genus, Crassostrea,  has two SMAVs that differ by a factor?
of  16.7, possibly  due to different exposure conditions between species. Only

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the data from Reish et al.  (1976) were used  for  Capitella  eapitata, only  data
from Martin et al. (1981) and Nelson et al.  (1976) were used  for Mytilus
edulis, only data from Sullivan et al.  (1983) were used for Buryte/nora
affinis, only data from Cripe (1994) were used for Penaeus- duorarum,  and  only
data from Park et al. (1994) were used for Rivulus mazmorat'us to avoid  using
test results from studies in which the life  stage tested is known  to  be less
sensitive or in which the life stage tested  is unreported  and the  higher  LCSOs
may be due primarily to the use of less sensitive life stages.  The
sensitivities of the four most sensitive genera  differed by a factor  of 2.7,
which includes two myaids, the striped bass  and  the American  lobster.
      The saltwater Final Acute Value for total  cadmium calculated from the
Genus Mean Acute Values in Table 3 is 80.55  f^g/L.  This Filial Acute Value is
below the SMAV for the mysid, Mysidopsis bigelowi  (110 M9/L) ,  but  is
aproximately three percent above the American lobster  (78  pg/lt), approximately
seven percent higher than the striped bass  (75.0 jj.g/1,) , and approximately 95
percent above the SMAV for the mysid, Americamysis bahia  (41.29 ^g/L,
geometric mean of two flow-through measured  tests}.  The resultant saltwater
Criterion Maximum Concentration  (CMC) for total  cadmium is 40.28 pg/L (FAV/2
or 80.55 ^g/L/2).  If the total cadmium CMC  is converted to dissolved cadmium
using the 0.994  factor determined experimentally by EPA, the  saltwater  CMC for
dissolved cadmium is 40.03 jug/L- The resultant 40.03 ug/L  CMC for  dissolved
cadmium is below all of the saltwater SMAVs  presented  in Table 3  (Figure  3).
Chronic Toxicitv to Aquatic Animals
      Acceptable chronic toxicity tests have  been conducted on cadmium  in
freshwater with 21 species, including seven invertebrates  and 14  fishes in  16
genera.   Several related values are in Table  6.   Among the unused values in
Table 6,  a 21-day Daphnia magna test in which the test concentrations were  not
measured, Biesinger and Christensen  (1972) found a 16  percent reduction in
reproduction at 0.17 jug/L.  Bertram and Hart  (1979)  and Ingersoll and Winner
 (1982)  found chronic toxicity to Daphnia pulex at less than 1 and 10 ^g/L,
respectively.  The 200-hr LC10 of 0.7'^g/L obtained  with rainbow  trout  (Table
6)  by Chapman (1978) probably would be close  to  the  result of an  early  life-
                                      10

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stage test because of the extent to which various  life stages were
investigated.  Effects on other salmonids and  many invertebrates  have  been   ,
observed at 5 M9/L or less  (Table 6) .  These species  include decomposers
(Giesy 1978) , protozoans  (Fernandez-Leborans and Noville-Villajes 1993;
Niederlehner et al. 1985), Ceriodaphnia duJbia  (Winner 1988; Zuiderveen and
airge 1997), D. magna (Enserink et al. 1993; Winner and whitford  1987),
zooplankton  (Lawrence and Holoka 1987), crayfish (Thorp et al.  1979),
amphipods  (Borgmann et al. 1991; Phipps et al.  1995),  copepods  and  annelids
(Giesy et al. 1979), midges  (Anderson et al. 1980), and mayflies  (Spehar et
al. 1978) .
      An acceptable C. dubia seven-day static -renewal toxicity  teat was
conducted by Jop et al.  (1995) using reconstituted soft laboratory  water.  The
<24-hr old neonates were exposed to 1, 5, 10,  19 and  41 M9/L measured  cadmium
concentrations in addition to a laboratory water control at 25'C.   The NOEC
and LOEC were 10 and 19 ^g/L cadmium, respectively, with a resultant chronic
value of 1.4 Mg/L cadmium  (Table 2).
      The effects of water hardness on the toxicity of cadmium  to D. magna was
evaluated by Chapman et al.  (Manuscript) under static-renewal conditions at  a
temperature of 20 ±2"C.  As part of the experimental  design, the  total
hardness level was adjusted  to either 53, 103  or 209  mg/L (as CaCOj) in three
distinct tests.  Daphnids were individually exposed to six measured cadmium
concentrations  (exposures ranged from 0.15 to  22.1 A*g/L cadmium among  the
three tests) and a control  (0.08 A*g/L cadmium)  for 21 days.  Based  on  an
analysis of variance hypothesis testing procedure,  they reported  reproductive
(mean number of young per adult) chronic values of 0,1523, 0.2117 and  0.4371
Mg/L cadmium at hardness  levels of  S3, 103 and 209 mg/L,  respectively  (Table
2) .  These  same data were also subjected to a  regression analysis procedure,
whereby  the 20 percent reproductive  (mean number of young per adult)
inhibition concentration  (IC20) was estimated  for  each hardness level.   The
resultant  IC20 values were  0.07, 0.23 and 0.33 M9/L cadmium for the 53,  103
and 209  mg/L hardness  levels, respectively.  Overall,  the results obtained by
the two  different  procedures are similar.
      The  effect of  cadmium on  the  reproduction strategy of D.  magna was
                                       11

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investigated by Bodar et al.  (1988b).  After a 25-day exposure of  the  12 ± 12-
hr old neonates to .0  (control), 0.5, 1.0, 5.0, 10.0, 20.0 and 50 pq/L  cadmium
at 20 ± 1°C,  the authors compared the survival, number of neonates per female,
first day of reproduction and neonate size of the cadmium exposures  to the
                                                         -.„
controls.  The 25-day reproductive NOEC was 5.0 ^g/L cadmium, and  the  . - .
reproductive LOEC was 10.0 ^g/L cadmium.  However, • a more sensitive  endpoint
was the length of the 5th and 6th broods  of neonates;  where the 25-day NOEC and
LOEC were estimated to be 0.5 and 1.0 t*g/L cadmium, respectively.  The
resultant chronic value was  0.7071 ng/L cadmium  (Table 2).  .
      Borgman et al.  (1989)  also investigated the effect of cadmium  on.D.
magna reproduction.  The 21-day static-renewal test was  conducted  at 20°C
using measured exposure concentrations of 0.22  (control), 1.86,  4.10,  7.78 and
22.9 Mg/L cadmium.  Reproduction was significantly reduced at the  lowest
measured exposure concentration of  1.86 ^g/L cadmium.  Thus,  the reproductive
NOSC and LOEC were <1.86 and 1.86 Mg/k cadmium,  respectively, with a chronic
value of <1.86 Mg/L cadmium (Table  2),
      Brown et al.  (1994) exposed 270-day old  rainbow  trout  to  cadmium under
flow-through conditions for 65 weeks using borehole water with  a total
hardness of 250 mg/L  (as CaC03) .  Mean cadmium concentrations during the
exposure of adult  fish were 0.47  (control), 1.77, 3.39 and 5.48 pg/L.  After
65  weeks of exposure, the three  most mature males and  females were selected •
from each  treatment,  anesthetized and striped  of their gametes  when  possible,
with the milt and  ova combined in a bucket.  The fertilized  eggs from each
treatment  group were  then divided  into  four approximately equal-sized
subsamples and exposed  for  seven weeks  in 30-liter  aquaria under flow-through
conditions to nominal concentrations of  0 (control),  2,0, 5.0 and  8.0 ^g/L
cadmium.   Second generation fry development was  significantly affected when
 the parents were exposed 1.77 ^g/L  cadmium, but  not  when exposed to  0.47  M9/L
 cadmium.   However,  second generation embryo  survival  for all groups  was  less
 than 60 percent,  which may have influenced the fry  development  effect levels.
 A more representative endpoint was  the  ability of  the first  generation adults
 to reach sexual maturity,  with a statistically derived NOEC  and LOEC of  3.39
 ami 5.48 ng/L cadmium.   The  resultant  chronic value was  4.310 Mg/L cadmium
                                       12

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(Table 2).
      Brown et al.  (1994) also exposed two-year old brown  trout  to  cadmium
under flow-through conditions for 95 weeks using the same  borehole  water.
Mean cadmium concentrations during the exposure of adult fish were  0.27
(control),  5.13, 9.34 and 29.1 /ig/L.  After  60 weeks of exposure, the  three
most mature males and females were selected  from each  treatment,  anesthetized
and striped of their gametes, with the milt  and ova combined in  a bucket.  The
fertilized eggs from each.treatment group were then divided into four
approximately equal-sized subsamples and exposed for SO days in  30-liter
aquaria under flow-through conditions to cadmium concentrations  similar  to
those in which the parents were  exposed.  After the 90 week exposure,  the
survival NOEC and LOEC were 9.34 and 29.1 A*g/L cadmium, respectively,  with a
resultant chronic value  of 16.49 pg/l cadmium (Table 2).
      A  32-day fathead minnow early life stage toxicity test was conducted by
Spehar and Fiandt  (1986) under flow-through  conditions using sand filtered
Lake Superior dilution water  (Table 2).  They reported a chronic value of 10.0
,ug/L cadmium, which when coupled with their  96-hour LC50 of 13.2 ^g/L  cadmium,
gives an acute-chronic ratio of  1.320.
      Cope et al.  (1994) examined  the sublethal responses  of juvenile
bluegills exposed to cadmium under flow-through conditions at  an average total
hardness of  134 mg/L  (as CaCO3)  and temperature of 21.7"C.   The  fish were
exposed  to a control  (0.02 Mg/L  cadmium)  and seven measured cadmium
concentrations  that ranged from  2.8  to  32.3  f^g/L.  At  the  end  of the 23-day
test, test fish survival or  growth was  not adversely  affected,  resulting in a
NOSC of  >32.3 Mg/L cadmium and a chronic value  of  >32.3 t^g/l> cadmium (Table
2) .
       Ingersoll and Kemble  (unpublished)  investigated the  chronic toxicity of
cadmium to the  amphipod Hyaleila azteca.   The organisms  were exposed under
 flow-through measured conditions at a mean temperature of  23°C and a total
 hardness of 280 mg/L (as CaCO,) , and a  3-m nylon mesh  substrate  was provided
 during the test.   The seven- to eight-day old amphipods were exposed to water
 only mean total cadmium concentrations of 0.10 (control),  0.12,  0.31,  0.51,
 2.0 and 3.5 t^g/L for 42 days.   The most sensitive endpoint was survival, with
                                       13

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an NOEC and LOEC of 0.5 and 2.0 ng/lt cadmium, respectively, after both 28 and
42 days of exposure.  The resultant chronic value was  1.000 ,ug/L total cadmium
(Table 2) .
      Ingersoll and Kemble  (unpublished) also exposed  the midge Chironamus
tentans to cadmium under the same conditions listed  above for  the amphipod,
except that a thin 5 ml layer of sand was provided as  a  substrate.  The <24-hr
old larvae were exposed to water only mean measured  total cadmium
concentrations of 0.15  (control), O.SO, 1.5, 3.1, 5.8  and IS. 4 ^g/L for 20
days.  The mean weight, biomass, percent emergence and percent hatch endpoints
all had 20-day NOEC and LOEC values of 5 . 8 and  16.4  £tg/L cadmium, respectively
(Table 2).  The resultant chronic value was 9.753 ^g/L total cadmium.
      Chronic values are available over a wide  range of  hardness for two
species (Table 2) .  To account for the apparent relationship of cadmium
chronic toxicity to hardness, an analysis of covariance  (same  as the analysis
performed on the acute data) was performed to calculate  the pooled slope for
hardness using the natural  logarithm of the chronic  value as the dependent
variable, species as the treatment or grouping  variable, and the natural
logarithm of hardness as the cpvariate or independent  variable.  This analysis
of covariance model was fit to the data in Table 2 for the two species for
which definitive chronic values are available over a range of  hardness such
that the highest hardness is at least three times the  lowest,  and the highest
is also at least  100 mg/L higher than the lowest (other  species in Table 2
either did not meet these criteria or did not show any hardness -toxicity trend
probably due to differences in exposure methods, species age,  etc.).  The
slopes for the two species  ranged from 0.9786 to 1.003,  and the pooled slope
for  these two species was 0.9917  (Table 2b) .  A plot of  the chronic effect
level versus total hardness is provided in Figure 4.
      The slope of  0.9917 was used to adjust each chronic value to a hardness
of  50 mg/L.  Generally, , replicate adjusted chronic values for  a species agreed
well,  as did values  for species within a genus. The two values for Atlantic
salmon are very different,  but one agrees well  with  the  value  for the other
 tested species  in the  same  genus.  Twenty-one Species  Mean Chronic Values^ were
 then calculated,  and from  these,  the sixteen Genus Mean  Chronic Values were
                                       14

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calculated and ranked  (Table  3b).
      A freshwater Final Chronic Value was  calculated from the sixteen Genus
Mean Chronic Values using  the procedure used  to calculate a Final Acute Value.
This approach seemed appropriate since a number of  chronic tests have been
conducted with a large variety of species.  Thus,' the freshwater Final Chronic
Value for total cadmium is 0.0861 M9/L at a hardness of SO mg/L, and the Final
Chronic Value  (in M9/D =  e(0-9917Cln(hardness) ]-6-332) .  For dissolved cadmium,
the Final Chronic value is 0.0809 M9/L  (0.94  x 0.0861 Mg/L)  at a hardness of
50 mg/L, or = 0.94  te«J-99"fln«"]-«-»2>]  .  At a hardness of  50 mg/L,  all
Genus Mean Chronic Values  are above the dissolved Final Chronic Value (Figure
5) .
      Another option for calculating the Final Chronic Value is to use the
Final Acute-Chronic Ratio  in  conjunction with the Final Acute Value.   However,
the acute-chronic ratios ranged  from 0.9021 for the chinook salmon to 433.8
for the flagfish  (greater  than a factor of  ten),  with other values scattered
throughout this range  (Tables 2c and 3).  These ratios do not seem to follow
any of the patterns  (Table 3) recommended in  the guidelines,  and so it does
not seem reasonable to use a  freshwater Final Acute-Chronic Ratio to calculate
a Final Chronic Value.
      Three chronic toxicity  tests have been  conducted with the saltwater
invertebrate, Americamysis Jbahia, formerly  classified as Mysidopsis bahia
(Table 2).  Nimmo et al.  (1977a) conducted  a  23-day life-cycle test at 20 to
28"C and salinity of 15 to 23 g/kg.  Survival was 10 percent  at 10.6  Mg/L,  84
percent at the next lower  test concentration  of 6.4 Mg/L, and 95 percent in
the controls.  No unacceptable effects were observed at 6.4 Mg/L or any lower
concentration.  The chronic toxicity limits,  therefore, are 6.4 and 10.6 Mg/L,
with a chronic value of  8.237 M9/L.  The  96-hr LC50 was 15.5 Mg/L,  resulting
in an acute-chronic ratio  of  1.882.
      Another  life-cycle  test was conducted on cadmium with Americamysis bahia
under different environmental conditions,  including a constant temperature of
21°C and salinity of 30 g/kg  (Gentile et al.  1982;  Lussier et al.  Manuscript).
All organisms  died  in  28  days at 23 M9/L.   At 10 Mg/L a series of
                                       15

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morphological abberations occurred at the onset of sexual maturity.  External
genitalia in males were abberant, females failed to develop brood pouches, and
both sexes developed a carapace malformation that prohibited molting after the
release of the initial brood.  Although initial reproduction at this
concentration was successful, successive broods could not be born because
molting resulted in death.  No malformations or effects on initial or
successive reproductive processes were noted in the controls or at 5.1
Thus, the chronic limits for this study are 5.1 and 10 ,ug/L for a chronic
value of 7.141 Mg/L-  The LC50 at 21*C and salinity of 30 g/kg was 110
which results in an acute-chronic ratio of  15.40  from  this study.
      These two studies showed excellent agreement between the chronic values
but considerable divergence between the acute values and acute-chronic ratios.
Several studies have demonstrated an increase in  acute toxicity of  cadmium
with decreasing salinity and increasing temperature  (Table S) .  The observed
differences in acute toxicity to the mysids might be explained on this basis.
Nimmo et al.  (1977a) conducted their acute  test at 20.  to 28°C and salinity of
15 to 23 g/kg, whereas the other test was performed at 21°C and salinity of  30
g/kg.
      A third Americamysis bahia chronic study was conducted  by Carr  et al.
 (1985) at  a salinity of 30 g/kg, but the temperature varied from 14 to 26°C
over the 33 day study.  At test termination,  >50  percent of the organisms  had
died. in cadmium exposures *8 ptg/L.  After 18 days of exposure, growth in the 4
pg/L treatment group, the lowest exposure concentration was significantly
reduced when  compared to the controls.  The resultant  chronic limits  for this
study  are  <4  and 4  pig/L cadmium.  Acute data were not  presented by  the
authors.   The lower chronic value observed  for this study as  compared to the
 two  studies described above may have been due to  unexpected temperature
 fluctuations  over  the study period  (mechanical problems) .
       Gentile, et  al .  (1982) also conducted a life-cycle test with  another
 mysid,  Mysidopsis  bigelowi, and the  results were  very  similar to those for A.
 bahia.   Thus, the  chronic value was  7.141 ng/L and  the acute-chronic  ratio was
 15.40.
       Because they covered  such a wide  range,  it  would be inappropriate to use
                                       16

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any of the available freshwater acute-chronic ratios  in the calculation of the
saltwater Final Chronic Value.  The two saltwater species for which acute-
chronic ratios are available  (Table 3} have Species Mean Acute Values in the
same range as the saltwater Final Acute Value, and so it seems reasonable to
use the geometric mean of these two ratios.  When the saltwater Final Acute
Value of 80.55 M9/L is divided by the mean acute-chronic ratio of 9.106, a
saltwater Final Chronic Value of 8.846 ;ig/L is obtained, or 8.793
dissolved cadmium {0.994 x 8.846 /zg/L) .
Toxicitv to Aquatic Plants
      Thirty-three acceptable tests are available  with  freshwater plant
species exposed to cadmium which lasted from  4  to  28  days  (Table 4).  Growth
reduction was the major toxic effect observed with freshwater aquatic plants,
and several values are in the range of concentrations causing chronic effects
on animals.  The influence that plant growth  media might have had on the
toxicity tests in unknown, but is probably minor at least  in the case of
Conway  (1978) who used a medium patterned after natural Lake Michigan water.
Because the lowest toxicity values for fish and invertebrate species are  lower
than  the lowest values for plants, water quality criteria  which protect
freshwater animals should also protect freshwater  plants.
      Toxicity values are available for five  species of saltwater diatoms and
two species of macroalgae  (Table 4).  Concentrations causing fifty  percent
reductions in the growth rates of diatoms  range from 60 ng/L for Oitylum
brightvelli to  22,390 Mg/L  for Phaeodaccylum cricornutum,  the most  resistant
to cadmium.  The brown macroalga  (kelp) exhibited  mid-range sensitivity to
cadmium, with an ECSO of  860  pg/1*.  The most sensitive  saltwater plant tested
was the red alga,  Champia parvula,  with  significant reductions  in the growth
of both the  tetrasporophyte plant  and female plant occurring at 22.8 Mg/L.
This  plant  is more  resistant  than  the chronically most  sensitive animal
 species tested.   Therefore,  water  quality criteria for cadmium that protect
 saltwater animals  should also protect saltwater plants.
                                       17

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Bioaccumulation
      Bioconcentration  factors  (BCFs)  for cadmium  in  fresh water (Table 5)
range from 3 for brook  trout muscle  (Benoit et al.  1976)  to 6,910 for the soft
tissue of the snail  Viviparus georgianus  (Tessier  et  al.  1994b).   Usually,
fish accumulate only small amounts of  cadmium in .muscle as compared to most
other tissues and organs  (Benoit et  al. 1976; Sangalang and Freemen 1979).
Also, cadmium residues  in fish  reach steady-state  only after exposure periods
greatly exceeding 28 days {Benoit et al. 1976; Sangalang  and Freeman 1979}.
Daphnia magna, and presumably other  invertebrates  of  about this  size or
smaller, often reach steady-state within a few days  (Poldoski 1979).   Cadmium
accumulated by fish  from water  is eliminated slowly  (Benoit et al.  1976:
Kumada et al. 1980), but Kumada, et  al. (1980) found  that cadmium accumulated
from food is eliminated much more rapidly.  If all variables,  except
temperature, were kept  the same, Tessier et al. (I994a) found that  increased
exposure temperatures generally increased the soft tissue bioconcentration
factor observed for  the snail,  Viviparus -georgianus,  but  not for  the mussel,
Elliptic complanata.  Poldoski  (1979)  reported that humic acid decreased  the
uptake of cadmium by Daphnia magna,  but winner (1984)  did not find  any effect.
Ramamoorthy and Blumhagen (1984) reported that fulvic  and humic acids
increased uptake of  cadmium by  rainbow trout.
      The only BCF reported for a saltwater fish is a  value of 48 from a 21-
day exposure of the mummichog (Table 6).  However,  among  ten species  of
invertebrates, the BCFs range from 22  to 3,160 for whole  body and from S to
2,040 for muscle (Table 5).   The highest BCF was reported for the polychaete,
Ophryotrocha diad&na (Klockner  1979).  Although a  BCF  of  3,160 was  attained
after sixty-four days exposure  using the renewal technique,  tissue  residues
had not reached steady-state.
      BCFs for four  species of  saltwater bivalve molluscs  range from  113 for
the blue mussel (George and Coombs 1977) to 2,150  for  the  eastern oyster
(Zaroogian and Cheer 1976).   In addition,  the range of reported BCFs  is rather
large for some individual species.   BCFs for the oyster include 149 and 677
(Table 6) as well as 1,220,  1,830 and  2,150 (Table 5).  Similarly,  two studies
with the bay scallop resulted in BCFs  of 168 (Eisler et al.  1972) and  2,040
                                       18

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(Pesch and Stewart 1980) and three studies with the blue mussel reported BCFs
of 113, 306, and 710  (Tables 5 and 6).  George and Coombs  (1977)  studied the
importance of metal speciation on cadmium accumulation in  the  soft  tissues of
Mytilus edulzs.  Cadmium complexed as Cd-EDTA, Cd-alginate, Cd-humate,  and Cd-
pectate (Table 6) was bioconcentrated at twice the rate of inorganic cadmium
(Table 5).  Because bivalve molluscs usually do not reach  steady-state,
comparisons between species may be difficult and the length of exposure may be
the major determinant in the size of the BCF.
      BCFs  for five species of saltwater crustaceans range from 22  to  307 for
whole body  and from 5 to 25 for muscle  (Tables 5 and fi).   Nimmo et  al.  (1977b)
reported whole-body BCFs of 203 and  307 for two species of grass  shrimp,
Palaemonetes pugio and  P.  vulgaris.  Vernberg et al.  (1977) reported a factor
of 140 for  P. pugio at  2S°C (Table 6), whereas Pesch and Stewart  (1980)
reported a  BCF of 22  for the -game  species exposed at 10°C, indicating  that
temperature might be  an important  variable.  The commercially  important
crustaceans, the pink shrimp and  lobster, were not  effective bioaccumulators
of cadmium  with  factors of 57  for whole body and 25  for muscle,  respectively
 (Tables 5 and  6).
       Mallard  ducks are a  native  wildlife  species whose chronic  sensitivity to
cadmium has been studied.   These  birds  can  be expected to ingest  many  of  the
freshwater  and saltwater plants  and animals listed  in  Table  4.  White  and
Finley (1978a,b) and  White et  al.  (1978)  found  significant damage at a cadmium
concentration  of 200  mg/kg in food for 90  days.  Di Giulio and Scanlon (1984)
 found significant  effects  on energy metabolism  at  450  mg/kg,  but not at 150
mg/kg.  These  are  concentrations which would cause  damage to mallard ducks.
More recent information may be available,  but these data would not have been
 identified during  the literature search conducted for this update.
       The bioaccumulation data provided in this document is  for information
 purposes only.  Calculation of a Final Residue Value for cadmium will  not be
 presented at this time.

 Other Data
       A number of the values in Table 6 have already been discussed,   when
                                        19

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possible, the freshwater acute effect concentration  has  been  adjusted  to a
hardness of 50 using the pooled slope.  Cadmium-binding  proteins were  isolated
from Amoeba proteus  (Al-atia, 197B, 1980) and  rainbow trout  (Roberts et al.
1979) .   The cumulative mortality resulting from  exposure to cadmium for more
than 96 hours is clearly evident from the studies  with phytoplankton  (Findlay
                 •                                .•'*?•
et al.  1996; Fargasova 1993), duckweed  (Outridge 1992),  protozoa  (Niederlehner
1985),  zooplankton  (Lawrence and Holoka  (1987) ,  snails (Spehar  et al.  1978),
zebra mussels (Kraak et al. 1992), crayfish  (Thorp et al.  1979),
macroinvertebrates  (Giesy et al . 1979), polychaetes  (Reish et al. 1976),
bivalve molluscs, crabs, and starfish  (Eisler  and  Hennekey 1977), scallops,
shrimp, and crabs  (Pesch and Stewart 1980),  and  a  mysid  (Gentile et al. 1982;
Nimmo et al. 1977a) .
      Nimmo et al.  (1977a) in studies with the mysid,  Americamysis bahia,
reported a 96-hr LC50 of 15.5 ng/L (Table 1) and a 17-day LC50  of 11 ^g/L
(Table 6) at 25 to 28°C and salinity of 15 to 23 g/kg.   In another series of
studies with this mysid  (Gentile et al . 1982), the 96-hr LC50 was 110  f^g/'L
(Table 1) and the 16-day LC50 was 2B ^g/L  (Table 6)  at 20  °C  and salinity of
30g/kg.  These data suggest that short-term  acute  toxicity might be strongly
influenced by environmental variables, whereas long-term effects, even
mortality, are not.
      Considerable information exists concerning the effect of  salinity and
temperature on the acute toxicity of cadmium.  Unfortunately, the conditions
and durations of exposure are so different that  adjustment of acute toxicity
data for salinity is not possible.  Rosenberg  and  Costlow (1976) studied the
synergistic effects of cadmium and salinity  combined with constant and cycling
temperatures on the larval development of two  estuarine  crab  species.  They
reported reduction in survival and significant delay in  development of the
blue crab with decreasing salinity.  Cadmium was three times  as toxic  at a
salinity of 10 g/kg than at 30 g/kg.  Studies  with the mud crab resulted in a
similar cadmium-salinity response. ' In addition, the authors  report that
cycling temperature may have a stimulating effect  on survival of larvae
compared to constant temperature .
       Theede et al.  (1979) investigated the  effect of temperature and  salinity
                                      20

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on the acute toxicity of cadmium to the colonial  hydroid,  Laomedea loveni.  At
17.5 °C cadmium concentrations inducing irreversible  retraction  of half of the
polyps ranged from  12.4 pg/l> at a salinity of  25  g/kg to 3.0 /^g/L at 10 g/kg
(Table 6) .  At a temperature of 17.5"C, the toxicity  of  cadmium  increased as
salinity decreased  from 25 g/kg to 10 g/kg..
      A similar acute toxicity-salinity relationship  was observed by Hall et
al.  {1995} for the  copepod, Eurytemora affinig, whereby  the 96-hour toxicity
increased four-fold (from 213 to 51.6 ^g/L cadmium) when the salinity was
decreased from 15 to 5 g/kg at a test temperature of  25°C.   Hall et al. (1995)
also observed an approximate three-fold toxicity  increase to the sheepshead
minnow when the salinity was lowered in similar fashion  at the same
temperature. - Likewise, the 21-day toxicity of cadmium to the blue crab,
Calllneetes sapidus, increased over nine-fold  when the salinity was lowered
from 25 to 2.5 g/kg, and the temperature was held constant at 22-23 °C (Guerin
and Stickle 1995).   In contrast, Snell and Personne (1989b)  observed little
difference in the 24-hour toxicity of cadmium  to  the  rotifer, Brachionus
plica tills/ exposed under 15 and 30 g/kg salinity regimes and a temperature of
25 °C.
      The effect of environmental factors on the  acute toxicity of cadmium is
also evident from tests with the early life stages of saltwater vertebrates.
Alderdice, et al.  (1979a,b,c) reported that salinity  influenced the effects of
cadmium on the volume, capsule strength, and osmotic  response of embryos of
the  Pacific herring.  Studies with embryos of  the winter flounder indicated a
quadratic salinity-cadmium relationship  (Voyer et al. 1977), whereas Voyer et
al.  (1979) reported a linear relationship between salinity and cadmium
toxicity  to Atlantic silverside embryos.
      Several studies have reported chronic sublethal effects of cadmium on
saltwater fishes  (Table  6).  Significant reduction in gill tissue respiratory
rate was  reported for the cunner after a 30-day exposure to 50 ng/L (Maclnnes
et al. 1977).  Dawson et al.  (1977) also reported a significant  decrease in
gill-tissue  respiration of  striped bass at  0.5 ^g/L above ambient levels after
a 30-day, but not a 90-day,  exposure.  A similar  study with the winter
 flounder (Calabrese et  al.  1975) demonstrated  a  significant alteration in gill

                                       21

-------
tissue respiration rate measured  in  vitro after a  60-day exposure to 5
Unused Data
      Some data on the effects of  cadmium on aquatic organisms were not used
because the studies were conducted with species  that are not resident in North
                                                   ^ *
America, e.g., Abbasi and Soni  (1986),  Abel and  Papoutsoglou (1986), Abel and
Garner  (1986), Abel'and Barlocher  (1988),  Ahsanullah et al.  (1981), Ahsanullah
and Williams  (1991),  Amiard-Triquet et  al.''(1987), Annune et al.  (1994),
Arshaduddin et al. (1989), Austen  et al.  (1997),  Avery et al.  (1996), Azeez
and Banerjee  (1987),  Baby and Menon (1987),  Bambang  et al.  (1994),  Bednarz and
Warkowska-Dratnal  (1983/1984), Birmelin et al.  (1995),  Bresler and Yanko
(1995), Brooks et al.  (1996), Brunetti  et al.  (1991),  Calevro et al. (1998),
Canli and Furness  (1993, 1995),  Cassini et al.  (1986),  Castille and Lawrence
(1981), Centeno et al.  (1993), Chan (1988),  Chandini (1988,  1988, 1989, 1991),
Chandra and Garg  (1992), Charpentier et al.  (1987),  Chattopadhyay et al.
(1995), Cheung and Lam  (1998), Coppellotti (1994), D'Agostino and Finney
(1974), Dallinger et al  (1989),  Darmono (1990),  Dartnono et al. (1990),  Datta
et al.  (1987), Demon et al  (1989),  Den  Beaten et al. (1989,  1991),  De Nicola
Giudici and Guarino  (1989),  De Nicola Giudici and Migliore (1988),  Denton and
Burdon-Jones  (1986, 1986), Devi  (1987,  1996),  Devi and Rao (1989),  Devineau
and Triquet  (1985), Dorgelo  et al.  (1995), Douben (1989), Drbal et al.   (1985),
Duquesne and  Coll  (1995),  Evtushenko et al.  (1986),  Evtushenko et al.  (1990),
Ferrari et al.  (1993),  Fisher et al. (1996), Fisher  et al.  (1996),  Forget et
al.  (1998),  Francesconi  (1989),  Francesconi et al. (1994),  Forbes  (1991), Gaur
et al.  (1994), Gerhardt  (1992,  1995), Ghosh and  Chakrabarti (1990), Glynn
 (1996), Glynn et  al.  (1992,  1994),  Gopal and Devi (1991), Green et al.   (1986),
Greenwood and Fielder'(1983), Gupta and Rajbanshi (1991), Gupta et al.   (1992),
Hader et al.  '(1997), Hansten et  al.  (1996),  Heinis et al. (1990), Herkovits
and.Coll  (1993),  Hiraoka et  al.  (1985), Hu et al. (1996), Huebner and Pynnonen
 (1992), Husaini  et al.  (1991),  Ikuta (1987), Jenkins and Sanders (1985),
 Karlsson-Norrgren and  Runn (1985), Kasuga (1980), Keduo et al. (1987),
 Khangarot  and Ray.  (1987),  Khristoforova et al.  (1984), Kobayashi  (1971),
 Krassoi and Julli (1994),  Krishnaja et  al.  (1987), Kuhn and Pattard  (1990),
                                       22

-------
Kuroshima  (1987), Kuroshima and Kimura  (1990),  Kuroshima et al.  (1993),  Lam
(1996, 1996), Lam et al.  (1997), Lee and Xu  (1984),  Loumbourdis  et al (1999),
McCahon et al.  (1988), McCahon and Pascoe  (1988,  1988,  1988),  McCahon et al.
(1989), McClurg  (1984), Ma et al.  (1999),  Malea (1994),  Markich  and Jeffree
(1994, 1994), Martinez et al.  (1996), Metayer et  al.  (1982),  Michibata et al.
(1986), Michibata et al.  (1987), Migliore  and Giudici (1987),  Moller et al.
(1994), Mostafa and Khalil (1986), Muino et  al.  (1990),  Musko et al. (19903,
Nakagawa and Ishio  (1988, 1989, 1989), Nasairi et al. (1997),  Negilski (1976),
Nir et al. (1990), Noraho and Gaur  (1995), Notenboom et  al. (1992),  Nott and
Nicolaidou (1994), Nugegoda and Rainbow  (1995), Ojaveer  et al. (1980),  Pantani
et al. (1997), Papathanassiou  (1995), Pavicic et  al.  (1994),  Perez-Coll and
Herkovits  (1996), Pynnonen (1995), Rainbow and Kwan   (1995), Rainbow et al.
(1980), Rainbow and White (1989), Ralph and  Burchett (1998),  Ramachandran et
al. (1997), Rao and Madhyastha  (1987), Rebhun and Ben-Amotz (1984),  Reish et
al. (1988), Ringwood  (1990, 1992), Ritterhoff et  al.  (1996),  Romeo and
Gnassia-Barelli  (1995), Safadi  (1998), Sastry and Shukla (1994),  Sastry and
Sunita (1982), Saxena et  al.  (1990,  1993), Schafer et al.  (1994),  Sehgal and
Saxena (1987), Shanmukhappa and Neelakantan  (1990),  Shivaraj  and Patil (1988),
Simoes Goncalves  (1989),  Stuhlbacher and Maltby (1992),  Takaittura et al.
(1989), Temara et al.  (1996a,b), Ten Hoopen  et al. (1985),  Thaker and Haritos
(1989), Thebault et al.  (1996), Theede et  al. (1979), Tomasik et al. (1995),
Tyurin and Khristoforova  (1993), Udoidiong and Akpan (1991),  Valencia et al.
(1998), Van Gemert  (1985), Vashchenko and  Zhadan  (193),  Verriopoulos and
Moraitou-Apostolopoulou  (1981,  1982), Visviki and Rachlin  (1991),  Vogiatzis
and Loumbourdis  (1998), Vranken et al.  (1985), Vuori (1994),  Vymazal (1990,
1995), Walsh et  al.  (1995), Warnau et al.  (1995a,b,c, 1996a,b, 1997),
Westernhagen and Dethlefsen  (1975),  Westernhagen  et  al.   (1975, 1978),  Wildgust
and Jones  (1998), White  and Rainbow  (1986),  Wicklund and Runn (1988),  wicklund
et al. (1988),  Hu et  al.  (1997), Wundram et  al. (1996),  Zanders  arid Rojas
(1992, 1996),and Zou  and Bu  (1994).  Brown and Ahsanullah  (1971)  conducted
tests with a  brine  shrimp, which  species  are too  atypical to be  used in
deriving  national criteria.
       Data were also  not used if  cadmium was a component of a drilling mud,

                                       23

-------
effluent, mixture, sediment, or sludge  (Allen  1994,  1995;  Araiard-Triquet  et
al. 1988; Andres et al. 1999; Arnac and Lassus  1985;  Austen and McEvoy 1997;
Bartsch et al. 1999; Beiras et al. 1998; Bendell-Young 1994;  Bendell-Young et
al. 1986; Besser and Rabeni 1987; Biesinger et  al.,-1986;  Bigelow and Lasenby
                                                 j
1391; Bodar et al. 1990; Buckley et al. 1985;  Burden and  Bird-1994;  Busch et
al. 1998; Campbell and Evans 1991; Camusso et  al.  199S; Carlisle and Clements
1999; Casini  and Depledge  1997; Cuvin-Aralar 1994;  Cuvin-Aralar and Aralar
1993; Dallinger et al. 1997; de March  1988,-'  Elliott et al. 1986; Farag et al.
1994, 1998; Gully and Mason 1993; Hall et al.  1984,  1987,  1988; Hardy and
Raber 1985; Hare et al. 1991, 1994; Haritonidis et  al. 1994;  Hartwell 1997;
Haynes et al. 1989; Hendriks 1995; Hickey and  Clements 1998;  Hickey and Martin
1995;' Hickey  and Roper 1992; Hogstrand et al.  1991;  Hollis et al. 1996; Hooten
and Carr 1998; Hylland et  al. 1996; Inza et al. 1998; Jak et al. 1996;
Janssens de Bisthoven et al. 1992; Jop 1991; Keenan and Alikhan 1991; Kelly
and Whittoh 1989; Kettle and deNoyelles 1986;  Khan and Weis 1993; Khan et al.
1989; Kiffney and Clements 1996; Klerks and Bartholomew 1991; Kock et al.
1995; Koivisto et al. 1997; Kolok et al. 1998;   Kraak et  al.  1993, 1994;
Krantzberg 1989a,b; Krantzberg and Stokes 1988, 1989; Kumar 1991; Lee and
Luoma 1998; Lithner et al. 199S; Lucker et al.  1997; Macdonald and Sprague
1988; Maloney 1996; Manz et al.  1994;  Marr et  al.  1995a,b; Mathew and Menon
1992; Mersch  et al. 1996;  Nalewajko 1995; Nelson 1994; Odin et al. 1996,  1997;
Palawski et:-al 1985;  Pedersen and Petersen 1996; Pellegrini et al. 1993;
Playle et al. 1993; Polar,  and Kucukcezzar 1986; Poulton et al. 1995; Prevot
and  Soyer-Gobi Hard  1986;  Qichen et al. 1988;  Rachlin~and Grosso 1993;
Reynoldson et al.  1996;  Richelle et al. 1995;  Roch and McCarter 1984;
Roesijadi and Fellingham 1987;  Sanchiz et al.  1999; Schaeffer et al. 1991;
Smokorowski  et  al.  1997;   stephenson and Macki 1989; Stern and Stern 1980;
Talbot  1985,  1987;  Tessier et al 1993; Vuori  1993; Vymazal 1984;  Wall et al.
 1996;  Walsh  and  Hunter 1992;  Wang  et al.  1996; Warren et al. 1998; Weimin et
 al.  1994; Wong  et al.  1982; Woodling  1993; Woodward et al. 1995).  Reviews by
 Barnthouse et al.  (1987),  Bay et al.  (1993),  Cairns et al. (1985), Chapman et
 al.  (1968),  Dierickx and Bredael-Rozen (1996), Dyer et al. (1997), Eisler
 (1981),  Sisler et al. (1979),  Enserink et al.   (1991), Florence et al.  (1992),
                                       24

-------
Guilhermino et al.  (1997), Hare  (1992), Hornstrom (1990),  Jonnalagadda and Rao
(1993),  Khangarot and Ray  (1987),  Kooijman  and  Bedaux (1996),  Kraak et al.
(1994a,b), LeBlanc  (1984), Mark  and Solbe  (1998),  Meyer (1999),  Nendza et al.
(1997),  Oikari et al.  (1992),  Papoutsoglou  and  Abel  (1993),  Pesonen and
Andersson  (1997), Phillips and Russo  (1978),  Ramesha et al.  (1996),  Rice
(1984),  Skowronski et al.  (1998),  Spry and  Wiener (1991),  Thomann et al.
(1997),  Thompson et al.  (1972),  Toussaint et  al.  (1995),  Trevors et al.
(1986),  Van Leeuwen et al.  (1987), Vymazal  (1990), Wright and Welboum'  (1994),
and Wong  (1937). only contain data  that have been published elsewhere.
      Data were not used if the  organisms were  exposed to cadmium in food or
by injection or gavage  (e.g.,  Bodar et al.  1988;  Brouwer et al.  1992;  Chou et
al. 1986; Davies et al.  1997;.  Decho and Luoma 1994;  Gottofrey and Tjalve 1991;
Handy 1993; Kluttgen and Ratte 1994;  Kuroshima  1992;  Lasenby and Van Duyn
1992; Lawrence and Holoka 1991;  Lomagin and Ul'yanova 1993;  Malley and Chang
1991; Melgar et al. 1997; Mount  et al. 1994;  Munger  and Hare 1997; Postma et
al. 1994; Postma and Davids 1995;  Reinfelder  and Fisher 1994,  1994;  Reddy et
al. 1997; Rhodes et al.  1985;  Van  den Hurk  et al.  1998; Wallace and Lopez
1997; Wang and Fisher  1996; Wen-Xiong and Fisher 1996; Wong 1989).
      A number of studies of cadmium  toxicity examined physiological or
behavioral effects but provided  no interpretable concentration,  time,  response
data, and some papers  described  effects of  only a single,  often lethal,
concentration.  Included in-such studies are  those of Berglind (1985),  Bitton
et al.  (1994), Block and Part  (1992), Block et  al.  (1991), Blondin.et al.
(1989),  Bowen and Engel. (1996),  Bressan and Brunetti (1988), Castano.et al.
(1996),  Christoffers and Ernst (1983), Clausen  et al. (1993),  Fargasova
(1994),  Fernandez-Pinas  et.al. (1995), George et al.  (1983), Iftode et al.
(1985), Ilangovan et al. (1998), Issa et  al.  (1995),  Jana arid Sahana (1988),
Kluytmans  et al.  (1988), Kraak et  al. (1993b),  Kosakowska et al. (1988),
Lussier et al.  (1999),  Mateo  et  al.   (1993), Palackova et al. (1994), Pereira
et al.  (1993), Prasad  et al.   (1998),  Rachlin  and Grosso (1991),  Reader et al.
(1989), Reddy and Fingerman (1994),  Reid  and  McDonald (1991),  Ribo (1997),
Rombough  (1985),  Rosas and Ramirez (1993),  Sauvant et al. (1997), Skowronski
et al.  (1991), Sunila  and  Lindstrom  (1985), Trehan and Maneesha  (1994),

                                       25

-------
 Verbost et al.  (1987), Visviki and Rachlin  (1994), Wang et al.  (1995), Woodall
 et al.  (1988),  Wundram et al.  (1996), and Xue and Sigg  (1998).
       Battaglini et al.  (1993),  Borchardt  (1983), Craig et al.  (1998),
•Gargiulo et al.  (1996), Gomot (1998), Harvey and Luoma  (1985),  Kraal  et  al.
 (1995), Penttinen et al. (1995), Rouleau et al.  (1998), and  Sobhan  and
 Sternberg (1999)  presented no useable data on cadmium toxicity  or
 bioconcentration.
       Papers that dealt with the selection, adaptation, or acclimation of
 organisms for increased resistance to cadmium were not used, ,e.g.,  Anadu et
 al. (1989), Bodar et al. (1990), Currie et  al.  (1998), Ramo  et  al.  (1987),
 Herkovits and Perez-Coll (1995), Kaplan et  al.  (1995), McNicol  and  Scherer
 (1993), Madoni et al.  (1994), Nagel  and Voigt  (1995), Thomas et al.  (1985),
 and Van Steveninck et  al.  (1992).
       Data were not used if the results were only presented  graphically
 (Laegreild et al. 1983; Laube 1980;  Remade et  al.  1982),  if the organisms
 were not .exposed to cadmium in water (Foster 1982;  Hatakeyama and Yasuno
 1981a; O'Neill 1981),  or if there was no pertinent  adverse effect  (Carr  and
 Neff 1982; DeFilippis  et al.  1981; Dickson et  al.  1982;  Fisher and  Fabris
 1982,-  Fisher and Jones 1981,-  Tucker  and Matte  1980;  Watling  1981; Weis et al.
 1981).  Data in publications  such as Abbasi and Soni (1989), Ball  (1967),
 Belabed et al.  (1994),  Bendell-Young (1999), Bitton et  al.  (1995),  Bjerregaard
 and Depledge (1994),  Bolanos  et  al.  (1992), Burnison et al.  (1975), Calevro et
 al.  (1998),  Canton  and Slooff (1979), Carpene  and Boni  (1992),  D'Aniello et
 al.  (1990),  Davies  et al.  (1994), Department of the Environment (1973),
 Srrecalde et al.  (1998), Fennikoh et al.  (1978),  Fernandez-Leborans and
 Antonio-Garcia  (1988), Galic  and Sipos  (1987),  Glubokov (1990), Gorman and
 Skogerboe (1987),  Guanzon  et  al.  (1994),  Guerin et al.  (1994),  Hofslagare et
 al.  (1985),  Janssen and Persoone (1993),  Jaworska et al.  (1997), Kay et  al.
  (1986),  Kessler (1985), Khangarot  et al.  (1987),  Koyama et al.   (1992),  Landner
  and Jernelov (1969),  Lee and Oshima (1998), Liao and Hsieh  (1990),  Maas
   (1978), Mansour (1993), Ministry of Technology (1967),  Moza et al.  (1995),
  Munger et al.  (1999), Naylor et al.  (1992), Nwadukwe and Erondu (1996),  Pascoe
  and Shazili (1986), Pauli and Berger (1997),  Penttinen et al.  (1998), Peterson

                                        26

-------
(1991), Peterson et  al.  (1984),  Rayms-Keller  et  al.  (1998),  Rombough (1985),
Sandau et al.  (1996), Sekkat  et  al.  (1992), Shcherban (1977),  Sheela et al.
(1995), Sovenyi and  Szakolczai  (1993), Stom and  Zubareva (1994),  Stubblefield
et al. (1999), Tarzwell  and Henderson  (I960),  Verma  et al.  (1980),  Vykusova
and Svobodova  (1987), Wani  (1986), Witeska et al.  (199S),  Yamamoto and Inque
(1985), and Zhang et al.  (1992)  were not used because either the  materials,
methods,  or results  were insufficiently described.   High control  mortalities
occurred in testing  reported  by  Asato and Reish  (1988),  Hong and  Reish (1987),
Sauter et al.  (1976) and Wright  (1988).  The  96-hr values reported by Buikema
et al. (1974a,b) were subject to error because of  possible reproductive
interactions  (Buikema et al.  1977).  Bringmann and Kuhn (1982)  and Dave et al.
(1981) cultured daphnids in one  water and tested them in a different water.
      The acceptability  of the dilution water or medium used in some studies
(e.g., Brkovic-Popovic and Popovic 1977a,b;.Cearley  and Coleman 1973,  1974;
Nasu et al. 1983) was open to question because of  its origin or content.
Algal studies were not used if they were not  conducted in an appropriate
medium (Stary and Kratzer 1982;  Stary et al.  1983) or if the medium contained
too much of a complexing agent such as EDTA  (Baillieul and Blust  1999;  Brand
et al. 1986; Chen et al.  1997; Couillard 1989; Hockett and Mount  1996;  Huebert
et al. 1993;  Huebert and Shay 1991, 1992, 1993; Jenkins and Mason 1988;
Jenkins and Sanders  1986; Jenner and Janssen-Mommen  1993;  Kessler 1986;  Lue-
Kim et al. 1980; Macfie  et al. 1994; Meteyer  et  al.  1988;  Muller  and Payer
1979; Nasu et al. 1988;  Rebhun and Ben-Amotz  1986, 1988; Sloof  et al.  1995;
Sunda and Huntsman 1996;  Thongra-ar and Matsuda,1993;  Thorpe and  Costlow 1989;
Tortell and Price 1996;  Vasseur  and Pandard 1988;  Wright et  al  1985).   Some
papers were omitted  because of questionable treatment of test organisms or
inappropriate test conditions or methodology  (e.g.,  Babich and  Stotsky  1982;
Brown et al. 1984; Bryan 1971; Chan et al. 1981; Dorfman 1977;  Eisler and
Gardner 1973; Greig  1979; Hung 1982; Hutcheson 1975; Moraitou-Apostolopoulou
et al. 1979; Parker  1984; Pecon  and Powell 1981; Ridlington et  al.  1981; Sunda
et al. 1978; Wikfors and Ukeles  1982).
      Data on bioconcentration by aquatic organisms-were not used if the test
was conducted  in distilled water, was not long enough,  was  not  flow-through,
                                      27

-------
or if the concentrations'in water were not adequately measured  (e.g., Allen
1995; Amiard et al. 1993; Amiard-Triquet et al. 1986; Balogh and Salanki  1984;
Baudrimont et al. 1997; Seattle and Pascoe 1978; Bentley  1991;  Berglind 1986;
Bernds 1998; Bervoets et al. 1995, 1996; Bjerregaard 1982,  1985, 1991;  Block
and Glynn 1992; Brown et al. 1986; Burrell and Weihs 1983;  Carmichael and
Fowler 1981; Carr and Neff 1982; Chan et al.  1992; Chander  et al.  1991; Chawla
et al. 1991; Chitguppa et al. 1997; Chou and  Uthe  1991; Collard and Matagne
1994; Craig et al. 1999; Davies et al. 1981;  De Conto Cinier et al.  1997;  De
Conto Cinier et al. 1998; De Nicola et al.* 1993; Denton and Burdon-Jones  1981;
                                       .'*'.;.
Elliott et al. 1985; Engel 1999; Everaarts 1990; Fair and Sick  1983; Frazier
and George 1983; Freeman 1978, 1980; Giles 1988; Gottofrey  et al.  1988; Graney
et al. 1984; Gupta and Devi 1993; Haines and  Brumbaugh 1994; Hansen et  al.
1995; Hardy and O'Keeffe 1985; Hashim et al.  1997; Hatakeyama 1987;  Herwig et
al- 1989; Hollis et al.  1997; Irato and Piccinni 1996; John et  al.  1987;  Katti
and Sathyanesan 1985; Kerfoot and Jacobs 1976; Khoshmanesh  et al.  1996, 1997;
Klaverkamp and Duncan 1987; Koelmans et al. 1996;  Kohler  and Riisgard 1982;
Kwan and Smith 1991; Langston and Zhou 1987;  Les and Walker 1984;   McLeese and
Ray 1984; Maeda et al. 1990; Malley et al. 1989; Maranhao et al. 1999;  Mersch
et al 1993; Mizutani et  al 1991; Muramoto 1980; Mwangi and  Alikhan 1993;  Nolan
and Duke 1983; Morey et  al. 1990; Oakley et al. 1983; Olesen and Weeks  1994;
Papathanassiou 1986; Pawlik and Skowronski 1994; Pawlik et  al.  1993; Pelgrom
et al 1994; Pelgrom et al. 1997; Playle and Dixon  1993; Preaing et al.  1993;
Postma et al. 1996; Poulsen et al. 1982; Rai  et al. 1995; Rainbow  1985;
Ramjrez et al. 1989; Ray et al. 1981; Reichert et  al. 1979; Reinfelder  et al.
1997; Riisgard et  al. 1987; Ringwood 1989, 1992, 1993; Roseman  et  al. 1994;
Rubinstein et al.  1983;  Santojanni et al. 1998"; Sedlacek  et al. 1989; Sidoumou
et al. 1997; Simoes Goncalves et al. 1988; Sinha et al. 1994; Skowronski  and
Przytocka-Jusiak  1986; Srivastava and Appenroth 1995; Stary et  al.  1982;  Sunil
et al. 1995; Suzuki et al.  1987; Swinehart  1990; Taylor et  al.  1988; Tessier
et al. 1996; Thomas et al.  1983; Van Leeuwen  et al. 1985; Van Ginneken  et al.
 1999; Vymazal  1995; Wang and Fisher 1998; Watling  1983a;  White  and Rainbow
 1982; Williams et  al. 1998; Windom et al. 1982; Winner and  Gauss 1986;  Winter
 1996; Woodworth  and Pascoe  1983; "Xiaorong et  al. 1997; Yager and Harry  1964;

                                      28

-------
Zauke et al. 1995; Zia and McDonald 1994) .  The bioconcentracion  testa  of
Eisler (1974), Jennings and Rainbow (I979b), O'Hara  (1973b), Phelps  (1979),
and Sick and Baptist  (1979), which used radioactive  isotopes of cadmium, were
not used because of the possibility of isotope discrimination.  Reports on the
concentrations of cadmium in wild aquatic organisms,  such as Anderson et al.
(1978), Bouquegneau and Martoja  (1982), Boyden  (1977), Bryan et al.  (1983),
Frazier (1979), Gordon et al.  (1980),  Greig and Wenzloff  (1978),  Hazen  and
Kneip  (1980), Kneip and Hazen  (1979),  McLeese et al.  (1981), Noel-Lambot et
al. (1980), Pennington et al.  (1982),  Ray et al.  (1981), Smith et al.  (1981),
and Uthe et al. (1982) were not used for.the calculation of bioaccumulation
factors due to an insufficient number of measurements of the concentration of
cadmium in the water.
Summary
      Freshwater Species Mean Acute Values  for  cadmium are available for
species in 59 genera and hardness adjusted  values  range  from 1.656 M9/L for
brown trout to 78,579 pg/L for a midge.  The  antagonistic effect of hardness
on acute toxicity has been demonstrated with  10 species.  Chronic tests have
been conducted on cadmium with 14 freshwater  fish  species and seven
invertebrate species with hardness adjusted Species  Mean Chronic Values
ranging from 0.1811 M9/L for Hyalella azteca  to 34.19  ^9/L for Ceriodaphnia
dtubia.  Acute-chronic ratios are available  for  eight species and range from
0.9021 for the chinook salmon to 433.8 for  the  flagfish.
    .  Freshwater aquatic plants are affected  by cadmium  at concentrations
ranging from 2 to 20,000 Atg/L.  These values  are in  the  same range as the
acute toxicity values for fish and invertebrate, species, and are considerably
above the chronic values.  Bioconcentration factors  (BCFs) for cadmium in
fresh water range from 7 to 6,910 for invertebrates  and  from 3 to 2,213 for
fishes.
      Saltwater cadmium species mean acute  values  for  11 fish species range
from 75.0 M9/L for  striped bass to 50,000 M9/L  for sheepshead minnow.  Species
Mean Acute values for SO species of  invertebrates  range  from 41.29 ^g/L for a
mysid to  135,000 M9/L for an oligochaete worm.   The  acute toxicity of cadmium

                                      29

-------
generally increases as  salinity  decreases.  The  effect  of  temperature seems  to
be species^specific.  Two  life-cycle  tests with  Americamysis bahia under
different test conditions  resulted  in similar chronic values of 8.237 and
7.141 fj.g/L, but the acute-chronic ratios were 1.882  and 15.40,  respectively.
A third chronic test with  Americamysis bahia gave  a  slightly lower chronic
value, possibly due to  the unexpected temperature  (14-26 °C)  fluctuation.  The
acute values appear to  reflect effects of salinity.and  temperature,  whereas
the few available chronic  values apparently do not.  A  life-cycle test with
Mysidopsis bigelowi also resulted in  a chronic value of 7.141 (tg/L and an
acute-chronic ratio of  15.40.  Studies with microalgae  and macroalgae revealed
effects at 2 to 22,390  ^g/L.
      BCFs determined with a variety  of saltwater  invertebrates range from S
to 3,160.  BCFs for bivalve molluscs  were generally  above  1,000 in long
exposures, with no indication that  steady-state  had  been reached.  Cadmium
mortality is cumulative for exposure  periods beyond  four days.   Chronic
cadmium exposure resulted  in significant effects on  the growth of bay scallops
at 78
National Criteria
      The procedures described  in the  "Guidelines  for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic  Organisms and
Their Uses"  indicate that, except possibly: where  a  locally important  species
is very sensitive, freshwater aquatic  organisms and  their uses should  not be
affected unacceptably if the four-day  average concentration (in ^g/L)  of
dissolved cadmium does not exceed the  numerical value given by 0.94
, _(0. 9917 [In(hardness) 1-6.332),       ..
le                          J more  than once every three years on the average,
and if the one-hour average dissolved  concentration  (in Mg/D  does not exceed
the numerical value given by 0.97  [e'1'205 Hn (hardness) ] -3. 949) j  more than once
every three years on the average. : For example, at hardnesses  of  50, 100, ,and
200 mg/L as CaC03 the four-day average dissolved concentrations of cadmium are
0.08, 0.16 and 0.32 ^g/L, respectively, and the one -hour average  dissolved
concentrations are 2.1, 4.8, and 11 ng/L.  If brown  trout are  as  sensitive as
some data indicate, they may not be protected by this criterion.
                                      30

-------
      The procedures described in the "Guidelines for Deriving Numerical
National Water Quality Criteria for the Protection of Aquatic Organisms and
Their Uses" indicate that, except possibly where a locally important species
is very sensitive, saltwater aquatic organisms and their uses should not be
affected unacceptably if the four-day average dissolved concentration of
cadmium does not exceed 8.8 M9/L more than once every three years on the
average and if the one-hour average dissolved concentration does not exceed 40
ug/L more than once every three years on the average.  However, the limited
data suggest that the acute toxicity of cadmium is salinity-dependent;
therefore the one-hour average concentration might be underprotective at low
salinities and overprotective at high salinities.
      EPA believes that the use of dissolved cadmium will provide a more
scientifically correct basis upon which to establish water-column criteria for
metals.  The criteria were developed on this basis.  The use of dissolved
criteria reduces the amount of conservatism that was present in earlier
cadmium criteria.  It is recognized that a considerable proportion of
dissolved cadmium in organic-rich waters may be less toxic than freely
dissolved cadmium. On the other hand, some particulate forms of cadmium may
contribute to cadmium loading of organisms, possibly through ingestion.
      The recommended exceedence frequency of three years is the Agency's best
scientific judgment of the average amount of time it will take an unstressed
system to recover from a pollution event in which exposure to cadmium exceeds^
the criterion.  Stressed  systems, for example, one in which several outfalls
occur in a limited area, would be expected to require more time for recovery.
The resilience of ecosystems and their ability to recover differ greatly,
however, and site-specific criteria may be established if adequate
justification is provided.
      The use of  criteria in designing waste treatment facilities requires the
selection of an appropriate wasteload allocation model.  Dynamic models are
preferred  for the application  of  these criteria.  Limited data or other
factors  may make  their  use impractical,  in which case one should rely on a
steady-state model.   The Agency recommends the interim use of  1Q5 or 1Q10 for
Criterion  Maximum Concentration (CMC) design flow and 7QS or  7Q10  for the

                                       31

-------
Criterion Continuous Concentration (CCC) design flow in steady-state models
for unstressed and stressed systems respectively.  These matters are discussed
in more detail in the Technical Support Document for Water Quality-Based
Toxics Control (U.S. EPA 1985).
                                      32

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